Preface – Water-Quality Engineering K Hanaki, University of Tokyo, Tokyo, Japan & 2011 Elsevier B.V. All rights reserved...
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Preface – Water-Quality Engineering K Hanaki, University of Tokyo, Tokyo, Japan & 2011 Elsevier B.V. All rights reserved.
Water technology has been ever growing. It is an essential set of technologies for sustainable human society. Traditional technology, or better called just skill, to obtain, purify, and supply water was developed in the ancient era in various regions of the world. Great efforts have been made to obtain safe and adequate water as an essential resource to human life. However, still, billions of people in the world have no access to safe water. Moreover, large numbers of people have no chance to use a proper sanitation system, and this eventually deteriorates water quality and decreases the available safe water resources. Water resources are renewable theoretically. Used water does not disappear but is renewed to freshwater through evaporation by the power of solar energy. Solar energy is a natural distillation system to remove impurities present in water. However, the help of water technology is needed to maintain this renewing function in the modern world in which human activity overwhelms the natural purifying function. Conventional water technology was used as a black box through which water was purified without knowing the mechanisms, which control the physical, chemical, and biological reactions used in purification. However, such empirical use of technology cannot further improve or develop the technology. Many researchers and practitioners have developed theory-based technology, rather than mere empirical skill, for purifying water. The function of each unit process was studied and the mechanisms of separation, role of microorganisms, and process characteristics were clarified. A significant amount of knowledge has been accumulated. This knowledge improves process performance and reliability. Human beings also developed tools to examine the micro- or nanoscale reaction. Modern technology needs to be based on a deep and broad understanding of theory. Water technology is not isolated from other technologies. Many innovations to upgrade water-technology performance have been tried by applying new technologies from other fields. Membrane technology that originated in a field such as medical science or chemical engineering is an example. Nowadays, water treatment is one of the largest application areas of membrane technology. The purpose of water technology has been expanded from purification of water to water generation, energy and resource recovery. This is a practical and important area to which new
technology can be applied. Water availability is limiting human settlements. The supply of water produced from seawater or even moisture can break through this limitation. The requirements for water technology differ very much from one place to the other. The key factors are target compounds to be removed, resource and energy consideration, capacity of operating human resources, as well as economic resources. For example, a safe water-supply system in leastdeveloped areas needs technology, which can be used without frequent and sophisticated maintenance. However, such technology does not mean cheap and old technology. Newly developed innovative technology has a higher chance of implementation than old technology. Water management needs policy and system technology rather than simple connection of unit technologies. A distributed wastewater treatment system needs reliable and economically and technologically reasonable treatment technologies. A nutrient removal policy for eutrophication can be realized by introducing a technologically reasonable combination of secondary and advanced treatments. The water technology is a system technology. Resource and energy limitation has become a key factor for sustainability. Substantial amount of material use threatens the world’s resources, and energy use provokes the climate change problem. Saving resource and energy is now an indispensable aspect of water technology. The necessity of energy and resource saving further changes water technology. The current global situation regarding climate change and resource limitation enhances the recovery of resource and energy. Wastewater contains organic matter, which is biomass; therefore, obtaining carbon-neutral energy is possible. Water technology is now forming an important part of business worldwide. Every country needs safe water and environmental protection from wastewater. Technology development, implementation, and maintenance provide substantial opportunities for business. This volume includes theory, practice, and recent development of these wide range of water technologies, although all such innovative technologies cannot be included. There is no single answer to any of the particular cases. Among many options, one should choose a technology system considering the local social, economic, and engineering aspects. This volume would help such a technology choice.
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4.01 Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations G De Feo, University of Salerno, Fisciano (SA), Italy LW Mays, Arizona State University, Tempe, AZ, USA AN Angelakis, Institute of Iraklion, Iraklion, Crete, Greece & 2011 Elsevier B.V. All rights reserved.
4.01.1 4.01.2 4.01.3 4.01.4 4.01.5 4.01.6 4.01.7 4.01.8 References
Aqueducts Minoan and Greek Aqueducts Roman Aqueducts Cisterns and Reservoirs Water Distribution Systems Fountains Drainage and Sewerage Systems and Toilets Discussion and Conclusions
Prolegomena The past is the key for the future ‘Hydor (Water) is the beginning of everything’ Thales from Miletus (c. 636–546 BC).
Humans have spent most of their existence as hunting and food-gathering beings. Only in the last c. 9000–10 000 years, they discovered how to grow agricultural crops and tame animals. Such revolution probably first took place in the hills to the north of Mesopotamia. From there the agricultural revolution spread to the Nile and Indus Valleys. During this agricultural revolution, permanent villages replaced a wandering existence. About 6000–7000 years ago, farming villages of the Near East and Middle East became cities. Hydraulic technology began during antiquity long before the great works of such investigators such as Leonardo da Vinci (1452–1519) and Isaac Newton (1642–1727), and even long before Archimedes (287–212 BC) (Mays, 2008). During the Neolithic age (c. 5700–3200 BC), the first successful efforts to control the water flow were driven (such as dams and irrigation systems) due to the food needs and were implemented in Mesopotamia and Egypt (Mays et al., 2007). Urban water-supply and sanitation systems are dated at a later stage, in the Bronze Age (c. 3200–1100 BC). Regarding the technological principles related to water and wastewater, today it is well documented that many are not achievements of present day, but date back to 3000–4000 years ago. These achievements include both water and wastewater constructions (such as dams, wells, cisterns, aqueducts, sewerage and drainage systems, toilets, and even recreational structures). These hydraulic works also reflect advanced scientific knowledge, which allowed the construction of tunnels from two openings and the transportation of water both by gravity flow in open channels and by pressurized flow in closed conduits. Certainly, technological developments were driven by the necessities to make efficient use of natural resources, to make civilizations more resistant to destructive natural elements, and to improve the standards of life. With respect to the latter, the Greek (including Minoan) and
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Roman civilizations developed an advanced, comfortable, and hygienic lifestyle, as manifested from public and private bathrooms and flushing toilets, which can only be compared to the modern one, re-established in Europe and North America in the beginning of the last century. Minoan technological developments in water and wastewater management principles and practices are not as well known as other achievements of the Minoan civilization, such as poetry, philosophy, sciences, politics, and visual arts. However, archaeological and other evidence indicate that, during the Bronze Age in Crete, advanced water management and sanitary techniques were practiced in several palaces and settlements. This period was called by the excavator of the palace at Knossos, Sir Arthur Evans, as Minoan after the legendary King Minos. Thus, Crete became the cradle of one of the most important civilizations of mankind and the first major civilization in Europe. One of the major achievements of the Minoans was the advanced water and wastewater management techniques practiced in Crete during that time. The advanced water distribution and sewerage systems in various Minoan palaces and settlements are remarkable. These techniques include the construction and use of aqueducts, cisterns, wells, and fountains, the water-supply systems, the construction and use of bathrooms and other sanitary and purgatory facilities, as well as wastewater and stormwater sewerage systems. The hydraulic and architectural function of the water-supply and sewer systems in palaces and cities are regarded as one of the salient characteristics of the Minoan civilization. These systems were so advanced that they can be compared with the modern systems, which were established only in the second half of the nineteenth century in European and American cities (Angelakis et al., 2010). Water and wastewater technologies developed during the Minoan, Greek, and Roman civilizations are considered in this chapter. Emphasis is given to the water resources development such as aqueducts, cisterns, wells, distribution systems, wastewater and stormwater sewerage systems construction, operation, and management beginning since Minoan times (second millennium BC). The achievements to support the
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Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
hygienic and the functional requirements of palaces and cities during this time were so advanced that could be paralleled only to modern urban water systems that were developed in Europe and North America only in the second half of the nineteenth century (Angelakis and Spyridakis, 1996). It should be noted that hydraulic technologies developed during the Greek and Roman periods are not limited to urban water and wastewater systems. The progress in urban water supply was even more admirable, as witnessed by several aqueducts, cisterns, wells, and other water facilities discovered (Koutsoyiannis et al., 2008). These advanced Minoan technologies were expanded to the Greek mainland in later periods of the Greek civilization, that is, in Mycenaean, Archaic, Classical, Hellenistic, and Roman periods. In this chapter, a rather synoptic description of the main concepts of water and wastewater management during the Minoan, Greek, and Roman civilization is attempted. The main principles and challenges are also discussed.
4.01.1 Aqueducts Aqueducts were used to transport water from a source to the locations where the water was needed, either for irrigation or for urban water supplies, and have been used since the Bronze Age. Aqueduct bridges are man-made conduits for transporting water across rivers, streams, and valleys. As a matter of fact, a systematic evolution of water management in ancient Greece began in Crete during the early Bronze Age, that is, the Early Minoan period (c. 3500–2150 BC) (Myers et al., 1992; Mays, 2007). Starting the Early Minoan period II (c. 2990–2300 BC), a variety of technologies such as wells, cisterns, and aqueducts were used (Mays, 2007).
4.01.2 Minoan and Greek Aqueducts The water distribution system at Knossos, as well as the mountainous terrain and available springs made possible
the existence of an aqueduct (Mays, 2007; Mays et al., 2007). The Minoan inhabitants of Knossos depended partially on wells, and mostly on water provided by the Kairatos River to the east of the low hill of the palace, and on springs. Indications suggest that the water-supply system of the Knossos palace initially relied on the spring of Mavrokolybos (called so by Evans), a limestone spring located 450 m southwest of the palace (Angelakis et al., 2007; Evans, 1921–1935; Mays et al., 2007). In later periods with the increase of population, other springs at further longer distances were utilized. Thus, an aqueduct made of terracotta pipe could have crossed a bridge on a small stream south of the palace which carried water from a perennial spring on the Gypsadhes hill (Graham, 1987; Mays, 2007). A second example of an aqueduct was found in Tylissos (see Figure 1(a)). Parts of the stone aqueduct, with the main conduit at the entrance of the complex of houses, and other secondary systems led the water to a cistern dated at c. 1425–1390 BC (Mays et al., 2007). Other aqueducts were in Gournia, Malia, and Mochlos. These technologies were further developed during the Hellenistic and Roman periods in Crete, and were transferred to continental Greece as well as other Mediterranean locations (Angelakis et al., 2007; Angelakis and Spyridakis, 2010). In the Archaic and the Classical periods of the Greek civilization, aqueducts were built similar to the ones built by the Minoans and Mycenaeans. One of the most renowned watersupply systems is the tunnel of Eupalinos on Samos Island. In fact, it is the first deep tunnel in history that was dug from two openings with the two lines of construction meeting at about the central point of the distance. The construction of this tunnel was made possible by the progress in geometry and geodesy that was necessary to implement two independent lines of construction that would meet (Koutsoyiannis et al., 2008; Mays et al., 2007). The Samos aqueduct system includes the 1036-mlong tunnel and two additional parts for a total length greater than 2800 m. Its construction started in 530 BC, during the tyranny of Polycrates and lasted 10 years. It was in operation until the fifth century AD (Koutsoyiannis et al., 2008).
Figure 1 Ancient Minoan and Greek aqueducts: (a) aqueduct entering Tylissos showing the stone cover and (b) Peisistratean aqueduct consisting of terracotta pipe segments and elliptical pipe openings in each pipe. Copyright permission with LW Mays.
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Obviously, there are several other acknowledged aqueducts in Greek cities since water supply was regarded a crucial and indispensable infrastructure of every city (Tassios, 2007). Aqueducts (either tunnels or trenches) were always subterranean due to safety and security reasons. Usually, at the entrance of the city, aqueducts would branch out in the city to feed cisterns and public fountains in central locations. The aqueducts were pipes (usually terracotta) laying in the bottom of trenches or tunnels allowing for protection. One or more pipes in parallel were used depending upon the flow to be conveyed. The terracotta pipes (20–25 cm in diameter) fit into each other and allow access for cleaning and maintenance by elliptic openings that were covered by terracotta covers (Mays, 2007; Mays et al., 2007). Water conveyed by aqueducts typically originated from karstic springs. As the history teaches us, the presence of natural springs was a prerequisite for the selection of an area to settle. As a matter of fact, the Acropolis at Athens had an aquifer and a spring named Clepsydra. With the intensified urban development as well as the increase of population, the natural springs were not able to cover the water demand. Thus, the increasing water scarcity was remedied by transferring water from distant springs by aqueducts, digging wells, and constructing cisterns for rainwater storage. In Athens all these alternatives coexisted: the Peisistratean aqueduct (see Figure 1(b)) constructed by the end of the sixth century BC was accompanied with numerous wells and cisterns. Legislative and institutional tools were developed in Athens in order to wisely and effectively manage a water-supply system with public and private elements (Mays et al., 2007; Koutsoyiannis et al., 2008). Subsequently, the technologies developed in ancient Greece were transferred to the Greek colonies both to the east in Ionia (Asia Minor, nowadays Turkey) and to the west in the Italian peninsula, Sicily, and other Mediterranean sites, most of which were founded during the archaic period. A brilliant example of this was the founding of Syracuse (on Sicily) as a colony of Corinth in 734 BC (Mays et al., 2007). Later, during the Hellenistic period, further developments were accomplished by the Greeks in the construction and operation of aqueducts due to the progress in science which led a new technical expertise. Hellenistic aqueducts usually used pipes as well as they continued to be subterranean for safety reasons (war, earthquakes, etc.). The scientific progress
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in hydraulic (especially due to Archimedes, Hero of Alexandria) allowed the construction of inverted siphons at large scales to convey water across valleys (lengths of kilometers, hydraulic heads of hundreds of meters) (Koutsoyiannis et al., 2007, 2008; Mays, 2007; Mays et al., 2007).
4.01.3 Roman Aqueducts Springs, by far, were the most common sources of water for aqueducts even with the Romans. Water sources for the Greeks and Roman systems included not only springs, percolation wells, and weirs on streams, but also lakes that were developed by building dams. At ancient Augusta Emerita, at present-day Merida, Spain, the Roman water system included two reservoirs created by the construction of the Cornalvo and the Proserpina dams. The Proserpina dam is an earthen dam, approximately 427 m long and 12 m high. The Cornalvo dam is an earthen dam, approximately 194 m long and 20 m high with an 8 m dam crest width. Both of these dams are still used in the present day, obviously with modifications over the years. Dams were built in many regions of the Roman Empire. Aqueducts consisted of many components, including open channels and pipes. The main types of conduits used by the Romans are: (1) open channels (rivi per canales structiles), (2) lead pipes (fistuli plumbei), (3) earthenware (terracotta) pipes (tubili fictiles), and (4) wood pipes. Open channels were built using masonry or were cut in the rock and flows were driven by gravity, while the lead pipes were used for pressurized conduits including inverted siphons. A scheme representing the general path of a whole aqueduct with the basic elements is presented in Figure 2. Obviously, there are many system configurations that were built by the Romans and Greeks; however, the drawing presents the major components, including the siphon (inverted siphon) which was used in some systems. Various types of pipes constructed by the Romans included terracotta, lead, wood, and stone. One of the most impressive Roman aqueducts in Roman Greece is that in the Aegean island Lesvos (Figure 3). It is probably a work of late second or early third century AD. It was mainly used for water supply of Mytilene town, the capital of the island, and for water supply and irrigation of the southeastern area of the island, by transporting water from the lake of Megali Limni (big lake), at the Olympus mountain,
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Figure 2 Flow sheet and components of a Roman aqueduct: (1) source – caput aquae; (2) steep chutes (dropshafts); (3) settling tank; (4) tunnel and shafts; (5) covered trench; (6) aqueduct bridge; (7) inverted siphon; (8) substruction; (9) arcade; (10) distribution basin/castellum aquae divisorium; (11) water distribution system. From De Feo G and Napoli RMA (2007) Historical development of the Augustan aqueduct in Southern Italy: Twenty centuries of works from Serino to Naples. Water Science and Technology: Water Supply 7(1): 131–138.
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Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 3 Part of the impressive Roman aqueduct rises 600 m west Moria, a Lesvian village at 6 km from Mytilene town: (a) general view of the remains and (b) the base of columns. Copyright permission with AN Angelakis.
where the construction begins. The aqueduct was also fed by other secondary springs, such as the springs at the Agiassou area (i.e., Karini). It was passed through a very anomalous landscape relief; thus, it includes parts on the soil surface, tunnels, and bridges. The total length of the Lesvos aqueduct is 26 km, with a uniform slope of 0.0096 m m1. Its depth ranges from 0.65 to 1.10 m and its width from 0.35 to 0.64 m (Karakostantinou, 2006). Its maximum capacity is estimated to be of 25 000 m3 d1 a along the distance of 26 km, a route that was entirely supported by gravity. Today, the maximum water supply of the town (15 000 m3 d1) is pumping from springs of Ydata located in a lower level of that of Karini (Mytilene Municipal Enterprise for Water Supply and Sewerage, 2009, personal communication. Mytileni, Greece). Its remains at the village of Moria are 170 m long and 27 m in height and consist of 17 arches, also called Kamares laying on their column (Figure 3(a)). Each opening is divided in three successive arches based on columns. The masonry is constructed with the use of emplekton system (Karakostantinou, 2006). The columns and arches were constructed from large blocks of gray marble taken from the island; these materials were very strong and resistant to decay (Figure 3(b)). The distribution of the arches along the openings consists of three at a time – up and down – for every opening. The openings are delimited by columns, and each column has an abacus. Siphons (Figure 2(g)) were built by the Romans also, in fact many of the siphons may very well have been started by the Greeks and completed by the Romans. The siphons included a header tank for transitioning the open channel flow of the aqueduct into one or more pipes, the bends called geniculus, the venter bridge to support the pipes in the valley, and the transition of pipe flow to open channel flow using a receiving tank. Locations of siphons included Ephesus, Methymna, Magnesia, Philadelphia, both Antiochias, Blaundros, Patara, Smyrna, Prymnessos, Tralleis, Trapezopolis, Apameia, Akmonia, Laodikeia, and Pergamon (Mays et al., 2007; Tassios, 2007). These siphons were initially built with terracotta pipes or stone pipes (square stone blocks to which a hole was
carved) such as the inverted siphon at Patara (Turkey), shown in Figure 4 (Haberey, 1972). As shown in the figure this siphon was constructed from carved stone segments. Nevertheless, the need for higher pressures naturally led to the use of metal pipes, specifically from lead. One of the largest siphons was the Beaunant siphon of the aqueduct of the Gier River which supplied the Roman city of Lugdunum (Lyon, France). This siphon had nine lead pipes with a total length of 2.6 km. This siphon was 2600 m long and 123 m deep with an estimated (Hodge, 2002) discharge of 25 000 m3 d1. Pergamon was a city in western Turkey at the present-day city of Bergama. The Helenistic aqueducts constructed were the Attalos, the Demophon, the Madradag, the Nikephorium, and the Asklepieion. The Roman aqueducts constructed were the Madradag channel, the Kaikos, and the Aksu. The Madradag aqueduct which had a triple pipeline (terracotta pipe) of more than 50 km long included an inverted siphon (made of lead) longer than 3.5 km with a maximum pressure head of about 190 m (Mays et al., 2007; Tassios, 2007). It took another 2000 years later before another pipeline was constructed that could bear a higher pressure (Fahlbusch, 2006). In particular, the Attalos aqueduct was the first pipeline (buried of fired clay, and 13 cm inner diameter) in Pergamon, and it was probably constructed in the middle or second half of the third century BC, bringing water from a spring in the mountains north of Pergamon (Fahlbusch, 2006; Mays, 2007; Oziz, 1987, 1996). The Romans built mega water-supply systems including many magnificent structures. As a matter of fact, Roman aqueducts became very famous all over the world, with Rome’s water-supply system being considered one of the marvels of the ancient world (Hodge, 2002; De Feo and Napoli, 2007; De Feo et al., 2009b; Mays, 2007; Mays et al., 2007). In fact, the Romans were urban people and consumed enormous amount of drinking water in order to supply baths, public and decorative fountains, residences, garden irrigation, flour mills, aquatic shows, and swimming pools (Hodge, 2002; Tolle-Kastenbein, 2005; De Feo and Napoli, 2007; De Feo et al., 2009b; Mavromati and Chryssaidis, 2007). However, the
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
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Figure 4 Inverted siphons. (a) Inverted siphon at Patara (Turkey) made of stone pipes. (b) Reconstruction of siphon of the aqueduct of Gier, near Beaunant, France that supplied water to Ancient Lugdunum, showing ramp of siphon with header tank on the top and the nine lead pipes of the siphon. (a) From Mays LW (ed.) (2010) Ancient Water Technologies. Dordrecht: Springer and (b) From Haberey W (1972) Die ro¨mischen Wasserleitungen nach Ko¨ln. Bonn: Rheinland-Verlag.
Roman aqueducts were not built with the primary purpose of providing drinking water, nor to promote hygiene, but rather to supply the thermae and baths or for military purposes (Hodge, 2002; De Feo and Napoli, 2007; De Feo et al., 2009b). The description of the ancient Roman water-supply system is contained in some recommendations of the Latin writers: Vitruvius Pollio (De Architectura, book VIII), Plinio the Elder (Naturalis Historia, book XXXVI), and Frontinus (De Aquaeductu Urbis Romae). Roman hydraulic engineering borrowed from the experiences and techniques of the Greeks and Etruscans. However, the size of the works as well as the technical-organizational features of distribution started with them. The common Greek practice was based on underground conduits, following courses determined by terrain features (Martini and Drusiani, 2009). The Etruscan civilization flourished in central Italy from the VIII century BC onward. The Etruscan talent for water and land management is highlighted by the existence of an imposing number of works (tunnels and channels) spread over their territories of Latium and, to a lesser amount, of the other Etruscan areas (Bersani et al., 2010). The construction of an ancient Roman aqueduct was not different from the modern practice, with several modern technologies coming from Roman engineering. The building of an aqueduct started with the search for a spring. Water was collected after permeating through vaults and walls of
draining channels and settled. From the spring, water flowed into an open channel flow and air was present over the water surface (Monteleone et al., 2007). The water in the aqueducts descended gently through concrete channels. During the route, there were multitiered viaducts, inverted siphons, and tunnels to exceed valleys or steep points. At the end of its course, the channel entered into a so-called piscina limaria, a sedimentation tank to settle particulate impurities. Then, the channel flowed into a partitioning tank called castellum divisorium where there were some walls and weirs to regulate the water flowing into the urban pressure pipes (De Feo and Napoli, 2007; Monteleone et al., 2007). Rome originally used water directly from the river Tiber as well as wells and many small springs existed inside its town area, such as Acque Lautole, Acque Tulliane, Fonte Giuturna, and Fonte Lupercale. However, since the fourth century BC, Rome gradually built aqueducts (Bono and Boni, 1996). Aqua Appia was the first aqueduct built in Rome in 312 BC. It was entirely underground for a total length of around 16.561 km, equivalent to 11 190 passus (1 passus ¼ 1.48 m) and an average flow rate of 73 000 m3 d1, corresponding to 1825 quinariae (1 quinaria B40 m3 d1) (Table 1; Panimolle, 1984). It is important to specify that a quinaria has not been scientifically defined. As a matter of fact, a quinaria was a pipe of 2.3125 cm diameter and there is no unanimity on how much water is a quinaria (Rodgers, 2004). During the subsequent
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Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Table 1
Characteristics of the 11 Imperial Age Roman aqueducts
Location
Dating
Length (km)
Aqua Appia Anio Vetus Aqua Marcia Aqua Tepula Aqua Julia Aqua Virgo Aqua Alsietina Aqua Claudia Anio Novus Aqua Traiana Aqua Alexandrina Average Total
312 BC 273 BC 144 BC 127 BC 33 BC 19 BC 2 BC 52 AD 52 AD 109 AD 226 AD
16.561 63.640 91.331 17.800 22.830 20.875 32.882 68.977 86.876 58.000 22.000 45.616 501.772
Underground length (km (%)) 16.472 (99.5%) 63.312 (99.5%) 80.286 (87.9%) 12.470 19.040 32.814 53.620 72.964
(54.6%) (91.2%) (99.8%) (77.7%) (84.0%)
43.872 (86.8%) 350.978
Average slope (m km1)
Flowrate (m3 d1)
0.6 3.6 2.7 5 12.4 0.2 6 3.8 3.8 3.8 1 3.9
73 000 175 920 187 600 17 800 48 240 100 160 15 680 184 280 189 520 113 100 21 025 102 393 1 126 325
From Panimolle G (1984) Gli Acquedotti di Roma Antica (The Aqueducts of Ancient Rome). Rome: Edizioni Abete; Adam JP (1988) L’Arte di Costruire presso i Romani. Materiali e Tecniche (Roman Building: Materials and Techniques). Milan: Longanesi; Bono P and Boni C (1996) Water supply of Rome in antiquity and today. Environmental Geology 27: 126–134; Hodge AT (2002) Roman Aqueducts & Water Supply, 2nd edn. London: Gerald Duckworth; Rodgers RH (2004) Sextus Iulius Frontinus. On the Water-Management of the City of Rome. De Aquaeductu Urbis Romae. Cambridge: Cambridge University Press.
500 years, 10 more aqueducts were constructed. The last great aqueduct built in Rome in ancient times was the 22-km-long Aqua Alexandrina. On the whole, the 11 Imperial Age Roman aqueducts had a total flow rate of 1.13 106 m3 d1 and a total length of more than 500 km. Since the population of Rome at the end of the first century AD was about 500 000 inhabitants (Bono and Boni, 1996), a mean specific discharge of B2000 l inhabitant1 d1 was produced. This value is extraordinary if compared with the current specific water use of B200–300 l inhabitant1 d1. Nowadays, the popular but inaccurate image is that Roman aqueducts were elevated throughout their entire length on lines of arches, called arcades. Roman engineers, as their Greek predecessors, were very practical and therefore whenever possible the aqueduct followed a steady downhill course at or below ground level (Hansen, 2006). As a matter of fact, Table 1 shows that on average 87% of the length of the Rome’s aqueduct system was underground. The longest aqueduct in the Roman world was constructed in the Campania Region, in Southern Italy. It is the Augustan Aqueduct Serino-Naples-Miseno, which is not well known due to there being no remains of spectacular bridges, but it was a masterpiece of engineering. The Serino aqueduct was constructed during the Augustus period of the Roman Empire, probably between 33 and 12 BC when Marcus Vipsanius Agrippa was curator aquarum in Rome, principally in order to refurnish the Roman fleet of Misenum and secondarily to supply water for the increasing demand of the important commercial harbor of Puteoli as well as drinking water for big cities such as Cumae and Neapolis. The main channel of the Serino aqueduct was approximately 96 km long, and had seven main branches to towns such as Nola, Pompeii, Acerra, Herculaneum, Atella, Pausillipon, Nisida, Puteoli, Cumae, and Baiae (De Feo and Napoli, 2007; De Feo et al., 2010). In summary the Romans made great contributions to the advancement of the engineering of aqueducts. Fahlbusch
(2006) points out the following from examination of many aqueducts: 1. size of the aqueduct channel was chosen according to the estimated discharge and the size varied along the course of the aqueduct; 2. the cross section was large enough for people to walk through the channel for repair and maintenance, particularly to remove calcareous deposits; and 3. the cross section was kept constant allowing manifold uses for encasings, especially the soffit scaffoldings for the vaults in a kind of industrialized construction.
4.01.4 Cisterns and Reservoirs In general, cisterns were usually constructed in order to store rainwater for domestic use (private houses), with a volume in the order of dozens of cubic meters, while reservoirs were realized in order to store flowing water with a volume in the order of thousands of cubic meters (Tolle-Kastenbein, 2005; De Feo et al., 2010). The Minoan and Mycenaean settlements used cisterns a 1000 years before the classical and Hellenistic-Greek cities. Cisterns were used to supply (store runoff from roof tops and court yards) water for the households through the dry summers of the Mediterranean. In ancient Crete, in particular, the technology of surface and rainwater storage in cisterns for water supply was highly developed and has continued to be used in modern times. One of the earliest Minoan cisterns was found in the center of a pre-palatial house complex at Chamaizi dating back to the turn of the second millennium BC. It is located on the summit of a hill and its rooms were situated around a small open court with a deep circular rock-cut cistern, 3.5 m in deep and with a diameter of 1.5 m, lined with brickwork in its upper part (Davaras, 1976; Mays et al., 2007; Angelakis and
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Spyridakis, 2010). Four of the earliest Minoan structures which may be considered to be large cisterns were built in the first half of the second millennium BC at Pyrgos-Myrtos (Ierapetra), Archanes, Tylissos, and Zakros (Cadogan, 2007; Mays et al., 2007; Angelakis and Spyridakis, 2010). While, at Phaistos, water supplied to cisterns depended on precipitation collected from rooftops and courts, a supplementary system was needed to satisfy the needs of water supply, especially in this particular area where agriculture was widely practiced. Thus, water was probably taken from wells in a location southwest of the palace which was rich in groundwater and surface water, and from the river Ieropotamos located to the north, at the foot of the Phaistos hill (Gorokhovich, 2005; Mays et al., 2007; Angelakis and Spyridakis, 2010). There were also cisterns on the high grounds above the Minoan palace in Malia, in a site lying in a narrow plain between the mountains and the sea. At the famous Phaistos palace, cisterns depended on precipitation collected from rooftops and yards. A supplementary system of water supply was needed to satisfy the needs of water supply, especially in those areas where agriculture was intensive. The cisterns were connected to small channels collecting spring water and/or rainfall runoff from catchment areas. The use of cisterns preceded channels or aqueducts in supplying the palace and the surrounding community with water (Mays et al., 2007; Angelakis and Spyridakis, 2010). Most Greek houses had a cistern supplied by rainwater for purposes of bathing, cleaning, houseplants, domestic animals, and even for drinking during shortages of water. Most likely, the water was of a quality that would be subpotable using today’s standards. Aristotle in his Politics (vii, 1330 b) written around 320 BC asserted that ‘‘cities need cisterns for safety in war.’’ During this time a severe 25-year drought required the collection and saving of rainwater. Also about this time cisterns were built in the Athenian Agora for the first time in centuries (Crouch, 1993; Mays, 2007). In particular, in the ancient Greek city of Dreros on Crete, there is a rectangularshaped cistern with dimensions of approximately 13.0 5.5 6.0 m3 (Antoniou et al., 2006; Mays, 2007). In ancient Crete, the technology of surface and rainwater storage in cisterns is continued to be used even today. Four of the earliest Minoan structures which may be considered to be large cisterns were built in the first half of the second millennium BC (the time of the first Minoan palaces) at PyrgosMyrtos (Ierapetra), Archanes, Tylissos, and Zakros (Angelakis et al., 2010). The Tylissos cistern is shown in Figure 5(a). This technology has been further improved during the Hellenistic and Roman periods. An impressive pillar of two interconnected cisterns, 40 m deep cut in the rock, has been discovered in ancient city Eleutherna (Figure 5(b)). The dimensions of the two cisterns are 40 25 m2 and the depth 4.5 m. The city flourished in the early Christian times and the water was transported from a spring through an aqueduct of about 3 km long to the cisterns. The water supply of the city including the thermes was transported through a 150-m-long channel with dimensions of 1.5 2.0 m2. The advanced water-supply technologies developed in Minoan Crete were expanded and improved during the Roman domination of the Greek world. Two such examples with a relatively small but impressive cistern in Minoan city and one of the two huge cisterns
9
(of about 3000 m3 each) in Aptera city in the western Crete are shown in Figures 5(c) and 5(d), respectively. During the classical age (the period between the Archaic and Roman epoch), the political situation was characterized in the Greek world (mainly Greece and Asia Minor) by wars among the various cities. In this period, no springs or deep wells existed, so cisterns were constructed to collect rainfall during the winter season. These cisterns were dug into the rock and were mostly pear-shaped with at least one layer of hydraulic plaster that prevented water loss. The cisterns varied in size from 10 m3 to thousands of cubic meters and possibly supplied more than 10 000-people baths and thermes. To prevent contamination of water the mouth of the cistern was covered to keep out dust and debris, and to prevent light from entering, avoiding the growth of bacteria and algae. Reservoirs constructed by the ancient Romans were set low in the ground, or actually underground, and roofed over, by means of concrete vaulting. The roofing vaults were supported by rows of columns, piers, or wall pierced with doors to allow the water to circulate. In some cases, the floor was slightly concave with a drain in the middle, to permit cleaning (Hodge, 2002; De Feo et al., 2010). In general, in the Roman world the reservoirs had two functions: a reservoir could be a reserve for use when the aqueduct ran low or by adding in a little from the tank everyday to supplement supplies until the aqueduct discharge picked up again. When the daily consumption exceeded what the aqueduct could bring in, at least in the hours of daylight, the reservoir was topped up every night to meet the next day’s demands (Hodge, 2002; De Feo et al., 2010). An example of a Roman reservoir is the Bordj Djedid at Carthage in Tunisia, into which the Carthage aqueduct emptied after a run of no less than 90.43 km from its source. This great reservoir was oblong, 39.0 154.6 m2, the size of an entire city block, and subdivided into 18 transverse compartments. Its capacity was 25 000–30 000 m3, representing about a day and a half’s discharge for the aqueduct (Hodge, 2002; De Feo et al., 2010). Remaining in Tunisia, in the center of the city of Dougga/Thugga, there are two very large reservoirs. The first one is the Ain El Hamman reservoir with five aisles, while the second one is the Ain Mizeb reservoir with seven aisles. The two reservoirs have a combined storage volume of 15 000 m3 (Tolle-Kastenbein, 2005; De Feo et al., 2010). Large reservoirs were constructed not only in Northern Africa but also in Europe, especially in Italy and in Turkey. Since a Roman thermae required an enormous quantity of water for its functioning, a huge reservoir had to be constructed. As a matter of fact, the reservoir of the Baths of Caracalla (located in an area of over 100 000 m2) could contain over 80 000 m3 in the numerous cells, situated into two parallel aisles and onto two floors. The oldest baths of Traiano received water supply from a reservoir of around 10 000 m3 (Tolle-Kastenbein, 2005; De Feo et al., 2010). The greatest baths of Diocletian occupied about the same area as those of Caracalla (a rectangle of about 356 316 m2) and closely resembled them in the plans. The reservoir by which the baths were supplied was fed by the aqua Marcia, the volume of which was increased by Diocletian. It was trapezoidal in shape, 91 m in length, with an average width of 16 m. This reservoir, called Botte di Termini (Barrel of Termini), was
10
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 5 Minoan, Hellenistic, and Roman water collection and storage cisterns: (a) Minoan at the ancient town of Tylissos; (b) Hellenistic at the city of Eleutherna; (c) Roman at the Minoa town; and (d) Roman at town of Aptera. Copyright permission with AN Angelakis.
destroyed during 1876 in order to build the Termini railway station, whose name derives from that of the baths (De Feo et al., 2010). In the three centuries of the Roman imperial age, the reservoirs were designed in almost all the architectural forms and in almost all the techniques of masonry known: arcs (especially transversal arcs), turned (especially barrel vault), carrying pillars or groups of pillars, walls of stones and bricks, opus caementicium; while columns were still not used. In fact, the columns were introduced by architects famous for their works of hydraulic engineering in the present-day Istanbul. They created a host of columns hidden in the heart of the capital of the Roman Empire (Tolle-Kastenbein, 2005; De Feo et al., 2010). As a matter of fact, the name of the first reservoir means ‘with a 1001 pillars’. It is the Binbirdirek reservoir which was built under the order of Philoksenos, a Senate member in the Constantinus I period of the fourth century. During the Roman period, Istanbul’s water requirements were met by water brought from distant parts of Thrace. For this reason, the Byzantines built large reservoirs in order to be able to withstand long sieges (De Feo et al., 2010). The Binbirdirek reservoir covered an area of 3640 m2 and had a capacity of around 32 500 m3 of water. It measured 66 56 m2 and was carried by 224 columns consisting of
16 rows, each one having 14 columns, all of which are equal in length, and every column carries the signature of its master (‘1001’ was used to emphasize the great number of columns). There is a thick overlapping astragal running round the columns carrying the vaults and arches and they are in the form of a truncated pyramid and are without decoration. The relief cross on one of the columns is good proof that the reservoir was built in the fourth century, after the Byzantines accepted Christianity. In order to construct ceilings 14–15 m2 high, a second layer of columns was fixed over the marble rings on the first layer of columns. When the palace was destroyed in the sixth century, the cistern was restored. After the Ottoman conquest of Istanbul in 1453, new reservoirs were built and the Binbirdirek was no longer used (De Feo et al., 2010). One of the magnificent historical constructions of Istanbul is the Yerebatan Saray (or Basilica Cistern), located near the southwest of Ayasofya (Hagia Sophia). This huge reservoir was rebuilt by the emperor Justinian (527–565) after the Nika revolt (532). It is a large, vaulted space; the roof rests on 12 rows of 28 marble columns, which are about 9 m high. As the total surface is 65 138 m2, the maximum capacity is almost 85 000 m3, which was brought to this cistern from a well B20 km away with a new aqueduct, also built by
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Justinian. It was used to provide water to the imperial palace (hence the name, imperial cistern). The 336 columns (246 are still visible) were brought to the Basilica Cistern from older buildings. Again, it is narrated that 7000 slaves worked in the construction of the cistern. In fact, the cistern borrowed its name from Ilius Basilica in the vicinity (Lendering, 2008; Ku¨ltu¨r, 2008; De Feo et al., 2010). Another huge Roman reservoir in ancient Constantinopolis (today’s Istanbul) is the Sultan’s Cistern. We do not have any verifiable scientific evidence for its construction date; at the earliest, it could be late fourth century AD, judging by the presence of crosses carved into the upper parts of the column heads. It has a rectangular plan and the whole is divided into five equal rectangular parts by the use of 28 columns, with 7 in granite and 21 in marble, placed equidistant from each other, also supporting the roof with vaulted arches (De Feo et al., 2010). The last Roman underground hydraulic marvel is the spectacular Piscina Mirabilis in Misenum, in the Southern Italy. The Piscina Mirabilis is located in the present-day Municipality of Bacoli, in Miseno (the ancient Misenum), up the hill facing the sea in the bay of Naples. It was constructed during the Augustan Age in order to supply water to the Classis Praetoria Misenensis (Adam, 1988; Hodge, 2002; De Feo and Napoli, 2007; De Feo et al., 2010). The Piscina Mirabilis is a gigantic reservoir 72 m long and 27 m large, with a volumetric capacity of 12 600 m3 of water (Figure 6). It is dug in a tufa hill and has two step entrances in the northwest, the Ancient Roman entrance and southeast corners, the latter closed. Forty-eight pillars, arranged on four rows serving as a support to the barrel vault, divide it into five principal aisles on the long sides (Figure 7(a)) and 13 secondary aisles on the short sides (Figure 7(b)), giving it the majestic look of a cathedral. The long walls were built in opus reticolatum (reticular work) with brick bonding courses and by the technique of the tufa stone pillars, both covered with a thick waterproof layer of opus signinum (pounded terracotta). There is a basin of 1.10 m, probably a polishing pool, which is a waste bath for the maintenance of the reservoir, in the floor of the nave. It was used as a Piscina limaria for the periodical cleaning of the reservoir (Figure 7(c)). The water was lifted through a series of openings (doors) in the vault along the central nave, hydraulically to the covering terrace of the reservoir, and from there, flowed in channels to the urban area. These doors appear casually opened in the roof (Figure 7(d)), with an irregular realization being noted (Adam, 1988; Hodge, 2002; De Feo and Napoli, 2007; De Feo et al., 2010). Russo and Russo (2007) estimated a total daily demand of 12 000 m3 of water for Misenum, including 4000 m3 for the fleet and 8000 m3 for daily demands and for the thermal baths and gardens (based upon daily individual requirements of 100 liters per capita and equal requirements for thermal baths and gardens). The estimated total daily demand is similar to the capacity of the Piscina Mirabilis. Close to the Piscina Mirabilis are two other large cisterns, probably belonging to large villas, the Grotta Dragonaria and Cento Camerelle (Nerone’s jail). In Pozzuoli, the aqueduct served several cisterns, notably the Piscina Cardito (55 16 m2) from the second century, and the Piscina Lusciano (35 20 m2) from the first century AD (De Feo and Napoli, 2007; De Feo et al., 2010).
11
4.01.5 Water Distribution Systems Water distribution systems are aimed at distributing water from reservoirs or aqueducts to the end users. The modern systems are based on the use of pipes. Regarding this aspect, the Minoan society was surprisingly modern. As a matter of fact, in the Knossos palace, the water supply was furnished by means of a network of terracotta pipe conduits (60–75 cm flanged to fit into one another and cemented at the joints) beneath the floors at depths that vary from a few cm up to 3 m (Koutsoyiannis et al., 2008; Angelakis and Spyridakis, 2010). Possibly, the piping system was pressurized (Mays, 2007). Similar terracotta pipes were discovered in some other Minoan sites. In particular, Tylissos was one of the important cities in Ancient Crete during the Minoan era, flourishing (2000–1100 BC) as a peripheral center dependent on Knossos. From the aqueduct, secondary conduits were used to convey water to a sedimentation tank (Figure 8; Mays, 2010) constructed of stone before its storage to the cistern shown in Figure 5(a). Terracotta pipes have also been found at Vathypetro, as well as in the Caravanserai (Guest House), south of the Knossos palace with some also having been found scattered in the countryside (Angelakis and Spyridakis, 2010). The study of the ruins of Pompeii gives a clearer understanding of a Roman urban water distribution system. But this statement does not mean that all Roman cities are identical to Pompeii. The ending point of a Roman aqueduct was the castellum divisorium which had the double function of serving as a disconnection between the aqueduct and the urban distribution network as well as dividing the water flow to various uses and/or geographical areas of the city (Figure 9). In the beginning, Pompeii was not supplied by the Serino aqueduct. As there were no springs in Pompeii, wells were dug to supply water. It is also very likely that Pompeii received water via an aqueduct from the mountains due northeast of Avella. The town must have had a long-distance water supply, quite some time before the Augustan Age, probably around 80 BC. When the Serino aqueduct was built under Augustus, it crossed the course of the older Avella aqueduct between the Apennines and Mount Vesuvius, and both aqueducts were united into a single system (De Feo and Napoli, 2007). The castellum divisorium of Pompeii was housed inside a large brick building near the Vesuvian gate (Figure 10(a)). The supply channel entering the building is 30 25 cm (Figure 10(b)). The flow in this distribution structure was allowed to expand into a wide, shallow tank, separated into three equal compartments (masonry structures) (Figure 10(c)). Flow from each compartment entered a lead pipe. Some feel that the three pipes were connected separately to public fountains, the second to the thermal baths and the third to private users (Hodge, 2002; Russo and Russo, 2007). From the exits the water flowed into lead pipes. There is also the distinct possibility that the three pipes were directed to different geographical areas of Pompeii. Assuming that the pipes did convey water separately to the three major uses as presented by Hodge (2002), the central pipe was directed to the public fountains and had a 30 cm external diameter, whereas the two side ones were 25 cm in diameter. The three gates were of different heights. Thus, the highest gate, which was that serving private houses, cut off their supplies until and unless the water level in
12
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations 1.2
4.9
1.2
4.0
1.2
4.0
1.2
4.0
1.2
4.9
1.2
11.4
27.0
1.2
4.3 4.3 1.2 4.3 4.3
( Measures in meters )
1 2 3 4 5
2.0
1.2
4
Legend
N
11.4
9.4
A
Inlet water Ancient Roman entrance - 1 Piscina Limaria Outlet washing water Ancient Roman entrance - 2
1
B
1.2
4.9
1.2
4.3
2
1.2
4.3
Longitudinal section A-A
1.2
4.3
1.2
4.3
1.2
72.0
1.2
A
4.3 1.2
3
A
Plan of the Roman Piscina Mirabiliis
1.2
3.0
5
1.2
4.3
1.2
4.3
1.2
4.9
1.2
B
10.4
Trasversal section B-B Figure 6 Plan and sections of the Piscina Mirabilis. Modified from De Feo G, De Gisi S, Malvano C, and De Biase O (2010) The greatest water reservoirs in the ancient Roman world and the ‘‘Piscina Mirabilis’’ in Misenum. Water, Science and Technology: Water Supply 10(4) (in press).
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
13
Figure 7 Piscina Mirabilis: (a) a cross aisle; (b) a longitudinal aisle; (c) internal piscina limaria; and (d) a hole in the barrel vaulted roof.
the main body of the castellum rose high enough to spill over it and start flowing down the channel; on the contrary, the lowest gate (that in the center) governed access to the public fountains, which, if the water level sank, were thus the least to dry up. The private users had no minimum water entitlement until the needs of the public fountains and thermal baths had been satisfied (Hodge, 2002). From the castellum divisorium, the three pipes lead the water to different parts of the city filling water towers: the castellum secondarium or castellum privatum (Figure 10(d)). The water
towers were lead tanks positioned on top of brick masonry pillars, 6 m tall, located at crossroads and connecting small numbers of customers. They also supplied public fountains. The single user had to pay to obtain water for his premises. The water was metered by means of bronze orifices, the calices connecting the customers’ pipes (usually quinariae pipes) to the castellum privatum lead tank. In Pompeii, case calices were placed at the bottom of the lead tanks, and pipes fit into cavities left in the brick pillars (Hodge, 2002; Monteleone et al., 2007).
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Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 8 Water system at Tylissos, Crete, Greece with sedimentation tank in foreground with stone channel connecting to cistern in background. (Mays, 2010, Copyright permission with LW Mays).
Aqueduct
Castellum divisorium
Head 18 m
Castellum secondarium Head 6m
Figure 9 Flow sheet of a Roman urban water distribution systems based on Pompeii. Modified from Hodge AT (2002) Roman Aqueducts & Water Supply, 2nd edn. London: Gerald Duckworth.
The lead tank on the water tower acted as a disconnection between the system at high pressure upstream and the customers’ pipes downstream. Connecting water derivation pipes elsewhere in the castellum privatum was against the regulations. The only connection available had to be arranged with the water office discussing the quantities for consumption. This water-supply system clearly shows that water towers could break from the pressure built up in the mains descending from the initial castellum divisorium at the top point of the city, with excess water overflowing into streets drains. As shown in Figure 9, the maximum height of water over the tap was about 6 m, without accounting for the pressure losses in the delivering pipes (Hodge, 2002; Monteleone et al., 2007). Lead pipes (Figure 11) in Pompeii are of the same construction and appearance as found in other Roman cities. The water taps found in Pompeii were also similar to those found in other Roman cities. Only a small number of houses had
a water pipe that supplied a private bath or basins in the kitchen, in the toilet, or in the garden.
4.01.6 Fountains The Minoan civilization gave an extraordinary contribution to the development of water management practices also in terms of fountains. The main examples of Minoan fountains are subterranean structures supplied with water directly or from springs via ducts. The construction of steps or alternatively the shallow basins indicates that water was taken out with the use of a container. This recalls the type of fountain of the later Classical and Hellenistic period called arykrene. The most typical of them is that of the Zakro palace. Another fountain similar to the Tykte was found at the Guest House (Caravanserai) of Knossos in the Spring Chamber. A ritual function of
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
15
Figure 10 Pompeii: (a) brick building near the Vesuvian gate housing the castellum divisorium; (b) inside castellum divisorium; (c) supply channel; and (d) a castellum secondarium.
the particular fountains is also argued, as artifacts of ritual content have also been unearthed. Another type known in later periods as rookrene, which constantly provided freshwater, was also found in Zakro with two zoomorphic waterspouts. Finally, a remarkable fragment from a fresco composition depicting a fountain of a supposedly Minoan garden was found in the House of Frescoes in Knossos (Angelakis and Spyridakis, 2010). During the Roman period, public fountains were usually located in the street. For example, in Pompeii the fountains were located at fairly evenly spaced intervals of about 100 m, and it was rare for anyone to carry their water for more than 50 m (Hodge, 2002). The simplest form of street fountain was normally equipped with an oblong stone basin, typically about 1.5 1.8 m2 and 0.8 m high, into which the spout discharged, and which presumably was normally full. The fountains were deliberately designed to overflow in order to clean the street (Hodge, 2002; De Feo et al., 2010). Not far from the city of Pompeii, in the District of Salerno, there is a Roman gallery in rock in the village of Sant’Egidio del Monte Albino in the Sarno River basin. The gallery was constructed in order to supply a public fountain which stands on the structure of an ancient Roman villae (the Helvius
villae). The Helvius fountain was a public fountain, but it was quite different from the public fountains in nearby Pompeii (Figure 12(a)). As a matter of fact, the Helvius fountain was constructed neither by means of matched slabs nor in limestone nor in Vesuvian stone. It was built as a single block of white marble. Moreover, there is another particular aspect which differentiates the Helvius fountain from the Pompeian fountains (Figure 12(b)). The Helvius fountain has a sculptural decoration on the three available sides representing the river Sarno along its path from the spring toward the sea (De Feo et al., 2010). Figure 13 shows two additional Roman fountains that are quite different from those previously mentioned. Figure 13(a) shows a fountain in Chersonesos (Crete) and Figure 13(b) the Fountain of Trajan in Ephesus (Turkey), dedicated by Aristion, AD 102/114.
4.01.7 Drainage and Sewerage Systems and Toilets Drainage systems were used for the disposal of surplus water, and were found both in cities (to carry rainfall, overflow from fountains and bathrooms) and in the country (to prevent
16
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 11 Components of lead pipe system found in Pompeii: (a) lead pipe and joint found along the street; (b) junction box; and (c) manifold. Copyright permission with LW Mays.
flooding in the fields). Sewerage systems were used for the conveyance of domestic wastewater, and were only found in cities, where they were necessary due to a high population density (Hodge, 2002). However, in most cases, combined systems of flow rates composed mainly of rainfall runoff and wastewater were applied. The Minoan civilization also gave an extraordinary contribution to the development of water management practices in terms of drainage and sewerage systems. As a matter of fact, Minoan palaces were equipped with elaborate storm drainage and sewer systems (MacDonald and Driessen, 1988). Open terracotta and stone conduits were used to convey and remove stormwater and limited quantities of wastewater.
Pipes, however, were scarcely used for this purpose. Larger sewers, sometimes large enough for a man to enter and clean, were used in Minoan palaces at Knossos, Phaistos, and Zakro. These large sewers may have led to the conception of the idea of the labyrinth, the subterranean structure in the form of a maze that hosted the Minotaur, a hybrid monster. The end section of the main part of the sewerage system of the Knossos palace is shown in Figure 14(a). The outlet of the Phaistos palace system appears to be similar (Figure 9(b)). Note that Evans (1921–35) and Darcque and Treuil (1990) considered that the main part of the system had been planned and constructed originally in Middle Minoan time. The main disposal sites at the Knossos and Zakros palaces were directed
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
17
Figure 12 Public fountains: (a) in Pompeii (matched slabs) and (b) in the basin of the Sarno river (single block of white marble).
Figure 13 Roman fountains: (a) fountain in Hersonissos (Crete) and (b) remains of the fountain of Trajan in Ephesus (Turkey), dedicated by Aristion, AD, 102/114. Copyright permission with LW Mays.
to the Kairatos River and to the sea, respectively. However, there are indications that in the palace of Phaistos and in the villa of Agia Triadha, cisterns were also used as disposal sites of surface water, along with appropriate landforms. Particularly in the palace of Phaistos, agricultural land located in the south site of the palace was used as disposal site of the both the wastewater and stormwater instead of the river Ieropotamos crossing the northern site of the Phaistos hill. In all cases of palaces and cities, there is an increased slope of the central sewers toward of their outlets; thus, anaerobic conditions have been maintained and the odors have been avoided. In addition to the very effective drainage and sewerage systems, some palaces had toilets with flushing systems operated by pouring water in a conduit. However, the best example of such an installation was found on the island of Thera (Santorini) in the Cyclades, Greece. This is the most eloquent and best-preserved example belonging to the early late-Minoan period (c. 1550 BC) in the Bronze Age settlement of Akrotiri, which shares the same cultural context of Crete (Angelakis and Spyridakis, 2010). At the beginning, for some centuries, the collection and discharge of rainwater runoff was managed by separate sewers.
As a matter of fact, rainwater was carried in simple channels carved into the rock in cities with bedrock (i.e., the Acropolis of Athens). Otherwise, the channels were covered with rocks. A system for the simultaneous discharge of both rainwater and domestic sewage was invented during the Greek period (Tolle-Kastenbein, 2005). Ancient drainage and sewerage systems were usually developed on four levels. The initial channels coming from buildings (first order) ended in street channels of second order, which prosecuted in principal channels with an increasing size (third order) and ended in a final huge collection channel (fourth order), usually present only in big cities. The great drain of Athens was first designed as a rainwater drainage system. However, in the first quarter of the fifth century BC, it received domestic sewage and ended in a huge collection channel (fourth order) similar to the Roman Cloaca Maxima (Tolle-Kastenbein, 2005). The Cloaca Maxima is the best-known ancient urban drain. Tradition ascribes its construction to Tarquinius Priscus, king of Rome 616–578 BC. The Cloaca Maxima (4.2 m high, 3.2 m wide) was covered by stone vaulting, while its bottom was paved with basalt pavers. It combined the three functions of
18
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 14 Outlet of the central Minoan sewerage and drainage systems: (a) palace of Knossos and (b) palace of Phaistos. Copyright permission with AN Angelakis.
wastewater and rainwater removal and swamp drainage. As it is well known, the exit from the Cloaca Maxima drain into the river Tiber still exists in Rome, but now partly hidden by the modern Lungotevere Embankment (Hodge, 2002). The street drains of Pompeii are very famous. At the time of the famous Vesuvius eruption, the drains existed only in the area around the forum. The streets were a sort of open channel conveying water coming from public fountains, rainwater, and segregate sewage. Therefore, as shown in Figure 15, streets had raised sidewalks (50–60 cm high) with stepping stones (pondera) at the street corners to enable pedestrians to cross from one side to the other without stepping down (Hodge, 2002). Toilets have a long history. The first evidence of the purposeful construction of bathrooms and toilets in Europe comes from Bronze Age Minoan (and Mycenaean) Crete in the second millennium BC (Vuorinen et al., 2007). In the palace of Knossos, rainwater was probably used to flush the toilet near the Queen’s Hall (Figure 16; Angelakis et al., 2005). The Hellenistic period is considered more progressive for the sanitary and purgatory engineering during the antiquity, although the considerable spreading of these systems occurred during the Roman era. The Romans applied the earlier techniques in larger constructions, using the advantages of their
building methods with concrete walls and vaulted roofing. Moreover, due to their improved aqueduct technologies, they could provide natural water flow in most public latrines. It is also evident that such structures and installations have survived until the end of the ancient world and have been implemented during the beginning of the Byzantine period. The customs of the new religion, Christianity, modified some of the structures in terms of privacy in bathing facilities (Antoniou and Angelakis, 2009). During the Hellenistic era lavatories improved significantly, followed by their spread throughout the Roman Empire. The features of the typical ancient lavatory are the bench-type seats with keyhole-shaped defecation openings and an underneath ditch. The ditch was both a water-supply conduit for flushing and a sewer. Figure 17 shows remains of a public toilet in Ephesus (Turkey) illustrating the bench seats, the defection openings, and the small channel on the floor for cleaning the sponghia. The lavatory was usually situated in the area of the building most convenient for water supply and/or sewerage. In many cases, the water for the flushing was reused either after other domestic or communal activities. Despite privacy, lavatories were used in antiquity by many people simultaneously, from two to three people in the small domestic latrines and up to 60 people in the larger public latrines
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
19
Figure 15 Stepping stones (pondera) in Pompeii.
Wooden seat
Door jamb
Gypsum floor
likely lacked running water and they were commonly located near the kitchens. All this created an excellent opportunity for the spreading of intestinal pathogens (Vuorinen et al., 2007). Hygienic conditions in both types of toilets must have been very poor, and consequently intestinal diseases were diffused. Dysentery, typhoid fever, and different kinds of diarrheas are likely candidates for diagnoses. Unfortunately, descriptions of the intestinal diseases in the ancient texts are so unspecific that the identification of the causative agent is a very problematic venture. Studies of ancient microbial DNA might offer some new evidence for the identification of microbes spread by contaminated water (Vuorinen, 2010).
Sewer
Seat
Hood
Sewer
Flushing conduit 1m Doors
Figure 16 Section and plan of ground-floor toilet in the residential quarter of palace of Minos. From Angelakis AN, Koutsoyiannis D, and Tchobanoglous G (2005) Urban wastewater and stormwater technologies in ancient Greece. Water Research 39: 210–220.
(Antoniou, 2010). Lavatories were used throughout the Roman Empire, with a more or less monumental appearance. The reader is referred to Antoniou (2010) for a detailed discussion of ancient Greek lavatories. Toilets during the Roman era can be divided into two groups: public and private. A public toilet was frequently built near to or inside a bath so that it was easily entered from both inside and outside of the bath. The abundance of water that was conducted to the bath could also be used to flush the toilet. Piped water for flushing private toilets seems to have been a rarity. The Romans, however, lacked something similar to our toilet paper. They probably used sponges or moss or something similar. In public toilets, the facilities were common to all. They were cramped, without any privacy, and had no decent way to wash one’s hands. The private toilets most
4.01.8 Discussion and Conclusions In the Minoan, Greek, and Roman cities, and other settlements, water supply varied according to local conditions, determined by climate (mainly rainfall), surface and ground water, and terrain. In these periods, various water-supply and wastewater systems and techniques were developed and applied, such as collection and storage facilities, wells and groundwater abstraction aqueducts, water distribution and use, construction and use of fountains, sewers, bathrooms, and other sanitary facilities and even recreational uses of water. These advanced technologies, which have been used in prehistoric Crete since about 4500 years ago, were subsequently expanded during the Mycenaean and then the Archaic, Classical, and Roman periods. In light of these historical and archaeological evidences, it turns out that the progress of present-day urban water and wastewater technologies as well as comfortable and hygienic living is not as significant as we tend to believe (Angelakis and Koutsoyiannis, 2003). However, a burst of achievements in water and
20
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Figure 17 Public toilet in Ephesus (Turkey): (a) the bench-shaped seats were constructed of stone slabs with another vertical stone slab that covered the opening from the void between the floor and the seat and (b) the small channel (half-pipe-shaped cross-section) on the floor in front of the seat had a continuous flow of water for cleaning the sponghia (the toilet paper of the time). Copyright permission with LW Mays.
wastewater technology was accomplished throughout the centuries of the ancient Greek and Roman civilization. With a few exceptions, the basis for present-day progress in water transfer is clearly not a recent development, but an extension and refinement of the past. In fact, the surprising features are the similarity of ancient water methodologies with those of the present and the advanced level of water and wastewater management used by the ancients. Greek and Roman technological developments in water and wastewater management principles and practices as well as other achievements of those civilizations, such as poetry, philosophy, sciences, politics, and visual arts, are not known. To put in perspective the ancient water and wastewater achievements discussed in this chapter, it is important to examine their relevance to modern times and to harvest some lessons. The relevance of ancient hydraulic works should be examined in terms of the evolution of technology, the technological advances, homeland security, and management principles. The Romans, whose empire replaced the Greek rule in most part of this area, inherited the technologies and developed them further by changing their application scale from small to large and implementing them to almost every large city. The Greek and Roman water technologies are not only a cultural heritage but also the underpinning of modern achievements in water and wastewater engineering and management practices. Apparent characteristics of technologies and management practices in many ancient civilizations are durability and sustainability. Also, there have been integrated management practices, combining both large-scale and small-scale constructions and measures that have allowed cities to sustain for millennia. Currently, engineers use return period for the design of hydraulic structures as dictated by design standards and economic considerations. Sustainability, as a design principle, has
entered the engineering lexicon within the last decade. Naturally, it is difficult to estimate the design principles of ancient engineers but it is notable that several ancient works have operated for very long periods, some until recent times. Thus, wastewater and stormwater drainage systems were functioning in Bronze Age settlements and continued during the Greek and Roman periods. These include the construction and use of bathrooms and other sanitary and purgatory facilities, as well as wastewater and storm sewer systems. In fact, the hydraulic and architectural function of sewer systems in palaces and cities are regarded as one of the salient characteristics of Minoan civilization. They were so advanced that they can be justly compared with their modern counterparts. The durability of some of the constructions that operated up to present times, as well as the support of the technologies and their scientific background by written documents, enabled these technologies to pass to present societies despite regressions that have occurred through the centuries (i.e., in the Dark Ages). The development of science and engineering is not linear but often characterized by discontinuities and regressions. Bridges from the past to the future are always present, albeit oftentimes they are invisible to those who cross them! Thus, in addition to many ancient constructions that have been continuously or intermittently in operation to date, substantial information from ancient Greek and Roman written sources has also been preserved (Angelakis and Koutsoyiannis, 2003). Thus, the major achievements were accomplished during the Greek and Roman civilizations. As a result, they represent the state-of-the-art structures that were technically feasible at that time. For example, the aqueduct of ancient Samos, called ‘&mj´istomon’ or ‘bi-mouthed’ (thus pointing out that it was constructed from two openings), is an important hydraulic monument, indicating that it was possible in the ancient world to design and construct technologically advanced water transportation projects on a large scale.
Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
From the preceding synoptic discussion, certain conclusions might be suggested for further reflection and systematic investigation: 1. The water and wastewater hydraulics works in Minoan, Greek, and Roman civilizations are sometimes not too different from the modern practice, since present technologies descend directly from that time’s engineering. 2. Minoan, Greek, and Roman water and wastewater public works are characterized by simplicity, robustness of operation, and the absence of complex controls. 3. The meaning of sustainability in modern times should be reevaluated in light of Minoan, Greek, and Roman hydraulic works and water and wastewater management practices. 4. Technological developments based on sound engineering principles can have extended useful lives. 5. In areas of water shortage, development of a cost-effective and environmental friendly water resources management practice, based on Minoan, Greek, and Roman civilizations principles, is essential.
References Adam JP (1988) L’Arte di Costruire presso i Romani. Materiali e Tecniche (Roman Building: Materials and Techniques). Milan: Longanesi. Angelakis AN and Koutsoyiannis D (2003) Urban water resources management in ancient Greek times. In: Stewart BA and Howell T (eds.) Encyclopedia of Water Science, pp. 999--1007. New York: Dekker. Angelakis AN, Koutsoyiannis D, and Tchobanoglous G (2005) Urban wastewater and stormwater technologies in ancient Greece. Water Research 39: 210--220. Angelakis AN, Lyrintzis AG, and Spyridakis SV (2010) Urban water management in Minoan Crete, Greece. E-Water (in press). Angelakis AN, Savvakis YM, and Charalampakis G (2007) Aqueducts during the Minoan era. Water Science and Technology: Water Supply 7(1): 95--101. Angelakis AN and Spyridakis SV (1996) The status of water resources in Minoan times – a preliminary study. In: Angelakis A and Issar A (eds.) Diachronic Climatic Impacts on Water Resources with Emphasis on Mediterranean Region, pp. 161–191. Heidelberg: Springer. Angelakis AN and Spyridakis DS (2010). Water supply and wastewater management aspects in ancient Greece. Water Science and Technology: Water Supply 10(4) (in press). Antoniou G, Xarchakou R, and Angelakis AN (2006) Water cistern systems in Greece from Minoan to Hellenistic period. In: Angelakis AD and Koutsoyiannis D (eds.) Proceedings of 1st IWA International Symposium Water and Wastewater Technologies in Ancient Civilizations, pp. 457–462. National Agricultural Research Foundation, Iraklio, Greece, 28–30 October 2006. Antoniou GP (2010) Ancient Greek lavatories: Operation with reused water. In: Mays LW (ed.) Ancient Water Technology. Dordrecht: Springer. Antoniou GP and Angelakis AN (2009) Historical development bathrooms (toilets) and other sanitary and purgatory structures in Greece. In: Proceedings of 2nd IWA International Symposium on Water and Wastewater Technologies in Ancient Technologies. Bari, Italy, 28–29 May 2009. Bersani P, Canalini A, and Dragoni W (2010) First results of a study of the Etruscan tunnel and other hydraulic works on the ‘‘Ponte Coperto’’ stream (Cerveteri, Rome, Italy). Water Science and Technology: Water Supply 10(4) (in press). Bono P and Boni C (1996) Water supply of Rome in antiquity and today. Environmental Geology 27: 126--134. Cadogan G (2007) Water management in Minoan Crete, Greece: The two cisterns of one Middle Bronze Age settlement. Water, Science and Technology: Water Supply 7(1): 103--112. Crouch DP (1993) Water Management in Ancient Greek Cities. New York: Oxford University Press. Darcque P and Treuil R (eds.) (1990) The storm drains of the east wing at Knossos. Special Issue: L’habitat e´ge´en pre´historique. Bulletin de Correspondance Helle´nique, Supple´ment 19: 141–146. Davaras K (1976) Guide to Cretan Antiquities. Park Ridge, NJ: Noyes Press.
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De Feo G, De Gisi S, Malvano C, and De Biase O (2010) The greatest water reservoirs in the ancient Roman world and the ‘‘Piscina Mirabilis’’ in Misenum. Water, Science and Technology: Water Supply 10(4) (in press). De Feo G, De Gisi S, Malvano C, et al. (2010) The Roman aqueduct and the Helvius’ Fountain in Sant’Egidio del Monte Albino, in Southern Italy: A historical and morphological approach. In: Proceedings of 2nd IWA International Symposium on Water and Wastewater Technologies in Ancient Technologies. Bari, Italy, 28–29 May 2009. De Feo G, Malvano C, De Gisi S, and De Biase O (2009b) The ancient aqueduct from Serino to Beneventum in Southern Italy: A technical and historical approach. In: Proceedings of 2nd IWA International Symposium on Water and Wastewater Technologies in Ancient Technologies. Bari, Italy, 28–29 May 2009. De Feo G and Napoli RMA (2007) Historical development of the Augustan aqueduct in Southern Italy: Twenty centuries of works from Serino to Naples. Water Science and Technology: Water Supply 7(1): 131--138. Evans SA (1921–1935) The Palace of Minos at Knossos: A Comparative Account of the Successive Stages of the Early Cretan Civilization as Illustrated by the Discoveries, vols. I–IV, London: Macmillan (reprinted by Biblo and Tannen, New York, USA, 1964). Fahlbusch H (2006) Water management in the classic civilization. In: Proceedings of La Ingenieria Y La Gestion Del Agua a Traves de Los Tiempos. Universidad de Alicante, Spain, with the Universidad Politechnica de Valencia, Alicante, Spain, 30 May–01 June 2006. Gorokhovich Y (2005) Abandonment of Minoan palaces on Crete in relation to the earthquake induced changes in groundwater supply. Journal of Archaeological Science 32: 217--222. Graham JW (1987) The Palaces of Crete. Princeton, NJ: Princeton University Press. Haberey W (1972) Die ro¨mischen Wasserleitungen nach Ko¨ln. Bonn: RheinlandVerlag. Hansen RD (2006) Water and wastewater systems in imperial Rome. http:// www.waterhistory.org (accessed February 2010). Hodge AT (2002) Roman Aqueducts & Water Supply, 2nd edn. London: Gerald Duckworth. Karakostantinou A (2006) The Roman Aqueduct of Moria, Lesvos. Volos, Greece: Department of Elementary Education, University of Thessaly (in Greek). Koutsoyiannis D, Mamassi N, and Tegos A (2007) Logical and illogical exegeses of hydrometeorological phenomena in ancient Greece. Water Science and Technology: Water Supply 7(1): 13--22. Koutsoyiannis D, Zarkadoulas N, Angelakis AN, and Tchobanoglous G (2008) Urban water management in ancient Greece: Legacies and lessons. ASCE, Journal of Water Resources Planning and Management 134(1): 45--54. Ku¨ltu¨r AS¸ (2008) The History of the Basilica Cistern. Istanbul, Turkey. http:// www.yerebatan.com/english/itarihce.html (accessed July 2010). Lendering J (2008) Constantinople (Istanbul): Basilica Cistern. Istanbul, Turkey. http://www.livius.org (accessed July 2010). MacDonald CF and Driessen JM (1988) The drainage system of the domestic quarter in the Palace at Knossos. British School of Athens 83: 235--358. Martini P and Drusiani R (2009) History of the water supply of Rome as a paradigm of water services development in Italic peninsula. In: Proceedings of 2nd IWA International Symposium on Water and Wastewater Technologies in Ancient Technologies. Bari, Italy, 28–29 May 2009. Mavromati E and Chryssaidis L (2007) Aqueducts in the Hellenic area during the Roman Period. Water Science and Technology: Water Supply 7(1): 139--145. Mays LW (2007) Ancient urban water supply systems in arid and semi-arid regions. In: Proceedings of International Symposium on New Directions in Urban Water Management. UNESCO, Paris, France, 12–14 September 2007. Korea Water Resources Association, http://www.kwra.or.kr (accessed February 2010). Mays LW (2008) A very brief history of hydraulic technology during antiquity. Environmental Fluid Mechanics 8(5): 471--484. Mays LW (ed.) (2010) Ancient Water Technologies. Dordrecht: Springer. Mays LW, Koutsoyiannis D, and Angelakis AN (2007) A brief history of urban water supply in antiquity. Water, Science and Technology: Water Supply 7(1): 1--12. Monteleone MC, Yeung H, and Smith R (2007) A review of ancient Roman water supply exploring techniques of pressure reduction. Water Science and Technology: Water Supply 7(1): 113--120. Myers JW, Myers EE, and Cadogan G (1992) The Aerial Atlas of Ancient Crete. Berkeley, CA: University of California Press. Oziz U (1987) Ancient water works in Anatolia. Water Resources Development 3(1): 55--62. Oziz U (1996) Historical water schemes in Turkey. Water Resources Development 12(3): 347--383. Panimolle G (1984) Gli Acquedotti di Roma Antica (The Aqueducts of Ancient Rome). Rome: Edizioni Abete.
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Water and Wastewater Management Technologies in the Ancient Greek and Roman Civilizations
Rodgers RH (2004) Sextus Iulius Frontinus. On the Water-Management of the City of Rome. De Aquaeductu Urbis Romae. Cambridge: Cambridge University Press. Russo F and Russo F (2007) Pompei. La Tecnologia Dimenticata (Pompeii. The Forgotten Technology). Naples: ESA – Edizioni Scientifiche e Artistiche. Tassios TP (2007) Water supply of ancient Greek cities. Water Science and Technology: Water Supply 7(1): 165--191.
Tolle-Kastenbein R (2005) Archeologia dell’Acqua (Water Archaeology). Milan: Longanesi. Vuorinen HS (2010) Water, toilets and public health in the Roman era. Water Science and Technology: Water Supply 10(4) (in press). Vuorinen HS, Juuti PS, and Katko TS (2007) History of water and health from ancient civilizations to modern times. Water Science and Technology: Water Supply 7(1): 49--57.
4.02 Membrane Filtration in Water and Wastewater Treatment Y Watanabe and K Kimura, Hokkaido University, Sapporo, Japan & 2011 Elsevier B.V. All rights reserved.
Membrane Application to Water Purification Current Status Membrane Fouling Main foulant Affinity of main foulant for membranes Membrane Filtration Systems for Controlling Fouling Channel flocculation in monolith ceramic membrane Pre-coagulation/sedimentation in hollow-fiber UF/MF membrane Hybrid submerged MF membrane system PVDF Membrane filtration with pre-ozonation Membrane Application to Wastewater Treatment Current Status of MBRs Mechanism of Membrane Fouling Effect of membrane permeate flux on fouling Effect of membrane material on fouling Fouling potential of carbohydrate assessed by lectin affinity chromatography
4.02.1 Membrane Application to Water Purification 4.02.1.1 Current Status The mainstay of water purification technology in the twentieth century was sand filtration, but since the late 1980s, membrane filtration technology using RO/NF/UF/MF membranes has been applied to the water and wastewater treatment, desalination, and water reuse (RO, reverse osmosis; NF, nanofiltration; UF, ultrafiltration; MF, microfiltration).
3500 Start of RO research in USA (1953)
3000
Water / wastewater treatment (UF/MF)
President J.F.Kenedy approved RO desalination as a national project (1961)
2500 Cryptosporidium infection in Milwaukee (1993)
2000 Enhanced regulations of surface water in USA (1998)
1500 Enhanced water works law in Japan (2001)
1000
Brackish water desalination / wastewater reuse (NF / RO)
2005
2000
1995
1990
1980
1975
1970
1965
1960
0
1955
500
1950
Global accumulative amount of permeate (×104 m3 d−1)
23 23 23 24 30 36 36 40 43 45 47 47 48 49 54 57 60
Figure 1 shows the historical development of membrane technology in the water and wastewater treatment. Membrane filtration has small foot print, extremely high solid–liquid separation ability, and its maintenance is easy. Water purification plants in the United States, the Netherlands, France, Australia, and Japan have introduced the membrane filtration process. Figure 2 shows the recent increase in the amount of water produced by the membrane filtration, which includes water purification, desalination, and wastewater treatment.
1985
4.02.1 4.02.1.1 4.02.1.2 4.02.1.2.1 4.02.1.2.2 4.02.1.3 4.02.1.3.1 4.02.1.3.2 4.02.1.3.3 4.02.1.3.4 4.02.2 4.02.2.1 4.02.2.2 4.02.2.2.1 4.02.2.2.2 4.02.2.2.3 References
Sea water desalination (RO)
“If we could produce fresh water from salt water at a low cost that would indeed be a great service to humanity, and would dwarf any other scientific accomplishment” John F. Kennedy Figure 1 Development of membrane filtration. MF, microfiltration; NF, nanofiltration; RO, reverese osmosis; UF, ultrafiltration.
23
24
Membrane Filtration in Water and Wastewater Treatment Global amount of water produced by membrane processes 35 000 000 32 000 000 m3 d–1, 2006 SWRO
Amount of water (m3 d–1)
30 000 000
Increase by 25% each year
NF+BWRO 25 000 000
LP+MF+UF
20 000 000 15 000 000 10 000 000
2006
2005
2004
2003
2002
2001
2000
1999
1998
1997
1996
1995
1994
1993
1992
1991
0
1990
5 000 000
Figure 2 Increase in purified water by membrane filtration. BWRO, brackish water reverese osmosis; LP, low pressure; MF, microfiltration; NF, nanofiltration; SWRO, seawater reverese osmosis; UF, ultrafiltration.
Table 1
Large-scale water purification plants in world wide
Country
Place (plant name)
Capacity (103m3d1)
Construction year
Membrane
Water source
USA Canada Singapore USA USA USA Canada UK Germany USA
Minneapolis (Fridley Plant) Mississanga, Ontario Chestnut Minneapolis (Columbia Heights) Racine, Wisconsin Thornton, Colorado Kamloops, British Columbia Clay Lane Roetgen/Aachen San Joaquin, California
360 302 273 265 189 187.5 160 160 144 136
2011 (to be built) 2006 2003 2005 2005 2005 2005 2001 2005 2005
UF UF UF UF UF UF UF UF UF UF
Surface Lake Surface Surface Surface Surface Surface Ground Reservior Surface
Source: Japan Water Research Center, Hot News in water works, No. 56.
Table 1 shows the large-scale water purification plants using membrane filtration. All plants in the table use the UF membrane but a plant using monolith ceramic MF membrane with the capacity of 173 000 m3 d1 is under construction in Japan. There has been a significant progress in the development of new robust MF membranes with new polymers such as PVDE and FTFE for water and wastewater treatment. Combining robust MF membranes and the other processes such as coagulation, ozonation, biological/chemical oxidation, and powdered activated carbon adsorption and chemically enhanced physical cleaning makes very efficient water purification system. They are very effective in the application to the large-scale water purification plant. The trend toward membrane filtration is expected to spread worldwide during this century. However, there are several limiting factors applying the UF membrane and MF membrane to the water purification. Among them, fouling in membrane is a major obstacle to widespread use of this technology. The authors have been studying the mechanism and control of membrane fouling in water treatment. This chapter
summarizes the authors’ research on membrane application to the water purification.
4.02.1.2 Membrane Fouling Several physical membrane cleaning methods such as hydraulic backwashing and air scrubbing have been developed and used routinely in many existing membrane plants to minimize membrane fouling. Despite routine physical membrane cleaning, membrane filtration resistance gradually increases over a long period of operation, indicating that membrane fouling cannot be completely controlled by physical cleaning. Fouling that cannot be controlled by physical cleaning is defined here as physically irreversible fouling. Control of physically irreversible fouling is important for the reduction of operation cost in a membrane process because this type of fouling develops even when a very efficient physical cleaning is carried out. Physically irreversible membrane fouling can only be canceled by chemical cleaning. However, chemical cleaning of the membrane should be limited to a minimum frequency because repeated chemical
Membrane Filtration in Water and Wastewater Treatment
cleaning may shorten the membrane lifetime and disposal of spent chemical reagents poses another problem. Membrane fouling strongly depends upon the structure of membrane (average size, size distribution, and density of pores). Surface morphology and roughness are surely involved in it. However, this chapter describes the effect of only nominal pore size and materials of membrane on the membrane fouling.
4.02.1.2.1 Main foulant In a number of previous studies on fouling of membranes used for water treatment, natural organic matter (NOM), composed of a variety of nonbiodegradable organic compounds including humic substances, has been shown to be the major constituent causing membrane fouling. However, it is still not clear which fraction of NOM causes membrane fouling. In early works, hydrophobic fractions of NOM, such as humic substances, were considered to be the major foulants. Hydrophobic interaction and electrostatic interaction were the explanations for the binding between hydrophobic NOM and membranes. More recently, hydrophilic NOM with features of carbohydrate or protein has been reported by several researchers to be the major foulant. As explanations for the binding between hydrophilic NOM and membranes, van der Waals attraction and hydrophobic interaction between membranes and hydrophobic domains in hydrophilic NOM have been suggested. In addition to NOM, metals and metal– NOM complexes have been reported as the constituents affecting membrane fouling (Yamamura et al., 2007a, 2007b). Physically reversible fouling and physically irreversible fouling have not been distinguished in many previous studies. In addition, many previous studies were based on short-term experiments, which are not sufficient for observing physically irreversible fouling. As a result, knowledge of physically irreversible fouling occurring in membrane filtration in drinking water treatment is very limited; therefore, further studies need to be carried out with special emphasis on physically irreversible fouling for more efficient use of membranes. In particular, investigation of the characteristics of components that cause physically irreversible fouling would be useful for the establishment of a new protocol of fouling control. In this study, three MF/UF membranes that had been fouled in long-term filtration of surface water used as a drinking water source were investigated in terms of the recovery of water permeability by chemical cleaning and the characteristics of the foulant causing physically irreversible fouling. Based on the results obtained from various analyses, a hypothesis regarding the evolution of physically irreversible fouling is proposed. Three different hollow-fiber membranes were used in this study. Two of them were MF membranes and the other was a UF membrane. The two MF membranes had the same nominal pore size of 0.1 mm but were made from different polymers such as polyethylene (PE; Mitsubishi Rayon, Tokyo, Japan) and polyvinylidene fluoride (PVDF; Asahikasei Chemicals, Tokyo, Japan). The UF membrane had a molecular weight cut-off of 100 000 Da and was made from polyacrylonitrile (PAN; Toray Industries, Tokyo, Japan). Using these three different membranes, pilot-scale membrane
25
filtration tests were carried out in parallel using the Chitose River surface water. This river flows through peat area and its surface water contains many humic substances. The concentration range of total iron and aluminum was 0.7–1.7 and 0.05 and 0.7 mg l1. About 75% of them were larger than 0.45 mm. The PVDF and the PE membranes were submerged in separate tanks and were operated under vacuum. The PAN membrane was housed in a vessel and was operated under pressure. All membranes were operated in the outside-in flow mode. The three membranes were operated with identical run cycles (filtration: 30 min; air scrubbing: 30 s; hydraulic backwashing: 60 s) at the same constant flux of 0.65 m3 m2 d1. Hydraulic backwashing was not accompanied by the addition of chlorine. When membrane fouling became significant in the submerged MF membranes despite the implementation of periodical backwashing, membrane modules were taken out from the tanks and were cleaned by spraying pressurized water on the membrane surface. The average quality of the feed water and that of membrane permeates are shown in Table 2. In the feed water, large portions of aluminum (78%) and iron (75%) were present as suspended solids (40.45 mm), while manganese, calcium, and organic matter were mainly present in dissolved forms. Aluminum and iron were effectively removed by the tested membranes due to the strict solid–liquid separation. On the other hand, removal of manganese, calcium, and organic matter was not significant in any of the membranes. This implies that the sizes of manganese, calcium, and dissolved organic carbon (DOC) were smaller than the pore sizes of the tested membranes. The UF membrane showed slightly higher rates of removal of DOC and UV absorbance than those of the two MF membranes, reflecting the difference between membrane pore sizes of the MF and UF membranes. However, the concentration of aluminum in the PAN membrane was slightly higher than the concentrations in the MF membranes. No reasonable explanation for this is available at present. Figure 3 shows the changes in transmembrane pressure (TMP) in the three membranes. The rates of increase in TMP in the three membranes were considerably different. As expected, the tightest membrane (PAN) showed the highest rate of increase in TMP. The rates of increase in the two MF membranes were different despite the fact that they had the same nominal pore size. This clearly indicates that the materials of the membrane have a substantial influence on the
Table 2
Average raw water quality during experiment
Temperature (1C) pH Turbidity (NTU) UV absorbance at 220 nm (cm1) UV absorbance at 260 nm (cm1) TOC (mg 11) DOC (mg 11) THMFP (mg 11) Manganese (mg 11) Soluble manganese (mg 11) Ammonia Nitrogen (mg 11)
11.5 7.11 16.54 0.411 0.099 2.43 2.29 0.086 0.100 0.074 0.22
DOC, dissolved organic matter; THMFP, trihalomethane formation potential; TOC, total organic carbon.
26
Membrane Filtration in Water and Wastewater Treatment 200
PAN
Time of additional physical cleaning
PVDF PE
TMP (kPa)
160
120
80
40
0
0
10
20
30
40
50
Operation time (days) Figure 3 Time course changes in transmembrane pressure (TMP) difference adjusted to 20 1C equivalent value considering the change in water viscosity. PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride; TMP, transmembrane pressure.
evolution of membrane fouling. Interestingly, the results obtained in this study showing that the PE membrane was less fouled than the PVDF membrane are opposite to the results of a previous study focusing on membrane fouling in membrane bioreactors (MBRs) used for municipal wastewater treatment. This implies that characteristics of foulants in the case of drinking water treatment were different from those in the case of wastewater treatment. Further investigation is needed to understand the influence of membrane material on the rate of fouling. In all of the tested membranes, increase in TMP was not constant and rapid increases in TMP were seen several times. After the rapid increases in TMP, however, the value of TMP gradually declined due to the periodical backwashing except for the case of the PVDF membrane. On days 31 and 41, an additional physical cleaning (spraying pressurized water on the membrane surface) was needed to maintain the permeability of the PVDF membrane. This additional physical cleaning worked well and substantial reduction in TMP in the PVDF membrane was seen after cleaning. Chemical cleaning was not carried out at that time. Based on the observations mentioned above, it is assumed that the rapid increases in TMP shown in Figure 3 were caused by the accumulation of cake on the surfaces of the membranes. The three dashed lines shown in the figure are assumed to represent the evolution of physically irreversible fouling in the three membranes, which accumulated and remained despite of the implementation of periodical backwashing and additional physical cleaning. As seen in Figure 3, the rates of occurrence of physically irreversible fouling in the three membranes were different. To investigate the features of constituents that were responsible for physically irreversible fouling, the foulants were desorbed from the fouled membranes at the termination of the operation and then their chemical characteristics were analyzed. When the pilot operations were terminated, fouled membranes were taken out from the filtration units. The
membrane fibers were immediately brought to the laboratory in a container filled with distilled water. First, each membrane fiber was manually wiped with a sponge and thoroughly rinsed with distilled water, which was carried out to minimize the influence of the accumulated cake causing physically reversible fouling in subsequent tests. By visual inspection, no accumulated cake was found on the membrane after wiping with a sponge. Using the wiped membranes, tiny membrane modules of 40 cm2 in membrane area were assembled and pure water permeability of the fouled membrane was measured by applying 30 kPa of pressure difference. Filtration was continued until a constant permeate flow rate was achieved (typically in 15 min). After measuring the pure water permeability, tiny membrane modules were soaked in various chemical solutions at 20 1C for 24 h. The chemical solutions used for cleaning were Milli-Q water, NaClO (700 ppm as free available chlorine), NaCl (0.1 M), NaOH (pH 12), HCl (pH 2), ethylenediaminetetraacetic acid(EDTA) (20 mM), and oxalic acid (0.5%). Recoveries in pure water permeability by the chemical cleaning were evaluated and the chemical solutions containing the foulants desorbed from the membranes were analyzed. Membrane specimens that were not used for assembling the tiny membrane modules were divided into two portions and were soaked in a solution of sodium hydroxide at pH 12 or hydrochloric acid at pH 2. Because a large amount of membrane specimens was available in this study, this process enabled extraction of a sufficient amount of organic matter for advanced analysis (e.g., Fourier transform infrared (FTIR) and nuclear magnetic resonance (NMR) spectra). Figure 4 shows the degree of restoration of the fouled membranes in terms of pure water flux by chemical cleaning with various reagents. In this figure, the ratio of pure water flux after chemical cleaning (J1) to the flux before chemical cleaning (J0) is used to express the degree of flux restoration. As described earlier, chemical cleaning was carried out after
Membrane Filtration in Water and Wastewater Treatment PVDF
PE 7.0
NaCIO Oxalic
PAN
NaCIO
NaCIO
Oxalic
Oxalic
HCl
HCl
HCl
EDTA
EDTA
EDTA
NaOH
NaOH
NaOH
NaCl
NaCl
NaCl
MQ
MQ
MQ
1
2 J1/ J0
3
1
27
2 J1/ J0
3
1
2 J1/ J0
3
Figure 4 Effect of chemical membrane cleaning (J0: pure water flux before chemical cleaning, J1: pure water flux after chemical cleaning). EDTA, ethylenediaminetetraacetic acid; MQ, milli-Q water; PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride.
manually removing reversible cake that had accumulated on the membrane. Therefore, it can be considered that the restoration shown in Figure 4 represents removal of the foulants causing physically irreversible membrane fouling. Actually, manual sponge cleaning carried out prior to chemical cleaning had little effect on the permeability of the fouled membranes, indicating that fouling seen at the termination of the longterm operation could be attributed mainly to physically irreversible fouling. As seen in the figure, in the case of the PVDF and PAN membranes, NaCl (0.1 M) and EDTA (20 mM) were not effective in mitigation of physically irreversible fouling in this study. Figure 4 also shows that alkaline solution (NaOH) was more efficient than acid solutions (oxalic acid and HCl) for recovery of permeability of the PVDF and PAN membranes. The oxidizing agent (NaClO) exhibited the best cleaning performance in recovery of permeability of the PVDF and PAN membranes. This implies that organic matter was mainly responsible for the evolution of physically irreversible membrane fouling in the PVDF and PAN membranes. In contrast, in the case of the PE membrane, which exhibited the least membrane fouling in the continuous run (Figure 3), the degree of recovery of water permeability following cleaning with acid, alkaline, and oxidizing reagents were comparable. This suggests that the contribution of metals to the physically irreversible fouling in the PE membrane was significant. Desorption of membrane foulants was carried out at the termination of the pilot operation. As stated above, to ensure that physically reversible cake was removed from the membrane surface, each membrane fiber was carefully wiped with a sponge prior to desorption tests. Although both aluminum and iron in the raw water were effectively removed by the membranes tested, only iron was desorbed from the fouled membranes at a significant amount. This suggests that aluminum in the feed water was rejected or deposited on the membrane surface and subsequently removed by the periodical backwashing. In contrast, iron was likely to cause the physically irreversible fouling to some extent. In the cleaning with HCl solution, not only metals but also organic matter were desorbed from the fouled membranes, particularly from the PVDF membrane. Figure 5 shows the FTIR spectra of the foulants desorbed from the fouled membranes by HCl solution. Interestingly, there were significant similarities among the three spectra. All of
1080
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Wave number (cm–1) Figure 5 FTIR spectra of membrane foulant desorbed with HCl (pH 2) solution. PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride.
the spectra had a dominant peak near 1080 cm1, which is an indication of their carbohydrate character. Therefore, the carbohydrate-like organic matter was thought to be the main constituent in the foulants desorbed with HCl solution regardless of membrane type. In a study by Kabsch-Korbutowicz et al., it was shown that a large portion of organic matter desorbed from the fouled membrane by acid or chelating agents formed complexes with metals. Similarly, in the present study, the carbohydrate-like organic matter and metals (mainly iron) desorbed with HCl solution were assumed to form complexes and cause physically irreversible fouling. It has been reported that carbohydrate can form a complex with iron. As previously mentioned, NaOH solution restored the membrane permeability to a larger extent and desorbed a larger amount of organic matter from the fouled membranes than did HCl solution. Therefore, analysis of the foulants desorbed from the membrane with NaOH solution would be more useful in understanding the fouling, compared to the case of HCl solution. The value of specific ultraviolet absorbance (SUVA) is considered to be a surrogate measurement of aromacity of organic matter, and a high SUVA value corresponds to organic matter consisting of a large amount of double-bond or aromatic structures. The values of SUVA determined for the foulants desorbed by NaOH solution were much lower than those for the feed water on average. This
28
Membrane Filtration in Water and Wastewater Treatment 1080 1660
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Figure 6 Infrared spectra of membrane foulant desorbed with NaOH (pH 12) solution. PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride.
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implies that a relatively hydrophilic fraction of the organic matter in the feed water was responsible for the physically irreversible fouling. Interestingly, the value of SUVA determined for the foulants was similar among the foulants desorbed from the three membranes. This indicates that the characteristics of the foulants desorbed from the three membranes might be similar, but this turned out to be false as discussed later. FTIR spectra of the foulants desorbed with NaOH solution from the three membranes are presented in Figure 6. There were significant similarities in the spectra obtained for the three membranes. In these spectra, peaks near 1660 and 1540 cm1 were significant. They are assigned to amido-I and II bands, respectively. In all spectra, a broad peak near 1080 cm1 was seen. This peak is an indicator of carbohydrate character. FTIR spectra shown in Figure 6 are not similar to those of humic substances. This suggests that humic substances were relatively minor components in the foulant responsible for the physically irreversible fouling. CPMAS 13C NMR spectra of the foulants desorbed with NaOH solution from the membranes are presented in Figure 7. A general similarity among the foulants desorbed from the three membranes was found in NMR analysis as well. Although a proteinaceous nature of the foulants in the membranes can be seen by peaks near 175 and 55 ppm, carbohydrate (peak at 75 ppm) was dominant in the foulant regardless of the membrane type. The aromatic carbon signal (110–165 ppm) was minor in the spectra for the two MF membranes (PVDF and PE) but was pronounced in the spectrum for the PAN membrane. This indicates that the contribution of the humic fraction of NOM to the evolution of physically irreversible fouling was more significant in the PAN membrane than in the two MF membranes. The humic fraction would be smaller than carbohydrate, as shown later. Thus, it is reasonable to assume that the contribution of the small humic fraction would become more significant in a UF membrane (PAN in this case) than in MF membrane (PVDF and PE in this case). The amount of calcium desorbed with NaOH solution was significant in the case of the PAN membrane. This calcium might have formed a complex with humic substance as suggested by several researchers. Nevertheless,
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120 80 Chemical shift (ppm)
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Figure 7 CPMAS 13C NMR spectra of membrane foulants desorbed with NaOH (pH 12) solution. PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride.
carbohydrate was dominant in the foulant desorbed form the PAN membrane as well, as shown in Figure 5. As shown above, both FTIR and NMR analyses demonstrated that carbohydrate was a dominant component causing physically irreversible fouling regardless of the type of membrane. Carbohydrate has, however, a hydrophilic nature, and hydrophobic interaction between the membranes and carbohydrate is therefore not a reasonable explanation for the participation of carbohydrate in physically irreversible fouling. To elucidate the fouling mechanisms involved in the continuous operation, changes in rejection rate of both humic acid and carbohydrate in the operation were investigated using HPLC-SEC with UV/DOC detectors. Figure 8 shows the representative molecular weight distribution of organic matter contained in the feed water used in this study. As seen in the figure, organic matter contained in the feed water could be roughly divided into two fractions: large molecules with a hydrophilic nature (little UV absorbance) and small molecules with a hydrophobic nature (high UV absorbance). A similar molecular weight distribution of organic matter was found in previous studies. It is thought that large molecules mainly consisted of carbohydrate, while small molecules mainly consisted of humic acid. Figure 9 shows changes in the removal of the large and small molecules by the three membranes determined by HPLC-SEC with UV/DOC detectors. In the case of the PVDF membrane, about 15% of the fraction of smaller organic molecules mainly composed of humic substances was initially removed. As the operation period became longer, however, the rate of removal of the small organic molecules declined and eventually no removal of small molecules was achieved by the PVDF membrane. The size of the small molecules should be considerably smaller than the nominal pore size of the PVDF membrane (0.1 mm); therefore, the sieving effect was discounted as an explanation for the initial removal of small organic molecules by the PVDF membrane. Rather, the initial
Membrane Filtration in Water and Wastewater Treatment
29
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Figure 8 Molecular size distribution of dissolved organic matter in the Chitose river surface water. DOC, dissolved organic carbon.
removal of the small molecules can be attributed to adsorption on/in the PVDF membrane. In contrast to the small molecules, the rate of removal of the large organic molecules by the PVDF membrane gradually increased during the operation. When the removal of the small molecules declined to a negligible level, the removal of large organic molecules increased by almost 100%. A similar trend was also seen for the other two membranes. Based on these observations, the following hypothesis regarding the evolution of physically irreversible fouling is presented. First, small molecules mainly composed of humic substances are adsorbed on/in membranes by hydrophobic interaction. As a result of adsorption of the small molecules, the sizes of membrane pores decrease and it becomes possible for large molecules mainly composed of carbohydrates to plug the pores and cause physically irreversible fouling. Also, adsorbed humic substances could work as glue for carbohydrates and facilitate the capture of carbohydrates on/in membranes. The examined PVDF was assumed to be more hydrophobic than the PE membrane because hydrophilic modification was provided for the PE membrane by the manufacture. It is likely that the hydrophobic PVDF membrane adsorbed humic substances more rapidly than did the hydrophilic PE membrane. As a result, the PVDF membrane should achieve complete rejection of carbohydrates earlier than the PE membrane (Figure 9). In discussion made above, it is assumed that foulant causing physically irreversible fouling originated from the feed water. Another possible origin of the foulant might be biofilms that cannot be removed by backwashing. It was reported that both carbohydrate and humics were excreted by microorganisms. Although the possibility that excretion from biofilms was the main source of the foulant which cannot be completely eliminated, it would be discounted by the following reasons: (1) evolution of reversible fouling (indication of biofilm formation) did not always dominate in the operation of the membranes as shown in Figure 3; (2) occasional increases in physically reversible fouling shown in Figure 3
could be explained by increases in turbidity in the feed (data not shown); and (3) water temperature was low (i.e., 5–10 1C) in the operation. To deal with the issues discussed above more precisely, establishment of the methods that can distinguish the origin of organic matter is indispensable. The following points were derived from the measurement of the zeta potentials of membranes before and after the longterm operation. The decrease in rejection of small molecules during the operation might be attributable to a decrease in favorable electrostatic interaction (repulsion) since the zeta potential of the tested membranes became slightly less negative after operation as a consequence of carbohydrate deposition. In this study, it was assumed that the decrease in favorable electrostatic interaction was not the main reason for the decrease in rejection of small molecules both because of the initial zeta potential that was close to neutral and because of the small changes in the zeta potentials after use. However, further investigation is needed to determine the influence of surface conditions of membranes on binding of NOM to membranes. To confirm the experimental results showing that carbohydrate-like substances are main substances causing the physically irreversible fouling, the authors carried out the bench-scale study where the surface water samples taken from four different sources such as Toyohira River (central Hokkaido), Kusiro River (eastern Hokkaido), Inba Lake (Chiba prefecture), and Yodo River (Osaka prefecture). Toyohira River water (total organic carbon, TOC ¼ 0.8 mg l1) is relatively clean. Kushiro river water (TOC ¼ 0.9 mg l1) is rich in humic substances. Inba Lake water (TOC ¼ 5.7 mg l1) is polluted and eutrophicated by the domestic wastewater. Yodo River water (TOC ¼ 1.8 mg l1) contained a lot of treated wastewater. Tiny membrane module with the surface area of 1.44 103 m2 was prepared with hollow-fiber membranes made of PVDF. The pore size of membranes was 0.1 mm. Membrane filtration was carried out by a peristaltic pump, and the constant-flow-rate mode of operation was applied. Permeate flux was fixed at 1.5 m d1 for all filtration experiments.
30
Membrane Filtration in Water and Wastewater Treatment PVDF 30
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Figure 9 Changes in removal rate of large molecules (carbohydrate) and small molecules (humic acid). PAN, polyacrylonitrile; PE, polyethylene; PVDF, polyvinylidene fluoride.
Hydraulic backwashing was performed every 15 min. The duration and pressure of backwashing were 30 sec and 50 kPa, respectively. The organic matter in the four water sources was concentrated using RO (Nanomax 95, Millipore), and its recovery, defined as (DOC mass after concentration by RO)/(DOC mass before concentration), was 0.95, 0.81, 0.80, and 0.91 for Toyohira River, Inba Lake, Kushiro River, and Yodo River, respectively. Fractionation of organic matter contained in the isolates was carried out using the procedure described by Croue et al. They used the DAX-8 and XAD-4 resins. The portion that passed through both the DAX-8 and XAD-4 column was denoted the hydrophilic (HPI) fraction. The portion that retained on DAX-8 resin was denoted the hydrophobic (HPO) fraction. The portion that retained on XAD-84 was denoted the transphilic (TPI) fractions. The HPO and TPI fractions were eluted by backwashing with 2 l of 0.1 N NaOH at 100 ml min1. Each of the three fractions was desalinated by the electric dialysis until its electric conductivity became less
than 0.5 mS O1. The HPI and HPO fractions were diluted to a concentration of 2.0 mg TOC l1 with Milli-Q water and used as the feed water for the bench scale experiment. Figure 10 shows the FTIR spectra of the organic matter in the hydrophobic and hydrophilic fractions of the water from each of the four sources. FTIR analysis is a powerful tool for identifying the functional groups in organic matter and, together with the SUVA, provides useful information about the characteristics of organic matter in the feed waters. As seen in the spectra of the HPO fractions, the organic matter in the hydrophobic fraction was highly aromatic. For all the spectra of HPO fraction, a general similarly was seen in two broad peaks around 1400 and 1620–1660 cm1. These peaks are an indication of their aromatic character. The HPO fractions also seemed to contain alkyl aromatic sulfonates, as evidenced by the peaks of an aromatic sulfonic acid group (1035 and 1009 cm1) and the alkyl group (2930 cm1). In the spectra of the HPI fractions, on the other hand, a high peak at 1080 cm1 is seen for all the sources. This peak is assigned to C–O stretching of polysaccharide or aliphatic alcohol, which represent the carbohydrate-rich nature of HPI organic matters. The spectra of the HPI fractions of Inba Lake water and Yodo River water not only show the signature of carbohydrate-like substances but also have sharp peaks at 1620 and 1660 cm1 corresponding to carboxylic acid. These peaks, in combination with the peak at 1080 cm1, might indicate the presence of alginate-like substances in the feed water. The changes in TMP during filtration through the MF membrane made of PVDF differed between the HPO fraction and the HPI fraction are shown in Figure 11. Regardless of the NOM source, the TMP for the HPO fraction increased by less than 7 kPa and the TMP for the HPI fraction increased by more than 30 kPa. This clearly indicates that the HPI fraction of NOM is a major component affecting the development of physically irreversible fouling. The major differences in the characteristics of organic matter between the HPO fraction and the HPI fraction are in aromaticity and size. The organic matter in the HPO fraction consisted mainly of aromatic humic substances less than 6000 Da in size, while the HPI fraction was rich in carbohydrate-like substances having sizes between 100 000 and 1000 000 Da. These findings indicate that the development of physically irreversible fouling was caused not by aromatic humic substances but by carbohydrate-like substances. In authors’ study investigating the affinity between NOM and membrane surfaces, it was concluded that the physico-chemical interaction with the surface of membrane was more significant for carbohydrate-like substances (with hydroxyl groups) than for humic-like substances (with carboxyl groups). As a consequence, it can be hypothesized that large carbohydrate-like substances can accumulate on the membrane surface, interact with it strongly, and thereby cause physically irreversible fouling. Although some researchers suggested that physically reversible fouling is largely due to the HPO fraction of NOM, the development of physically reversible fouling was not obvious for the HPO feed waters, probably because the organic particles in the HPO fraction are smaller than the membrane pores in this study. Rather, some of the HPI fractions were found to contribute to the physically reversible fouling as well.
Membrane Filtration in Water and Wastewater Treatment Hydrophobic
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Figure 10 Fourier transform infrared (FTIR) spectra of the natural organic matter (NOM) in hydrophobic (HPO) and hydrophilic (HPI) fractions of raw water from different sources: HPO fraction of water from (a) Toyohira river, (b) Lake Inbanuma, (c) Kushiro river, and (d) Yodo river; HPI fraction of water from (e) Toyohira river, (f) Lake Inbanuma, (g) Kushiro river, and (h) Yodo river.
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32
Membrane Filtration in Water and Wastewater Treatment
In particular, the filtration of the HPI fractions of Lake Inbanuma water and Yodo River water induced the evolution of the physically reversible fouling to a large extent. These HPI fractions were found to contain a large amount of macromolecular polysaccharides with a negative charge at neutral pH, in which the electrostatic repulsion between the negatively charged polysaccharide and accumulated polysaccharides or membrane surface would occur. Such an electrostatic repulsion would help to weaken the binding of organic molecules to each other and thereby enable the accumulated organic matter to be easily removed by physical cleaning.
4.02.1.2.2 Affinity of main foulant for membranes In our previous study on pilot-scale filtration using hydrophilic and hydrophobic membranes, NMR analysis of the foulant demonstrated significant contribution of carbohydrate-like substances to the evolution of fouling. It was also shown that the nature of membrane materials affected the rate of accumulation of carbohydrate-like substances. However, the reason for the preferential binding of carbohydrate-like substances to membranes remains unclear. Elucidation of the physicochemical interactions between membranes and carbohydratelike substances is needed for understanding the mechanism of fouling involving carbohydrate. Several research groups have already demonstrated the usefulness of atomic force microscopy (AFM) force measurement for the quantification of the affinity between a carboxylmodified microsphere and the surfaces of NF/RO membranes. Carboxyl-modified microspheres were used as a surrogate of humic substances in their studies. Taking into account the hydroxyl-rich characteristics of carbohydrate, AFM force measurement using hydroxyl-modified microspheres and membranes would provide useful information about the affinity of carbohydrate-like substances to membranes, which has been reported in recent studies on fouling as reviewed above. Two MF membranes with the same nominal pore size of 0.1 mm were used in this study. One membrane was made of PE (Mitsubishi Rayon Engineering, Tokyo, Japan) and the other was made of PVDF (Asahi Kasei Chemicals, Tokyo, Japan). These two membranes were chosen because they are now used in many full-scale plants. Prior to the AFM force measurement, new membranes were filtered with Milli-Q water for 6 h so as to wash out impurities remaining on the membrane surface. Because of hydroxyl-rich nature of carbohydrate, Polybeads-hydroxylate microspheres (Cosmo Bio, Tokyo, Japan) were used as surrogates for carbohydrate-like substances. For comparison, Polybeads-carboxylate microspheres (Cosmo Bio, Tokyo, Japan) were also used in the AFM force measurement. In previous studies, carboxyl-modified microspheres were used as surrogates of humic substances. Both microspheres used in this study were made of polystyrene (3 mm). The characteristics of these microspheres are shown in Yamamura et al. (2008). The colloidal probes used in the AFM force measurement were prepared by attaching the microspheres to the top of a silicon nitride tip (NP-S: Veeco Instruments Inc., New York, USA) as previously described (Figure 12). Attachment of the microspheres to the cantilever tip was carried out with a micromanipulator with the aid of a
Figure 12 Scanning electron microscope image of a polystyrene bead (3 mm) glued to the top of a cantilever tip.
scanning electron microscope (TINY SEM, Technex Lab, Tokyo, Japan). After preparation, the colloidal probes were stored in a refrigerator (4 1C) prior to use. The spring constants of carboxyl- and hydroxyl-colloidal probes determined by thermal fluctuation method were 84 and 92 pN nm1, respectively. These values were used for converting cantilever deflections to loading forces. An atomic force microscope (MFP-3D, Asylum Research, Santa Barbara, CA) was used for the force measurements. Measurements were carried out in buffered water (1.0 mM NaHCO3, pH 6.8) with a trigger point of 50 nm. Divalent cations such as calcium or magnesium were not added to the buffered solution so as to prevent the formation of a bridging between polystyrene of microspheres and the membrane surface. Taking the heterogeneities of local membrane surfaces into account, measurements of force curves were made at three different locations. At each location, more than five force curves were obtained. All force curves obtained by the AFM force measurement were originally expressed as a function of force determined on the basis of the scanner position in the AFM instrument. The scanner position was converted to the separation distance by determining the onset of constant compliance between the scanner position and cantilever deflection (i.e., where cantilever deflection becomes a linear function of piezo-scanner position) and subtracting this value from all other scanner position values. In AFM force curves, the separation distance at which the interaction became either repulsive or attractive was identified as the point where the measured force is either positive or negative, respectively. At separation distances greater than this value, no force was considered to be acting on the colloidal probe and the zero force region of the plot was determined. An AFM force measurement gave two force curves: an approaching force curve and a retraction force curve. The affinity of the colloidal probe to the surface of the membrane was expressed by the adhesion force, Fad, which is defined as the force needed to separate the two from contact. Fad is determined on the basis of the maximum value of cantilever deflection in a retraction force curve (dmax) as shown in
Membrane Filtration in Water and Wastewater Treatment
33
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Figure 13 A representative retraction force curve.
Figure 14 Adhesion forces of (a) carboxyl-modified and (b) hydroxylmodified microspheres to polyethylene (PE) and polyvinylidene fluoride (PVDF) membrane surfaces in buffered solution (pH 6.8).
Figure 13. In contrast, interaction between the colloidal probe and the membrane surface when the probe was approaching the membrane surface (similar to the situation in which carbohydrate-like substances approach membranes by convection flow) was also assessed by the effective distance of the forces shown in an approaching force curve. The affinity between a carbohydrate-like substance and membrane surface would change as a result of fouling. Therefore, AFM force measurement was also carried out with membranes previously fouled in a pilot operation to investigate the change in affinity. Because of the difficulty in regular sampling of membrane specimens from the PVDF membrane module used in the pilot study, the investigation of change in affinity of the carbohydrate-like substance was carried out only with the PE membrane. Pilot-scale membrane filtration was carried out at the Kamiebetsu water purification plant (Ebetsu, Japan) using Chitose River surface water as raw water. Characteristics of the raw water used for the pilot operation are described elsewhere. In authors’ previous study using the same water, it was found that carbohydrate-like substance was dominant in the foulant causing physically irreversible fouling. After passage of the grit chamber, the raw water was delivered to the membrane units without any pretreatment. The PE membrane, which had the same properties as those described before, was assembled (3 m2) and horizontally immersed in a 300 l submersion tank. The operation was conducted using a vacuum. The filtration flux was set at a constant value of 1.0 m3 m2 d1. During the operation that continued for 49 days, periodic physical cleaning was carried out by filtration for 30 min, air scrubbing for 30 s, and hydraulic backwashing for 60 s, as recommended by the manufacturer. When the membrane was rapidly fouled or the value of TMP became excessive, the submerged membrane module was taken out from the submersion tank and was cleaned by spraying pressurized water on the membrane surface. During the pilot-scale operation, membrane fibers were sampled from the center of the membrane module six times: on days 1, 3, 5, 16, 23, and 39. After cutting the fibers, corresponding channels were closed with epoxy glue to prevent leakage, and the permeate flow rate was adjusted to maintain a constant flux of 1.0 m3 m2 d1. To check for membrane breakage, turbidity of the permeate was monitored. After
membrane fibers had been cut, they were immediately brought to the laboratory in a container filled with distilled water (resulting pH of 6.570.5), and the surface of the membrane specimen was manually wiped with a sponge and rinsed with distilled water thoroughly. This step was carried out to ensure the removal of the accumulated cake (i.e., effect of physically reversible fouling) and to specifically focus on physically irreversible fouling in this study. It was found that manual sponge cleaning had little effect on permeability of the fouled membranes at the termination of the operation, indicating that physically irreversible fouling was dominant in the pilot operation. A portion of membrane fibers was examined in a zeta potential meter (ELS-8000, Otsuka Electronics, Osaka, Japan) at pH 7.0 and 5 mM KCl. The other membrane fibers were stored in Milli-Q water until use for AFM force curve measurements. Figure 14 shows the adhesion forces (Fad) of (a) carboxylmodified and (b) hydroxyl-modified microspheres to clean PVDF or PE membranes, which were determined from the maximum values in the retraction force curves (Figure 13). From Figure 14, it is obvious that the adhesion force of the hydroxyl group was much greater than that of the carboxyl group regardless of membrane. The difference in values shown in Figure 14 is explained by differences in a balance of three relevant forces: (1) electrostatic interaction, (2) hydrogen bond (or electron transfer interaction), and (3) van der Waals interaction as seen in Figure 15. The hydrogen bond and the van der Waals interaction work as attraction forces, while the electrostatic interaction is considered to be repelling force because of the negatively charged nature of both microspheres and membrane surfaces. The electrostatic repulsion is governed by Coulomb’s force, which is proportional to the product of the two different charges to be considered. The charges of the two functionally modified microspheres were comparable. Thus, similar levels of Coulomb’s force would be exerted on carboxyl- and hydroxyl-modified microspheres with the membranes. On the other hand, the van der Waals interaction between a microsphere and a flat surface is known to be proportional to the radius of the microsphere The two types of microspheres used in the present study had the same radius of 3 mm, and therefore the levels of van der Waals attraction were also considered
34
Membrane Filtration in Water and Wastewater Treatment
Carboxyl group (COOH)
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Figure 15 Three relevant forces: (1) electrostatic interaction, (2) van der Waals interaction, and (3) hydrogen bond.
to be similar. Considering the balance of the three relevant forces, it is reasonable to conclude that the difference in the adhesion force shown in Figure 14 may be attributed to the hydrogen bond. A hydrogen bond is generated by electron transfer reaction between electronegative atoms (e.g., O, N, F, and Cl) and H atoms that are covalently bound to similar electronegative atoms. The two functional groups examined in the present study (i.e., carboxyl and hydroxyl groups) have the possibility of forming hydrogen bonds due to their high polar nature, but the bounding power largely depends on their pKa values. If pKa value is larger than pH of the solution, the functional group is protonated, contributing to the formation of a strong hydrogen bond. In contrast, in the condition of pKa being less than pH of the solution, the functional group dissociates, resulting in an insignificant hydrogen bond. To make sure the dissociation condition of two functional groups, an investigation of the adhesion force as a function of pH is considered to be appropriate. However, because of low resistance of available AFM cells to extreme pH condition, the authors could not figure out the dissociation condition of two functional groups. In previous studies in which the pKa values of hydroxyl- or carboxyl-modified microspheres were investigated, it was estimated that carboxyl groups have pKa values between 3 and 6 and hydroxyl groups have pKa values between 9 and 13. Assuming that the pKa values obtained in those previous studies could be applied to the present study, the carboxyl groups were dissociated whereas the hydroxyl groups were not dissociated in the adhesion force measurements carried out at pH 6.8 (Figure 14). In the present study, the difference stated above presumably caused the remarkable difference in adhesion force of the two types of microspheres. An additional remark that should be made for Figure 14 is that the adhesion forces of hydroxyl-modified microspheres to the PVDF membrane and the PE membrane were quite different. As shown in Figure 14, the binding power of the hydroxyl group was much greater for the PVDF membrane than for the PE membrane. According to Ducker et al., the adhesion value possibly varies depending on surface roughness. The difference between the roughness of the PVDF membrane and that of the PE membrane was insignificant, suggesting a limited effect of roughness on the difference in adhesion force. Rather, difference in polymer materials seemed to affect the binding force of hydroxyl-modified microspheres: binding power of the hydrogen bond largely depends on hydrogen
bonded pairs. It is known that PVDF has two fluoride atoms that are arranged symmetrically with a center carbon atom, while PE has only hydrogen atoms along with carbon chain. Generally, the higher the electronegativity of the bounded atom, the greater the binding energy of the hydrogen bond becomes. Because of the high electronegative nature of fluoride atoms, a strong hydrogen bond would be formed between the surface of the PVDF membrane and hydroxyl-modified microspheres. Based on the fact that carbohydrate has many hydroxyl groups in its structure, the hydrogen bond seems to play an important role in the accumulation of carbohydrate-like substances on membranes, as indicated by previous studies on fouling. The hydrogen bond is considered as a semi-irreversible reaction, and the value of binding energy is between 10 and 40 kJ mol1, which is stronger than that of typical van der Waals attraction (B1 kJ mol1). Because of such a strong and semi-irreversible binding ability of the hydrogen bond, it is probably very difficult to remove carbohydrate-like substances from membranes by physical cleaning (e.g., backwashing) once they have adhered to the membranes by hydrogen bonds. The data shown in Figure 14 suggest that more carbohydrate would accumulate on a membrane made from polymers containing atoms with high electronegativity. For the prevention of accumulation of carbohydrate on membranes used for water treatment, it would be desirable to choose membranes that are fabricated with polymers that do not contain atoms with high electronegativity in their structure. Figure 16 shows the approaching force curves repeatedly measured with (a) carboxyl- and (b) hydroxyl-modified microspheres for new PE and PVDF membranes. As shown in the figure, features of approaching force curves were completely different depending on the type of microspheres. As the carboxyl- modified microspheres approached the membrane surface (Figure 16(a)), they encountered repulsive interaction due to repulsive electrostatic interactions between the negatively charged microspheres and the negatively charged membrane surface. It is shown in Figure 16(a) that the interaction became apparent within a distance of about 20 nm for both membranes, demonstrating that the two membranes exerted similar electrostatic repulsion against the carboxyl-modified microspheres. This is consistent with the results of measurement of zeta potentials of the membranes: the two membranes exhibited similar negative charges.
PE 0.5 0.25
0.5 0.25
0.0
0.0 0
25
50
75
0
Distance (nm)
(a)
50
25
75
Distance (nm)
0.5
0.5 PE
PVDF
0.25
0.25
0.0
0.0
–0.25
–0.25
Force (nN)
Force (nN)
35
PVDF
Force (nN)
Force (nN)
Membrane Filtration in Water and Wastewater Treatment
–0.5
–0.5
–1.0
–1.0
–1.5
–1.5
–2.0
–2.0 25
0
50
75
0
Distance (nm)
(b)
25
50
75
Distance (nm)
Figure 16 Approaching force curves of (a) carboxyl-modified microspheres and (b) hydroxyl-modified microspheres to the PE membrane (left panels) and the PVDF membranes (right panels) in buffered solution (pH 6.8).
6 Adhesion force (nN)
Adhesion force (nN)
6 5 4 3 2 1
4 3 2 1 0
0 0 (a)
5
5
10 15 20 25 30 35 40 Operation time (days)
0 (b)
5
10 15 20 25 30 35 40 Operation time (days)
Figure 17 Changes in adhesion force of carboxyl-modified microspheres (a) and hydroxyl-modified microspheres (b) to PE membranes that were sampled during the pilot-filtration test.
In contrast, as the hydroxyl-modified microspheres approached the membrane surface (Figure 16(b)), rapid decrease in the bending stresses of the cantilever or jump-in attraction forces appeared after gradual increase in repulsion force. The increase in attractive force was probably due to hydrogen bonds between the hydroxyl groups of microspheres and the membrane surface. The effective distances of hydrogen bonds were around 15 and 5 nm in the case of the PVDF and the PE membranes, respectively. This was in accordance with the strong adhesion force of the hydroxyl-modified microspheres to the PVDF membrane discussed above. The results shown in Figure 16 suggest that hydrogen bonds between foulants and membranes can be significant only when they are transported to the region where the
membrane surface is very close. Before entering the region where hydrogen bonds can be significant, foulants need to overcome repulsive forces if they bear negative charges. Otherwise, they do not adhere to the membrane surface and subsequently cause membrane fouling. Strongly negativecharged particles/molecules (e.g., humic acid) are less likely to reach the membrane surface: in contrast, it is expected that carbohydrate-like substances relatively easily access to the membrane surface because of their electrostatically neutral nature. This is an additional explanation why carbohydratelike substances have recently been reported to be major foulants. Figure 17 shows the changes in adhesion forces (Fad) of (a) carboxyl- or (b) hydroxyl-modified microspheres to the PE
36
Membrane Filtration in Water and Wastewater Treatment
membranes, which were sampled during the pilot-scale filtration on days 0, 1, 3, 5, 16, 23, and 39. Adhesion force shown in the figure was determined by the same procedure as that used for obtaining the data shown in Figure 14. As clearly shown in Figure 14, adhesion forces of both hydroxyl-modified and carboxyl-modified microspheres changed to a large extent as a result of fouling. In the case of carboxyl-modified microspheres, the adhesion force decreased rapidly to a value of 0.06 nN within 1 day and remained at a low level until the end of operation. One possible reason for the reduction in binding force of carboxyl-modified microspheres was the increase in electrostatic repulsion. The charge of the membrane surface changed from 11 to 28 mV during the pilot operation (Yamamura et al., 2008), which resulted in greater electrostatic repulsion between negatively charged microspheres and the membrane surface. In authors’ previous fouling study using the PE membrane carried out at the same plant, it was shown that negatively charged substances (e.g., humic substances) also accumulated on/in the membrane during the long-term filtration. Accumulation of such negatively charged substances presumably decreased the charge of the membrane surface. As shown in Figure 17, adhesion force of the hydroxylmodified microspheres also declined rapidly, but the values of Fad for the hydroxyl-modified microspheres were much larger than those for the carboxyl-modified microspheres except for on day 39. This result partially explains why hydrophilic NOM dominated over humic substances and was shown to be a major foulant in previous studies on fouling: hydrophilic NOM actually has a great binding power to the membrane due to hydrogen bonding.
The exponential reduction of adhesion force seen with hydroxyl-modified microspheres could presumably be explained by the decrease in binding sites available on the membrane surface due to membrane fouling and/or by the increase in repulsive forces between negatively charged microspheres and the negatively charged membrane surface.
4.02.1.3 Membrane Filtration Systems for Controlling Fouling In order to reduce the membrane fouling, we need to produce the membrane resistant to fouling and to construct hybrid membrane systems which include the existing treatment processes such as coagulation, activated carbon adsorption, and biological/chemical oxidation. Figure 18 describes such a concept, considering the size, concentration, and chemical properties of the substances to be removed.
4.02.1.3.1 Channel flocculation in monolith ceramic membrane Coagulation–flocculation process has been widely used to form aggregates (flocs), which include many fine particles contained in the raw water, for the efficient solid–liquid separation in the sedimentation basin and sand filter. Tambo and Watanabe published several papers describing the floc density and flocculation kinetics for the better understanding of flocculation process. They presented the floc density function and GC0T value. The floc density function describes the quantitative relationship between the size and effective (buoyant) density of flocs. The exponent Kr in the function is related to the fractal dimension (D) for the aggregates formed
Impurities mm Suspended matters
Organic–inorganic soil (clay, microorganisms, highmolecular-weight humics, etc.) Silts MF–UF Algae filtration Protozoa (Cryptosporidium, Giardia, etc.) Bacteria
µm Impurity size
Protein Colloidal matters
Coagulation + MF–UF
Coagulation / sedimentation + MF–UF filtration
Oxidized substances (SiO2, Fe2O3, Al2O3, MnO2, etc.) Humic acids Virus
nm
Dissolved matters
Å
Adsorption, ion exchange + MF–UF
Saccharoid
Ozonation, activated carbon adsorption, biological oxidation + MF–UF filtration
Taste and odor producing inorganic ions (Fe2+, Mn2+, etc) Fulvic acids NF filtration Synthetic organic compounds (DDT, BHC, PCB,) Inorganic compounds (arsenic, antimony, seleninum, etc.)
Concentration Figure 18 Design matrix of hybrid membrane filtration systems. DDT, dichlorodiphenyltrichloro ethane; BHC, benzene hexachloride; MF, microfiltration; NF, nanofiltration; PCB, polychlorinated biphenyl; UF, ultrafiltration.
Membrane Filtration in Water and Wastewater Treatment
in cluster–cluster aggregation (CCA) as D ¼ 3 Kr. Kr is a function of the aluminum to turbidity (ALT) ratio, which is defined as Al dosage(mg/l)/suspended solid concentration (mg l1) in raw water, and has the value of 1.00 and 1.25 for the ALT ratio of around 1/100 and 1/20, respectively. These values coincide with the fractal dimension D determined for the reaction and diffusion limited case (2.05 and 1.75), respectively. Tambo and Watanabe have proposed that the GC0T value is more useful than GT value proposed by Camp as the criterion of flocculation. These research results have been included in the membrane filtration process to improve the filterability of the membrane (Yonekawa et al., 2004). In Japan, membrane filtration plant has increased its treatment capacity since the mid-1990s. Tokyo Metropolitan Water Works Authority constructed a plant with the total capacity of 80 000 m3 d1 in April 2007 using hollow-fiber MF membranes made of PVDF. It is currently the largest plant in Japan. There has also been innovation in the membrane material and membrane module. The monolith ceramic membrane was developed in 1988 and its advances have been remarkable as seen in Figure 19. Figure 20 describes the detail of the monolith ceramic membrane. By the end of 2008, 81 plants with monolith ceramic membrane have been under operation in Japan and the maximum capacity of the plants is about 40 000 m3 d1. The pre-coagulation has been provided to all of these plants to strengthen filterability for stable filtration performance for a
Configuration
Unit
Length
mm
Diameter
mm
Channel number
wide range of raw water turbidity and enhancement of the removal of viruses and dissolved organic substances. The authors have clarified the characteristics unique to monolith ceramic membrane with pre-coagulation by referring to the behavior of microparticles. The region exists in the monolith channel with the optimum G and GC0T value for good flocculation. The flocculation of microparticles offers the reduction in the membrane fouling. The laminar flow model within dead-end hollow-fiber membranes has been presented in many studies. For example, Fujita and Takizawa developed Equation (1) from the energy equation and the material balance in the course of filtration:
dp v 8m ¼ 1 dv g rdkðp p0 Þ
ð1Þ
where p is the static pressure (m), v the axial velocity within hollow fiber (m s1), g the gravitational acceleration (m s2), m the viscosity (kg m1 s1), r the water density (kg m3), d the internal diameter of hollow fiber (m), k the membrane filterability (s1) and p0 the external pressure of membrane (m). Considering the characteristic values (d ¼ 4 104 m, k ¼ 6 106 s1) of the typical hollow fiber, the first term in Equation (1) is much smaller than the second term. Neglecting the first term, an appropriate equation to calculate an expanded approximate axial velocity in a fibre can be derived. In the case of monolith ceramic membrane (d ¼ 2.5 103 m,
Tube 1985
Stage
Monolith 1988
1990
1994
2001
2006
1000 10
1500
30
1
19
37
180 61
2000
Channel diameter
mm
7
4
3
Membrane area
m2
0.02
0.24
0.35
0.48
15
24
Packing density
m2 l–1
0.25
0.34
0.50
0.63
0.6
0.63
1.5m3 8.9m2 1000 Module capacity (m3 d–1 module)
2.5
Industrial use
Application
100
10
Figure 19 Advance in monolith ceramic membrane.
37
13m2
Water purification m–2
d–1
1.8m3 73m2
m–2
d–1
2.5m3 m–2 d–1
5m3 m–2 d–1
150m2
240m2 module–1
38
Membrane Filtration in Water and Wastewater Treatment
Figure 20 Detail of monolith ceramic membrane (META water product).
k ¼ 5 105 sec1), however, the first term in Equation (1) cannot be neglected to derive an appropriate equation for calculating axial velocity in a monolith channel. Without neglecting the first term in Equation (1), the authors have developed Equation (2) to calculate an expanded approximate axial velocity in a monolith channel:
v2 v ¼ vf coshðaxÞ b pf pe þ f sinhðaxÞ 2g rgdk 4dk 2 ; a¼ 2 b ¼ 8m d b
ð2Þ
where pf and vf are the pressure (m) and velocity at inlet of monolith channel (m s1), respectively. On the other hand, the membrane filterability k in the monolith membrane has a certain distribution. To facilitate analysis of the flow pattern on the basis of this distribution, a five-channel model with three levels of filterability was created, as described in Figure 21. Solving Equation (3) under the material balance and appropriate boundary conditions, the equation for axial velocity in the five channel model has been derived as
vi ¼ vfi coshðai xÞ bi ðpfi pe Þsinhðai xÞ ði ¼ 0; 1; 2Þ
ð3Þ
where pfi is the total pressure at channel inlet (m) and pe the external static pressure of membrane (m). The calculated flow pattern in the monolith ceramic membrane module is shown in Figure 22. A concentrate flowing out through outlets of channels 1 and 2 with lower filterability is drawn into channel 0 with higher filterability. It was also confirmed that the dead-end point is located at the position with an axial velocity vi ¼ 0 in channel 0.
In the channel of 1 m long, axial velocities calculated by Equation (9) are shown in Figure 21 for the membrane flux of 2 m3 m2 d1. The G value in the channels 0–2 was calculated at about 40 s1, which is in the range of optimum values proposed by Camp. On the other hand, the mean hydraulic residence time in the channels 0–2 was about 50 s. Therefore, the GT value in the channel is only about 2000, which is too small compared with the Camp’s proposed values. However, good flocculation was observed in the channel, because the GC0T value in the part of channel is high enough for good flocculation, explained as below. Using the data shown in Figure 23, the distribution of the local G values within the channel 2 under the membrane flux of 2 m3 m2 d1 is described as seen in Figure 23. Considering the velocity distribution in the channel and high concentration of coagulated microparticles reflected by membrane filtration, the GC0T value may be high enough for a good flocculation in the region with the local G value of 40–100 s1. In this context, C0 is defined as the coagulated microparticle concentration near the entrance of such a region. Figure 24 shows the experimental setup (large and small monolith membrane module) and sampling points. The top and bottom portion of the both modules were made of transparent material to enable a visual observation of flocs using video camera. Raw water was taken from the Kiso River near Nagoya city. The dosage of coagulant (polyaluminum chloride, PACl) was fixed at 1 mg Al l1. For rapid mixing condition, G value was fixed at 150 s1 and hydraulic detention time at 300 s. The filtration mode was dead end and membrane flux was fixed at 2 m3 m2 d1. The specifications and operation conditions of the two membrane modules are described in this chapter.
Membrane Filtration in Water and Wastewater Treatment
39
Eq. (3) Velocity equation for a 5-channels model
5 channels
Average velocity in channels (m s−1)
Channel no.0 Permeability k = 5.80 × 10–5[s–1]
Channel no.1 Permeability k = 4.65 × 10–5[s–1]
Channel no.2 Permeability k = 4.07 × 10–5[s–1]
Channel no.1 Permeability k = 4.65 × 10–5[s–1]
Channel no.0 Permeability k = 5.80 × 10–5[s–1]
i = fi cos(i x ) – i (p f,i – pe) sinh(i x ) (i = 0, 1, 2)
0.04
Channel no. 0 Channel no. 1 Channel no. 2
0.03 0.02
0.01
2 m3 m–2 d–1
0 0
0.2
0.4
0.6
0.8
1.0
Channel axial coordinate (m) Figure 21 Five-channel model and filterability k.
Module casing
Dead-end point
90%
Membrane
96.5%
Feed
Figure 22 Flow pattern in monolith module.
With the laser diffraction scattering-type particle-size distribution cell holder (Horiba LY-073), the particle-size distribution was measured to verify the predicted flocculation phenomena and its effect on the filterability of the monolith ceramic membrane. The behavior of microparticles with the size of 0.5–15 mm in the channel with lower filterability was also measured to identify the critical particle size. Polystyrenetype latex particles (JSR Stadex/Dynospheres: 0.5, 3, 5, 10, 15 mm, specific density of 1.05) were used as model particles.
The authors also investigated the correlation between microparticle concentration and TMP using the effluent from a conventional rapid sand filtration process, as shown in Figure 25. There exists a clear relationship between them. It would suggest a significant effect of the flocculation on the filterability in the monolith channel, because the microparticles, larger than 1 mm in the shear field, are subject to a lift force such as the lateral migration and shear-induced diffusion which are proportional to square and cubic power of the equivalent particle diameter, respectively, as described in Figure 26. There were no visual flocs in the bottom portion of the module where coagulated microparticles entered. Visual flocs, however, blew out at the maximum velocity of 3–8 mm s1 from the lower filterability channels in the upper portion of the module. From the analytical result with five channel model, the average outflow velocity at the membrane top was estimated to be 2–4 mm s1. The maximum flow velocity in laminar flow is twice the average velocity. Therefore, the analytical result has been confirmed by the visual experiment. The authors measured the concentration of polystyrenetype latex particles with the size range of 0.5–15 mm in the influent and effluent of the membrane. There were almost no particles in the effluent. It demonstrated that the latex particles of smaller than 15 mm are deposited onto the membrane surface in the course of membrane filtration. This result can explain the correlation of the variation of microparticle number in raw water and TMP as seen in Figure 25. The experiment on the flocculation in the monolith channel was carried out to prove that good flocculation occurs in the channel and will improve the filterability of the membrane. From the theoretical analysis, the average flow velocity in the channel with lower filterability is about 0.5 mm s1 in
40
Membrane Filtration in Water and Wastewater Treatment Recovery 90%
Channel diameter 2.5 mm
Let’s consider “flocculation condition in channel” Channel length 1000 mm
G value 20 sec–1
Especially, near the membrane surface G value 20−100 s –1 : desirable value for flocculation Contact Time
40
60
80
Enough : laminar velocity is very low 100 Concentration Highly concentrated : accompanied by filtration Flux 2 m3 m–2 d–1 Figure 23 Profile of G values in monolith channel with lowest k.
Frequency (volume based ) (%)
20
SP3
16
SP4
12 SP2 SP1
8
Sp4
4 Filtrate
0 1
10
100 Particle size (µm) Small membrane
Coagulant (PACI) M
M
P SP1
SP2
SP3
Figure 24 Experimental setup and sampling points (SPs). PACl, polyaluminum chloride.
the region of 1–200 mm from the surface, so the detention time is between a few tens of minutes and few hours. The G value in the zone is between 20 and 100 s1. The floc size distribution in each sampling point is seen in Figure 24. Flocs are lifted up by laminar flow and carried away from the outside of the channel. Therefore, the space near the membrane surface might be considered to be a high efficient field for coagulating the charge neutralized microparticles.
Figure 27 shows a schematic image of phenomenon occurred in the channel when the pre-coagulation is prepared. In order to confirm the flocculation effect on the improvement of ceramic membrane filterability, the authors carried out an additional experiment using the small module with the Nishitappu River water. It is a very clean water with annual average turbidity and DOC of 1.3 TU and 0.6 mg l1, respectively.
Membrane Filtration in Water and Wastewater Treatment
41
CSF treated water
Run 6
Pore size
Flux
Interval
Pressure
Recovery
1.0 mm
20 m3 m–2 d–1
15 min
300 kPa
93.3%
TMP (kPa)
50 10
40 5 TMP Microparticle 30 29 Oct.
30 Oct.
Microparticle count (103 ml–1) (0.5–1.0 µm)
15
0
31 Oct. Date
Figure 25 Correlation between transmembrane pressure (TMP) and microparticle concentration. CSF, coagulation/sand filtration.
0 Monolith ceramic membrane
Back transport –2 log cm s–1
F1 ux Membrane
Log transport velocity (cm s–1)
2 Minimum size of particle that will not deposit on membrane
0
DpL = 56 µm
d–1
–4 0.8 µm
–6
DpL = 87 µm
–8 –10
Ultrafiltration flux
–4
–8 –4
m–2
Microfiltration flux
–2
–6
2
m3
Brownian R = 0.03 cm u = 133 cm s–1 T = 20 °C –3
–3
Shear Calculation conditions
–2 –1 0 1 Particle diameter: log Dp (µm)
2
3
Channel diameter = 2.5 mm Water temperature = 20 °C Channel entrance
Lateral migration
–2 –1 0 1 Log particle diameter (µm)
Middle point of channel 2
Back-transport velocity and critical flux Figure 26 Particle size and lift force.
Figure 28 shows the experimental result and confirms the effect of flocculation on the fouling reduction. Further improvement is possible using the chemically enhanced backwashing (CEB) with acidic solution. Coagulant addition of 1 mg Al l1 to the monolith ceramic MF membrane system also improved the virus log removal efficiency up to 7. Figure 29 shows the experimental verification of the effect of the CEB on the membrane filterability. The reason behind the
improvement may be the removal of microflocs attached to the membrane surface by the ECB.
4.02.1.3.2 Pre-coagulation/sedimentation in hollow-fiber UF/MF membrane The surface water from Chitose River and Nisitappu River was used as the raw water in the experiment. Table 2 summarizes
42
Membrane Filtration in Water and Wastewater Treatment
L = uo2 dp3/(32 ro2)
Lateral migration Shear-induced diffusion
S = 0.05 uo dp2/(4 ro2)
Ceramic membrane surface
Disaggregated floc particles
Lift force
Ceramic membrane surface u (membrane flux) (b) Aggregation
(a) Carrying near the membrane accompanied with filtration
(c) Lifting from membrane
Figure 27 Schematic image of effect of channel flocculation.
80
4 TMP Membrane flux
TMP (kPa at 25 °C)
3 Back washing interval: 4h
2h
40
2
20
1
0 1/4
Membrane flux (m d–1)
Precoagulation
60
0 1/19
2/3
2/18
3/5
3/20
4/4
Figure 28 Effect of channel flocculation on transmembrane pressure (TMP) change.
CEB (acid)
Experimental flow Coagulation
TMP (kPa)
Mn oxidization
Ceramic membrane
40
10 m3 m–2 d–1
30
8 m3 m–2 d–1
20
6 m3 m–2 d–1
10 0 04 Jan.
4 14 Jan.
m3
m–2
d–1,
24 Jan.
with CEB
without CEB 03 Feb. Date
13 Feb.
Figure 29 Effect of CEB, chemically enhanced backwashing on TMP change under high flux operation.
23 Feb.
Membrane Filtration in Water and Wastewater Treatment
the average raw water quality of Chitose river during the experiment (Jang et al., 2004). With Chitose River water, the pilot plant consists of a rapid mixing tank, a jet mixed separator (JMS) with inclined tube settlers, and three hollow-fiber UF or MF membrane filters as described in Figure 30. The JMS is a simple but effective solid/liquid separator with several vertical porous plates in a channel; microflocs are flocculated under the turbulent flow produced by the water jets and larger parts of grown flocs settle between the plates; subsequently, residual small flocs are removed in the inclined tube settlers. The effective volume of JMS with inclined tube settlers is 7.0 m3 and flow rate to the JMS was 120 m3 d1, corresponding to the hydraulic detention time of 84 min. The operating conditions of this pilot plant are summarized (refer to Jang et al., 2004). Four processes of the pilot experiment were carried out. In processs 1 and 2, the aluminum sulfate (AS) with activated silicate and PACl was used as coagulant. The water was fed from outside to inside of hollow-fiber UF membrane, which is made of specially polymerized PAN with nominal average pore size of 0.01 mm, at a constant permeate flow rate of 0.9 m d1. The physical cleaning with back washing and air scrubbing was carried out to prevent fouling in a time interval of 30–60 min. In processes 3 and 4, polysilicato-iron (PSI) which is inorganic polymeric iron coagulant was used as coagulant PSI has a molecular ratio of Fe to Si of 1:1–1:5, but we used the molecular ratio with 1:1 in this pilot plant experiment. Coagulant dosage and coagulation pH were 0.21 mmol Fe l1 and 6.2, respectively. Results of TMP trends, for Chitose River water, with increasing UF filtration time in process 1 are shown in Figure 31. Figure 32 shows the comparison of the TMP among processes 1, 2, and 3 using the same UF membrane and AS, PACl, and PSI as coagulant, respectively.
Transmembrane pressure at 25 °C (kPa)
Considering the data shown in Figure 33 and Table 3, it may be concluded that the higher DOC removal in the precoagulation/sedimentation gives better performance of UF membrane filtration. Even though the TMP used by PACl and PSI was almost the same at about 3300 h of filtration time (the actual TMP reached about 100 kPa, which is the recommended TMF for chemical cleaning), TMF used PSI has always been lower than that by AS and PACl. Figure 33 shows the comparison of removal efficiency of DOC among the three coagulants. PSI gave the best removal efficiency resulting in the best filtration performance. With Nishitapu River water, the TMP increased in each operating condition as seen in Figure 34. When the Nishitapu River water was filtered at constant flow rate of 1.1 m d1 directly by using UF membrane, the filtration time to reach 100 kPa of TMP was only 300 h in spite of low organic content and low inorganic content. However, the filtration time for coagulated water was 4 times longer than that. In addition, hypochlorite solution was added
150
Raw-UF Coa.-UF Sed.-UF
120 90 60 30
Flux: 0.9 m d–1
0 0
500 1000 1500 2000 Membrane filtration time (h)
Jet mixed separator (JMS)
P
Rapid mixing tank Permeate
Permeate
Permeate
Drain P
Compressor
Figure 30 Schematic description of pilot plant.
2500
Figure 31 Effect of pre-coagulation/sedimentation on performance of ultrafiltration (UF) membrane system.
Coagulant
Chitose river water
43
P
P
44
Membrane Filtration in Water and Wastewater Treatment
during backwashing term; the UF membrane filterability was significantly improved. These results were also obtained in the case of using MF membrane as seen in Figure 35.
4.02.1.3.3 Hybrid submerged MF membrane system
Transmembrane pressure at 25 °C (kPa)
The hybrid MF membrane system is a combination of submerged membrane and the other processes such as the powdered activated carbon adsorption and chemical/biological oxidation. The membrane system has been developed to purify raw waters with low quality containing a lot of soluble matter such as biodegradable organics, humic substances, manganese, and ammonia nitrogen. In the hybrid system, soluble less-biodegradable organics are adsorbed to the
120
Sed.-UF (PSI: 0.21 mmol-Fe l–1)
100
Sed.-UF (PACI: 0.19 mmol-Al l–1) Sed.-UF (AS: 0.37 mmol-Al l–1)
80
powdered activated carbon, and suspended particles including powdered activated carbon are separated by the membrane filtration. The soluble biodegradable organics, manganese and ammonia nitrogen, are biologically or chemically oxidized. In the case of chemical oxidation (with prechlorination), soluble manganese is oxidized with chlorine and the catalytic reaction of powdered activated carbon, and the oxidized manganese is removed by membrane separation. Ammonia nitrogen is also oxidized by chlorine in a pre-chlorination tank. In the case of biological oxidation (without prechlorination), the iron oxidizing bacteria and ammonia oxidizing bacteria, which are concentrated in submerged membrane tank, oxidize the soluble manganese and ammonia, respectively. A schematic diagram of the pilot plant is shown in Figure 36 (Suzuki et al.,1998). The volume of membrane submerged tank and the surface area of submerged membrane were 4 m3 and 86–120 m2, respectively. Detention time in the mixing tank was 10–15 min. The raw water was fed into the mixing tanks. Four types of polytetrafluoroethylene (PTFE) membranes were used. When the first, second, and third type of membranes were used, the
60 Table 3
Physically irreversible resistance
40 Run 1(125 days)
20
Run 2(73 days)
Module Module Module Module A B A B
0 0
500
1000 1500 2000 2500 3000 3500 Membrane filtration time (h)
Figure 32 Effect of various coagulants on performance of ultrafiltration (UF) membrane system.
Membrane flux (m3m2d1) Physically irreversible filtration resistance (1011 m1)
0.2 0.29
0.6 1.52
0.4 0.69
0.8 2.48
100 Removal efficieny of DOC (%)
Removal of UF
Removal of coagulants
Removal of sedimentation
80 PSI
60 PACI
40
AS
20
Run 1
Run 2
RW -U F C oa -U JM F SU F
RW -U F C oa -U JM F SU F
RW -U F C oa -U JM F SU F
RW -U F C oa -U JM F SU F
0
Run 3
Run 4
• Direct UF; around 15%, precoagulation / sedimentation; 35–60% • The highest DOC removal efficiency was obtained in run 4 Figure 33 Removal efficiency of dissolved organic carbon (DOC) with various coagulants. AS, aluminium sulfate; PACl, polyaluminum chloride; PSI, polysilicato-iron.
Transmembrane pressure at 25 °C (kPa)
Membrane Filtration in Water and Wastewater Treatment
45
Raw-UF (hypochlorite solution was not added, flux: 1.1 m d–1) Coa.-UF (hypochlorite solution was not added, flux: 1.1 m d–1) Coa.-UF (hypochlorite solution was added, flux: 1.1 m d–1) Coa.-UF (hypochlorite solution was added, flux: 1.7 m d–1)
140 120 100 80 60 40 20 0 0
500
1000
1500
2000
Membrane filtration time (h)
Transmembrane pressure at 25 °C (kPa)
Figure 34 Effect of operation condition on performance of ultrafiltration (UF) membrane system.
100
Raw-MF (hypochlorite solution was added) Coa.-MF (hypochlorite solution was added)
80 60 Flux: 1.4 m d –1
40 20 0 0
100 200 300 400 Membrane filtration time (h)
500
Figure 35 Effect of precoagulation on microfiltration (MF) membrane system.
raw water to the pilot plant was taken from the existing water purification plant, which had already contained the powdered activated carbon in the concentration of 5–30 mg l1. In the first tank, hypochlorite was added when the chemical oxidation was applied, and the sludge containing biomass and activated carbon were returned from the membrane submerged tank and mixed with the raw water in the second tank. The same powdered activated carbon (average diameter of 10 mm) was dosed into the second tank at the constant concentration of 13 mg l1 when the fourth PTFE membrane was used. PACl was added to coagulate the small suspended particles in the third tank. The pretreated water was fed into the submerged tank where the hollow-fiber PTFE membranes were submerged and intermittent aeration was performed to supply the oxygen to the microorganisms. Air was supplied for 1 min. with the intensity of 0.64 N m3 min1. every 4 min. The raw water in the pilot plant study was surface water from Chitose River. It contained many humic substances, soluble manganese, and ammonia nitrogen. The average raw water quality is given in Table 2. The dosage of hypochlorite in the mixing tank was 4–6 mg Cl2 l1. Coagulant dosage was 2–3 mg Al l1.
In the pilot plant experiments, symmetric or composite PTFE membrane with nominal pore size of 0.1 mm was used. The thickness of skin layer in the composite membrane was changed at 60, 30, and 15 mm. The pore density was about 80%. The skin layer thickness and pore density of the newest composite PTFE membrane are about 15 mm and 80%, respectively. The hybrid membrane system is able to efficiently remove the soluble matter such as organics, manganese, and ammonia nitrogen. The soluble manganese and ammonia nitrogen were oxidized biologically or chemically and small humic substances were adsorbed to the powdered activated carbon. The removal efficiency of TOC, E260, and trihalomethane formation potential (THMFP) is shown in Figure 37, and the comparison in the removal efficiency of the soluble manganese and ammonia nitrogen between chemical oxidation with prechlorination and biological oxidation without prechlorination was made in Figure 38 when the composite PTFE membrane with the skin layer thickness of 30 mm was used. Dosage of powdered activated carbon was fixed at 13 mg l1. As previously reported by the authors, chemical oxidation is necessary to oxidize soluble manganese when the raw water temperature become less than 10 1C. The authors also reported that improved filtrate quality can contribute to keep a higher flux. Figure 39 shows the change of the permeate flux, TMP, and raw water temperature during the experiment with the symmetrical membrane. In this experiment, the hybrid MF membrane system was operated without prechlorination. The average flux was relatively low at less than 0.3 m d1 and the TMP increased to 70 kPa after 5 months of operation. To improve the permeability of the PTFE membrane, the structure of membrane was changed from symmetrical to composite. Permeability of the composite membranes with different skin layer thickness was compared in the pilot plant experiment. The thinner the skin layer thickness, the better the permeability. Figure 40 shows the change of the membrane flux, TMP, and raw water temperature with increasing operation time when the newest composite PTFE membrane was used. It demonstrates that TMP was very stable under a high flux of 1.2 m d1. It is about 4 times higher than that in the symmetrical membrane.
46
Membrane Filtration in Water and Wastewater Treatment Hybrid submerged PTFE MF membrane system including coagulation, carbon adsorption and biological oxidation Submerged MF membranes PAC Cl2 PACl Raw water
C
P Suction pump
Compressor
P
Storage tank of permeate
P Circulation pump Hydraulic retention time = 1.5 h Figure 36 Schematic description of pilot plant. PAC, powdered activated carbon; PACl, polyaluminum chloride.
4 3.5
Raw water Raw water (soluble) Membrane filtrate
0.12 0.1
2003/9/1~2004/8/31 0.099 0.086
mg l–1
2.5
2.43
1/cm mg / l
3 2.29
2 1.5
1.17
0.08 0.06 0.04
0.031
1 0.02
0.5 0
0.017
0 TOC
E260
THMFP
Figure 37 Removal efficiency of total organic carbon (TOC), E260, and THMFP in hybrid system.
4.02.1.3.4 PVDF Membrane filtration with pre-ozonation Combination of ozonation with membrane filtration is effective for the prevention of membrane fouling. Japanese membrane manufacturing companies have developed the ozone-resisting membrane module made of PVDF with potting material having a high resistance to ozone. In the developed membrane module, water containing residual ozone can be directly filtered. It is reported that this system can provide consistently high permeate flux for various raw waters, especially high turbidity water and secondary treated municipal wastewater. We studied the effect of residual ozone on fouling reduction using the ozone resisting PVDF membrane.
Figure 41 shows the schematic diagram of the experimental system. The same raw water was used as the hybrid membrane system. The average water quality is shown in Table 2. In experimental runs 1-1 and 1-2, ozone dosage was 2.0 and 4.2 mg O3 l1, in which the residual ozone concentration was 0.73 and 1.13 mg O3 l1, respectively. The ozone contact time was of 7.8 min in all experimental runs. In runs 2-1 and 2-2, ozone dosage was 1.4 and 1.9 mg O3 l1, in which the residual ozone concentration was 0.41 and 0.61 mg O3 l1, respectively. Figure 42 shows the TMP change with increasing operation time in run 1 where the constant permeate flux mode operation under 3.5 m d1 with physical cleaning of backwashing
Membrane Filtration in Water and Wastewater Treatment 0.3
mg l–1
0.15
0.2
0.090
0
0.1
0.05
Manganese
0.01
0.016
0.05
0.15
0.002
0.1
2003/11/13–2004/8/31 with prechlorination (chemical oxidation)
0.100
0.15 0.105
mg l–1
0.25
2003/6/25–11/13 without prechlorination (biological oxidation)
0
Ammonia nitrogen
Manganese
0.01
0.25
Raw water Raw water (soluble) Membrane filtrate 0.23
Raw water Raw water (solouble) Membrane filtrate
0.079
0.3
0.2
47
Ammonia nitrogen
30
Flux Water temperature
0.6
25 20 15
0.4
10
0.2
5 0
30 Nov
31 Oct
1 Oct
1 Sep
3 Jul
100 90 80 70 60 50 40 30 20 10 0
2 Aug
0
Water temperature (°C)
0.8
3 Jun
TMP (kPa)
Membrane flux (m d–1)
Figure 38 Removal efficiency of Mn and NH4–N in hybrid system.
2002
0
30
60
90 Operating days
120
150
Figure 39 Transmembrane pressure (TMP) changes and permeates flux (symmetrical PTFE membrane with pore size of 0.1 mm).
and air scrubbing was carried out. It clearly demonstrates that the residual ozone reduced the membrane fouling (Lee et al., 2004). After the continuous operation for about 1800 h, chemical cleaning of fouled membrane was conducted. The following three chemical solutions were used: NaOH solution of 1%, NaClO solution of 5 mg l1, and oxalic and nitric acid of 2%
and 5%. Figure 43 shows the extracted TOC in each chemical solution. As seen in Figure 41, preozonation with residual ozone significantly decreased the attached organic substances to the membrane causing the physically irreversible fouling. It may come from the following two ozone-induced reactions: degradation of organic substances and destabilization of particles on the membrane surface. The ozone-induced particle
Membrane Filtration in Water and Wastewater Treatment 50
1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0
Average membrane Flux: 1.2 m d –1 40
:
30 Temperature
Flux
Flux
Water temperature
20 10
Water temperature (°C)
Membrane flux (m d–1)
48
0
100 90
TMP (kPa)
80 70 60
Chemical washing
50 40 30 20
0
30
60
90
120 150 Operating days
180
30 Aug
31 Jul
1 Jul
1 Jun
2 May
2 Apr
3 Mar
2 Feb
3 Jan
0
4 Dec
10
210
240
270
Figure 40 Transmembrane pressure (TMP) changes and permeates flux (new composite PTFE membrane with a skin layer of 15 mm and porosity of about 80%).
destabilization reaction has been reported by many researchers. In the other experiment, we measured the size distribution of fouling particles in the backwash water and found that the average size of the particles was about 20 and 50 mm without and with ozonation, respectively. Increasing particle size increased the rate of back transport of organic particles, leading to the decrease in the accumulation of organic particles on the membrane surface. In this experiment, permeate TOC (i.e., DOC) concentration in the membrane filtration system without and with preozonation was the same as 2.4 mg l1 but E260 and E260/DOC were 0.062 and 0.034 cm1, and 0.026 and 0.013 cm1 mg1 l1, respectively. These results demonstrate that the biodegradability of organic particles increased due to the oxidation by O3.
4.02.2 Membrane Application to Wastewater Treatment 4.02.2.1 Current Status of MBRs Necessity of recycling use of water has been recognized to resolve the shortage of water resources. Municipal wastewaters seem to be an important water resource for recycling use. MBR is a key technology for creating the reclaimed water resource. MBR has been applied to the municipal wastewater treatment since the 1980s. The first-generation MBR combines a crossflow-type membrane with outside bioreactor and mixed liquor
is recirculated into membrane module. The operation pressure is high and recirculation pump is needed. In addition, it is reported that microorganism activity decreases due to the recirculation of the mixed liquor. The second-generation MBR submerges membrane module directly in the bioreactor. As a result, circulation pump is not needed and the operating pressure is low. Submerged MBRs have been preferred due to their lower energy consumption and smaller footprint compared with recirculated MBRs. However, it is reported that accumulated dissolved organic matter in the bioreactor decreases the membrane permeability in the submerged MBR more seriously compared with the first-generation MBR. In 2005 the European Commission decided to boost the development and application of MBR processes for municipal wastewater treatment through financing a 3-year research project within the scope of the 6th framework program: AMEDEUS (accelerate membrane development for urban sewage purification). Within AMEDEUS an analysis of the potential for MBR standardization was carried out. Based on an extensive survey of the MBR industry, the White Paper was published to provide a comprehensive overview of the market interest/expectation and technical potential of going through a standardization process of MBR technology in Europe. Due to the predominance of submerged MBR system in municipal applications, representing 99% of the installed membrane surface in Europe in the period 2002–05, the study focuses only on this configuration.
Membrane Filtration in Water and Wastewater Treatment
49
Preozonation-PVDF MF membrane system O3
Permeate
Chitose River water Ozonationmembrane Residual O3
PVDF membrane Back wash P
Run-4.1, 4.2 Run-5.1, 5.2
Air
Ozonation tank Retention tank O3 O2 Ozonationmembrane No residual O3
Back wash P
Run-4.3 Run-5.3
Air
O3 removal tank Membrane Run-4.4 Run-5.4
Pressurized membrane ⇒ Pore size: 0.1 μm (MF) ⇒ PVDF (polyvinylidenefluoride)
Back wash P
Air Figure 41 Experimental system of preozonation and membrane filtration.
300 Run 1 Run 2 Run 3 Run 4
TMP at 25 °C (kPa)
250 200 150 100 50 0
0
300
600
900
1200
1500
1800
Operation time (h) Figure 42 Effect of preozonation and residual ozone on membrane filtration. TMP, transmembrane pressure.
Figure 44 shows the number of municipal and industrial MBRs in Europe. In Japan MBR technology has been applied to the water recycling for some large business, commercial and residential complex buildings such as Roppongi Hills and Tokyo Mid Town. Membrane fouling deteriorates the membrane permeability and consequently increases energy consumption in MBR. A seriously fouled membrane must be cleaned with chemical reagents, which are costly. In addition, disposal of chemical reagents after membrane cleaning is an issue of concern, and the frequency of chemical membrane cleaning
should therefore be minimized. Thus, there is a need for efficient control of membrane fouling in MBR. In order to develop methods for efficient MBR operation, a better understanding of the mechanism of membrane fouling in MBR is needed.
4.02.2.2 Mechanism of Membrane Fouling Membrane fouling is a major obstacle for wider application of MBRs. Membrane fouling results in reduced performance, severe flux decline, high-energy consumption, and frequent
50
Membrane Filtration in Water and Wastewater Treatment 14 NaOH NaClO (Oxalic + nitric) acid
Detached organic substances per unit permeate volume (mg-TOC m–3)
12 10
Without O3
8 6 4 Residual O3 2 0 Run 1-1
Run 1-2
Run 1-4
Figure 43 Effect of residual ozone on amount of detached organic carbon.
600
Number of installations
With standardization?
Industrial (> 20 m3 d–1) Municipal (> 500 p.e.)
500 400
> 50 per year 300 200 > 20 per year 100 0 7
99
<1
98
19
99
19
00
20
01
20
02
20
03
20
04
20
05
20
06
20
07
20
08
20
09
20
10
20
Year Figure 44 Number of MBRs in Europe. From AMEDEUS (2006).
membrane cleaning or replacement. In order to establish strategies for controlling membrane fouling, an understanding of the mechanisms of membrane fouling is indispensable. Many factors that might influence membrane fouling in MBRs have been reported. Attention has been given to various design and operating parameters such as airflow rate in the reactor, membrane configuration, membrane flux, concentrations of mixed liquor-suspended solids (MLSSs), and solids retention time. Figure 45 describes several factors affecting the membrane fouling in MBRs (Yamato et al., 2006). In this part, the authors focus on two important operating parameters, membrane flux and membrane material.
4.02.2.2.1 Effect of membrane permeate flux on fouling Experiments were carried out at the Soseigawa Municipal Wastewater Treatment Center, Sapporo, Japan. Characteristics of the raw wastewater of this plant can be found in Kimura et al. (2005). The examined wastewater is classified as weak. Feed wastewater for the MBR examined in this study was
delivered from the grit chamber of the facility. Hollow-fiber MF membrane modules made of PVDF that had a total surface area of 1.3 m2 each and nominal pore size of 0.4 mm (Mitsubishi Rayon, Tokyo, Japan) were used in this study. Two identical membrane modules were submerged in the same MBR tank (350 l) and filtered the same mixed liquor suspension side by side as described in Figure 46 (Kimura et al., 2005, 2008a, 2008b). To investigate the influence of difference in membrane flux on fouling, the two modules were operated at different fluxes. Any differences between the two modules could be attributed solely to the influence of membrane flux. In this study, two long-term operations of the MBR (runs 1 and 2) were carried out. In run 1, one module (module A) was operated at 0.2 m3 m2 d1, while the other (module B) was operated at 0.6 m3 m2 d1. In run 2, modules A and B were operated at 0.4 and 0.8 m3 m2 d1, respectively. In each run, new membranes were used. Continuous monitoring was initiated in September 2005 for run 1 and in September 2006 for run 2. In the MBR, aeration was continuously carried out at the flow
Membrane Filtration in Water and Wastewater Treatment * Water purification
51
Module configuration, material (PE, PVDF, PFTE) hydrophilicity/hydrophobicity, porosity, pore size
Membrane
Turbidity, humic substances Characteristics of raw water Temperature, TOC, pH, alkalinity
Membrane fouling
Operating conditions
Molecular-level analysis
Permeate flux, aeration intensity, TMP, HRT, SRT, F/M ratio
Characteristics of mixed liquor
Hybrid system
MBR Multifunctional system
MLSS, viscosity, EPS/SMP, dissolved matter, DO, particle-size distribution, activity of microorganisms
Figure 45 Factors relating to the membrane fouling in MBR. DO, dissolved oxygen; EPS, extracelluar polymeric substance; HRT, hydraulic retention time; MLSS, mixed liquor-suspended solid; PE, polyethylene; PVDF, polyvinylidene fluoride; SMP, soluble microbial product; SRT, solid retention time; TMP, transmembrane pressure; TOC, total organic carbon.
Real wastewater
No. 1
Flux Membrane material
No.1
No.2
0.2
m3
No. 2
m2 d–1
0.6 m3 m2 d–1
PVDF
Nominal pore size
0.4 µm
Surface area
1.3 m2
MLSS
12000 mg l–1
HRT
10.2 h
SRT
111 days Organic analysis
Characterization of foulant NaOH (pH 11) Figure 46 Experimental setup – effect of flux. HRT, hydraulic retention time; MLSS, mixed liquor-suspended solid; SRT, solid retention time.
rate of 3.5 m3 h1. Intermittent filtration (12-min filtration and 3-min pause) was also carried out. MLSS concentration in the MBR was maintained at 11 g l1 and resulting SRT was 110 days in run 1, while MLSS and SRT in run 2 were 12 g l1 and 65 days, respectively. The degree of membrane fouling was estimated by membrane filtration resistance calculated by the following equation:
J ¼ DP=m=Rt
where J is the membrane permeate flux (m3 m2 s1), DP the TMP difference (Pa), m the water viscosity (Pa s), and Rt the total membrane filtration resistance (m1). In run 1, operation of the MBR equipped with the two membrane modules was continued for 125 days. As mentioned before, the membrane fluxes were different for the two modules: 0.2 m3 m2 d1 for module A and 0.6 m3 m2 d1 for module B. As expected, increase in the filtration resistance in module B was much faster than that in module A. However, increase in the filtration resistance in module B was still slow,
52
Membrane Filtration in Water and Wastewater Treatment
and filtration using both modules could be continued for 125 days without any chemical cleaning in run 1. A similar result was obtained in run 2 where operation of the MBR was continued for 73 days: increase of filtration resistance in module A with the low flux (0.4 m3 m2 d1) was much slower than that in the other. Chemical cleaning of the membranes was not carried out in run 2, either. Membrane fouling can be categorized into two types: physically reversible fouling and physically irreversible fouling. Physically reversible fouling can be cancelled by physical cleaning, whereas physically irreversible fouling needs chemical cleaning to be cancelled. Control of physically irreversible fouling is essential for reduction of operating costs of MBRs because physically reversible fouling can be mitigated as long as an efficient physical cleaning is carried out. Most of the existing MBRs are operated with routine practices of physical cleaning (e.g., backwashing). To specifically focus on physically irreversible fouling in this study, each membrane module was intensively cleaned by spraying pressurized water and wiped with lab paper at the end of the operations, and then water permeability was measured. Table 3 summarizes the magnitude of physically irreversible resistance measured at the end of the operations. As mentioned above, in run 1, membrane flux of module B was set 3 times higher than that of module A. However, the difference in the degree of physically irreversible fouling between the two modules was larger than threefold. The degree of physically irreversible fouling in module B was about 5 times larger than that in module A. In the operation of module B, increase in physically irreversible fouling was 1.7 times more rapid than that of module A on the basis of volume of the suspension filtered. A very similar result was obtained in run 2. Although the difference in membrane flux between the two
modules was twofold in run 2, the degree of physically irreversible fouling in module B operated at the higher flux was about 4 times larger than that in module A operated at the lower flux. These indicate that the degree of physically irreversible fouling in an MBR is not directly linked to the volume of the suspension filtered but differs depending on membrane flux. To investigate the features of constituents that were responsible for membrane fouling in the pilot runs, organic matter was desorbed from the fouled membranes at the terminations of the operations and was then analyzed. When the pilot operations were terminated, membrane modules were taken out from the reactors and were disassembled. Each membrane fiber was manually wiped with lab paper in order to remove accumulated cake that could be physically removed. Desorption of organic matter from the fouled membranes was carried out by soaking the membranes in alkaline solution (sodium hydroxide) for 24 h. Solution pH was set at 11. After desorption, measurements of TOC were carried out. The remaining solutions were subsequently processed with electric dialysis for desalination and lyophilized for advanced analyses. Amounts and characteristics of the foulants that caused physically irreversible fouling in the two experiments were investigated by analysis of foulants extracted from the fouled membranes with sodium hydroxide. Figure 47 shows the amounts of foulants desorbed from the fouled membranes. Data in Figure 47 are shown on the basis of a unit membrane surface area. As seen from Figure 47, the amounts of organic matter, which were expressed in extracted concentrations of TOC, carbohydrate, and protein, were not significantly different between the two modules in both runs. Although the amounts of organic matter desorbed from module B were slightly larger than those desorbed from
90
Amount of desorbed organic matter (mg m−2)
80 70
Module A (run 1) Module B (run 1) Module A (run 2)
60 Module B (run 2) 50 40 30 20 10 90 Total organic carbon
Figure 47 Amount of foulants desorbed from fouled membrane.
Carbohydrate
Protein
Membrane Filtration in Water and Wastewater Treatment
module A, the differences were much smaller than the difference seen for the filtration resistances shown in Table 3. On the contrary, the amount of protein desorbed from module A was larger than that from module B in run 2. These data imply that quality of organic matter rather than quantity of organic matter should be focused for an explanation of the differences. Figure 48 shows the composition of monosaccharides in the foulants desorbed from the fouled membranes. Very interestingly, the presence of rhamnose was large in the foulants desorbed from module B that was operated with the higher fluxes in both runs. Figure 49 shows amino acid compositions in the foulants desorbed from the two membranes. Compared with the monosaccharide analysis shown in Figure 48, in the case of amino acid composition, differences between the two modules were not significant in both runs. However, in both runs, the presence of glutamic acid (GLU) and aspartic acid (ASP) was more pronounced in the foulants desorbed from module A operated with lower fluxes. As
Module B (run 2)
Fucose Rhamnose
Module A (run 2)
Arabinose Galactose
Module B (run 1) Glucose Mannose
Module A (run 1) 0
20
40 60 Percent
80 100
Figure 48 Composition of monosaccharides in foulants.
shown in Figures 48 and 49, it was found that the composition of foulants that accumulated in the two modules differed despite the fact that the two modules filtered the same suspension at the same time. These differences can be discovered by comprehensive measurements of carbohydrates and protein. FTIR spectra of the foulants obtained from the fouled membranes are presented in Figure 50. Difference in characteristics of the foulant caused by the difference in membrane flux was clearly shown in this analysis and a good reproducibility of the analysis is shown in Figure 50. In both runs, the foulants desorbed from the membranes operated at higher fluxes showed more pronounced peaks around 1550 cm1 than those operated at lower fluxes. Peaks around 1550 cm1 in an FTIR spectrum are assigned to amide groups. Figure 51 shows 13C NMR spectra of the desorbed foulants. Similar to the FTIR spectra, reproducibility of the analysis was confirmed. In the case of NMR analysis, a significant difference between the two modules can be found in the peaks at 105 ppm, which is attributed to anomeric carbon, and minor differences were repeatedly found in the region between 130 and 160 ppm, which corresponds to aromatic carbon. As described above, characteristics of foulants in MBRs differed significantly when membrane flux was different. This difference in characteristics of the foulants associated with difference in membrane flux can probably be explained by the size distribution of organic matter in the mixed liquor suspension of the MBR. Organic matter in the suspension exists in a variety of size distributions. It is likely that some types of organic matter tend to exist as large particles/molecules, while others exist as small ones. A particle/molecule never causes membrane fouling unless it is transported to the membrane surface. In the case of submerged MBRs, whether a particle/ molecule can reach the membrane surface or not, is totally
PHE LEU ILE MET VAL PRO ALA GLY LYS ARG HIS TYR CYS
Module B (run 2)
THR SER GLU ASP
Module A (run 1)
Module A (run 2) Module B (run 1)
0
2
4
6
8 Percent
Figure 49 Amino acid composition in foulants.
53
10
12
14
54
Membrane Filtration in Water and Wastewater Treatment
Module A (run 1)
2000
1800
1600
1400
1200
1000
800
600
Module A (run 2)
2000
1800
Wave number (cm–1)
1600
1400
1200
1000
Module B (run 1)
2000
1800
1600
1400
1200
800
600
Wave number (cm–1)
1000
800
600
Module B (run 2)
2000 1800
Wave number (cm–1)
1600
1400
1200
1000
800
600
Wave number (cm–1)
Figure 50 Fourier transform infrared (FTIR) spectra of foulants in fouled membranes.
Module A (run 1)
250
200 150 100 50 Chemical shift (ppm)
Module A (run 2)
0
250
200 150 100 50 Chemical shift (ppm)
0
Module B (run 2)
Module B (run 1)
250
200 150 100 50 Chemical shift (ppm)
0
250
200 150 100 50 Chemical shift (ppm)
0
Figure 51 Nuclear magnetic resonance (NMR) spectra of desorbed foulants.
dependent on a balance between the rate of convection flow toward the membrane associated with suction pressure (i.e., membrane filtration) and the rate of back-transport from the membrane mainly provided by turbulence caused by aeration. The size of a particle/molecule plays a key role in determining back-transportation rate. In this study, rates of convection flow toward the membranes were different in the two modules, and consequently different types of constituents could be transported to each module despite the fact that the two modules filtered the same suspension. As a result, characteristics of the membrane foulants in the two modules could become different. A high membrane flux can attract an increasing fraction of particles/ molecules to the membrane by overwhelming back-transportation rate. Larger particles/molecules with larger back-transportation rates are then pulled to the membranes and might
cause fouling when membrane flux is high. Among large particles/molecules in the mixed liquor suspension, there seem to be fractions that would cause severe membrane fouling. This would be a good explanation for the results showing that the degree of physically irreversible fouling caused by filtration of a specific volume of suspension differed considerably depending on membrane flux.
4.02.2.2.2 Effect of membrane material on fouling Two different bunches of hollow-fiber membranes made from different polymers (i.e., PE and PVDF) were separately bundled and submerged in a single reactor side by side. Both tested membranes had the same nominal pore size (0.4 mm). Total surface areas of the PE and PVDF membranes were 3 and 1.3 m2, respectively. These membranes were kindly supplied
Membrane Filtration in Water and Wastewater Treatment
Concentration (mg l–1)
by Mitsubishi Rayon Co., Ltd. (Tokyo, Japan). Filtration was carried out with the constant flow rate mode of operation using suction pumps. Membrane permeate flux was fixed at 0.4 m3 m2 d1 for both membranes. Intermittent filtration (12-min filtration and 3-min pause) was also carried out. MLSS concentration in the reactor was maintained at 11 g l1 by a daily extraction of excess sludge. Aeration was continuously carried out in the reactor at the flow rate of 3.5 m3 h1. Hydraulic retention time and solid retention time were 6.1 h and 34 days, respectively. When membrane fouling became significant, membrane modules were taken out from the reactor and were cleaned physically or chemically. Physical membrane cleaning was carried out by spraying pressurized water on the membrane surface. Chemical membrane cleaning was carried out by submerging the membrane modules in a solution of sodium hypochlorite (500 ppm) and hydrochloric acid (pH 2). The degree of membrane fouling was evaluated by membrane filtration resistance (Rt). Figure 52 shows time course changes in TOC concentrations measured in permeates from both membranes. There was no obvious difference in TOC concentrations in the permeates throughout the operation. It was confirmed that the difference in pore sizes of the two membranes was negligible. Figure 53 shows time course changes in Rt observed in the
pilot run. In the early stage of the operation, Rt of the PE membrane increased much faster than that of the PVDF membrane. The rate of increase in Rt of the PE membrane was fairly constant. Chemical cleaning of the PE membrane was carried out on day 76. The efficiency of the cleaning was so high that almost complete recovery of membrane permeability was observed. After the chemical cleaning, filtration using the PE membrane was restarted and Rt of the PE membrane increased at a rate that was comparable to that observed before the cleaning. Physical cleaning was carried out for the PE membrane at the end of operation (day 140) and was found to be ineffective. Almost no recovery of membrane permeability was achieved by the physical cleaning. This means that membrane fouling in the PE membrane that occurred after the chemical cleaning was irreversible fouling. Increase of Rt of the PVDF membrane was minimal in the early stage of the operation. Around day 60, Rt of the PVDF membrane suddenly started to increase. Physical cleaning of the PVDF membrane was carried out on day 89. Before the physical cleaning, accumulation of a gel-like substance that was considerably different from ordinary sludge cake was observed on the membrane surface by visual inspection. Rt of the PVDF membrane was substantially reduced by the physical cleaning and the operation was restarted. A rapid increase in Rt of the PVDF membrane was just after the physical cleaning. This rapid increase in Rt might have been due to changes in properties of the mixed liquor during that time (details given later). At the end of the operation (day 140), physical cleaning of the PVDF membrane was carried out and resulted in substantial reduction in filtration resistance. This was in contrast to the results obtained for the PE membrane. The filtration resistance that accumulated in the PVDF membrane was thought to be mainly reversible resistance. In Figure 53, evolution of irreversible fouling in the PVDF membrane can be estimated by connecting the points recorded just after the implementation of physical cleaning. As mentioned above, the increase in Rt of the PE membrane observed after day 76 directly reflected the evolution of
10 PE
PVD
5
0
0
20
40
60 80 100 Elapsed time (days)
120
140
Figure 52 Time course changes in total organic carbon (TOC) concentration in permeates.
MLSS = 10 g l–1 F/M = 20.4 SRT = 34 day
Total filtration resistance (1011 m–1)
Flux = 0.4 m d–1
PVDF
PE
10
Physical cleaning Chemical cleaning
9
55
PE PVDF
PE PVDF
8 7 6 5 PE
4 3 2
PVDF
1 0
0
20
40
80 100 60 Elasped time (days)
120
140
Figure 53 Time course changes in Rt (effect of membrane materials). The type of fouling was differed depending on membrane materials. PE: physically irreversible fouling was dominant; PVDF: most of the fouling observed for the PVDF membrane was physically reversible. MLSS, mixed liquor-suspended solid; SRT, solid retention time.
56
Membrane Filtration in Water and Wastewater Treatment
irreversible fouling. A comparison of the rates of increase in irreversible resistance observed in the two membranes showed that the rate of increase in the PE membrane was much faster. It should be noted that two membranes, having the same pore sizes, simultaneously filtered the same mixed liquor at the same permeate flux in the pilot run. Thus, the observed difference can be attributed to the properties of polymer materials. It can be concluded that PVDF is superior to PE in terms of prevention of irreversible fouling in MBRs used for treatment of municipal wastewater. In this study, the nominal pore size of both membranes was 0.4 mm. Particles with sizes close to the nominal pore size were assumed to affect membrane fouling. Thus, changes in the concentration of organic particles that were smaller than 1 mm (denoted hereafter as sub-micron-sized organic matter) were monitored. Whole sub-micron-sized organic matter was further divided into four fractions: between 0.65 mm and 1 mm, between 0.45 mm and 0.65 mm, between 0.1 mm and 0.45 mm, and less than 0.1 mm. This fractionation was carried out by successive filtrations using membrane filters with different pore sizes. Figure 54 shows the time course changes in the concentrations of each fraction in the mixed liquor of the pilot-scale MBR. As can be seen from Figure 54, between day 60 and day 100, the concentration of organic matter with particle size between 0.1 and 0.45 mm increased remarkably, although the concentration of TOC in both permeates was fairly constant. Possible reasons for this increase in sub-micron-sized organic matter are low temperature and/or increase in feed concentration. During the period between day 60 and day 100, a gradual decline in temperature and an increase in TOC concentration in the feed were observed at the same time. These changes might be responsible for the increase in concentration of organic matter with particle size between 0.1 and 0.45 mm at that time. It has been reported that production of soluble microbial product (SMP) is promoted as the organic loading rate increases. More accumulation of data is needed to identify the factors causing changes in characteristics of mixed
liquor in MBRs, which are thought to be related to membrane fouling in MBRs. In the case of the PVDF membrane in pilot run, as stated before, reversible fouling became significant during this period. On the other hand, in the case of the PE membrane, change in the rate of increase in filtration resistance was not significant during this period. These observations imply that organic matter that accumulated in the reactor during the period between day 60 and day 100 had a feature to promote formation of a cake layer (i.e., reversible fouling) on the PVDF membrane but that it did not affect fouling in the PE membrane. It is possible that features of the foulant differ depending on the membrane polymer material. Figures 55 and 56 show the changes in carbohydrate and protein, respectively. In these figures, data measured for permeates from both membranes and for mixed liquor in the MBR are shown. With respect to mixed liquor, measurements were carried out after filtering samples with a 0.5-mm filter. Concentrations of carbohydrate and protein in the permeates were relatively constant throughout the operation. The concentration of dissolved carbohydrate in the reactor significantly increased during the period between day 60 and day 100, when the concentration of sub-micron-sized organic matter increased (see Figure 54). At that time, increase in dissolved protein concentration in the mixed liquor was insignificant. Therefore, the increase in sub-micron-sized organic matter in the mixed liquor was likely to be due to the increase in carbohydrate. It is possible that carbohydrate accumulating between 60 and 100 days had different features from those observed in other periods. Unfortunately, however, the phenol–sulfuric acid method used for measurement of carbohydrate does not provide any information about features/composition of carbohydrate. To investigate the changes in composition of carbohydrate that accumulated in the reactor during the operation, monosaccharide composition analysis was conducted. Figure 57 shows the change in monosaccharide composition of dissolved organic matter in the MBR.
Flux = 0.4 m d–1 MLSS = 10 g l–1 F/M = 20.4 SRT = 34 day
Total filteration resistance (1011 m–1)
20 0.65 – 1.0 µm 0.45–1.0 µm 0.1–1.45 µm < 0.1 µm
15
10
5
0 0
20
40
60 80 100 Elasped time (days)
120
140
Figure 54 Time course changes in concentration of each fraction in mixed liquor (submicron-sized organic matter – effect of membrane materials). When the organic matter with particle size of 0.1–0.65 mm increased: PVDF – rapid increase in physically reversible fouling was observed; and PE – there is no obvious change in the rate of membrane fouling. MLSS, mixed liquor-suspended solid; PE, polyethylene; PVDF, polyvinylidene fluoride; SRT, solids retention time.
Membrane Filtration in Water and Wastewater Treatment 2.5 Dissolved PE permeate PVDF permeate
30
Concentration (mg l–1)
Concentration (mg l–1)
40
20
10
0
20
40
60 80 100 Elapsed time (days)
120
140
2 1.5 1 0.5
35
Dissolved PE permeate
30
PVDF permeate
25 20 15 10 5 0
20
40
60 80 100 Elapsed (days)
120
0
20
40
80 100 60 Elapsed time (days)
120
140
Figure 57 Changes in monosaccharide composition in dissolved organic matter.
40 Concentration (mg l–1)
Fucose Rhamnose Arabinose Galactose Glucose Mannose
0 0
Figure 55 Changes in carbohydrate. PE, polyethylene; PVDF, polyvinylidene fluoride.
0
57
140
Figure 56 Changes in protein. PE, polyethylene; PVDF, polyvinylidene fluoride.
From Figure 57, it can be seen that all types of monosaccharide exhibited an upward trend after day 60. Fucose, arabinose, and glucose increased significantly. Carbohydrates that were mainly composed of those monosaccharide increased during the period between day 60 and day 100 and might have caused reversible fouling in the PVDF membrane. On the other hand, it seemed that those carbohydrates did not affect membrane fouling in PE membrane as the rate of fouling in the PE membrane was fairly constant regardless of increase of those constituents at that time. To investigate features of constituents responsible for irreversible fouling that had occurred in the pilot run, organic matters were desorbed from the fouled membranes at the end of the continuous operation and were then analyzed. When the pilot operation was terminated, both the PE and the PVDF membrane modules were taken out from the reactor and were disassembled. Membrane fibers were rinsed with tap water and were manually wiped with a lab paper. This was done to remove accumulated cake that could be physically removed, allowing us to eliminate the bias caused by reversible fouling and to focus on the irreversible membrane fouling. Desorption of organic matter from the fouled membranes was carried out by soaking the membranes in an alkaline solution (sodium hydroxide) at 25 1C for 24 h. The solution pH was set at 11.
Analysis of organic matter desorbed from the fouled membranes was conducted at the termination of the continuous operation, and the difference between the PE and the PVDF membranes was investigated. As described in the experimental section, intensive physical cleaning was carried out prior to desorption of foulants from the membranes. Thus, the following discussion will be made for filtration resistance that cannot be cancelled by physical cleaning (i.e., irreversible fouling). Table 4 shows the results of analysis in the desorption test. The data shown in Table 4 are expressed on the basis of a unit surface membrane area (mg m2). As stated before, at the end of continuous operation, irreversible fouling resistance in the PE membrane was greater than that in the PVDF membrane. Nevertheless, a greater amount of organic matter was desorbed from the PVDF membrane than from the PE membrane on the basis of a unit membrane surface area. One possible explanation for this is that features/compositions of the organic matters desorbed from the two membranes were different and consequently the magnitude of fouling differed on the basis of a unit mass of organic matter. If this is the case, the data shown in Table 4 suggest that organic matter desorbed from the PE membrane had a higher fouling potential than that desorbed from the PVDF membrane. SUVA and carbohydrate/protein ratio (C/P) determined for the organic matters desorbed from the two membranes are also shown in Table 4. These two indexes are lumped ones and therefore provided limited information on features of organic matter. Nevertheless, the difference between the organic matters desorbed from the two membranes in terms of SUVA and C/P was obvious, and it can therefore be assumed that features of the foulant causing irreversible resistance differ depending on the membrane polymer material.
4.02.2.2.3 Fouling potential of carbohydrate assessed by lectin affinity chromatography In a number of previous studies, carbohydrates have been pointed out to be major foulants in MBRs. The authors’ previous study indicated that some fractions of organic matter contained in the mixed liquor of MBRs would cause severe membrane fouling than do other fractions. They also reported
58
Membrane Filtration in Water and Wastewater Treatment
Table 4
Characteristics of organic matter desorbed from fouled membranes
Membrane
TOC(mg m2)
Carbohydrate (mg m2)
Protein (mg m2)
Carbohydrate/Protein
UVA260 (cm1)
SUVA (m1 mg1 l)
PE PVDF
3.3 9.8
3.3 8.3
8.5 11.6
0.38 0.72
0.092 0.142
3.17 2.45
PE, polyethylene; PVDF, polyvinylidene flouride; SUVA, specific ultraviolet absorbance; TOC, total organic carbon; UVA260, ultraviolet absorbance at 260 nm.
Mixed liqour of MBR Filtration (0.45 µm) Dissolved organic matter
Control
Removal of polysaccharides by lectin affinity column
Effluent of each column Dead-end filtration test Figure 58 Experimental procedure. MBR, membrane bioreactor.
that composition of foulants was different depending on the membrane material implying that characteristics of organic matter that caused severe membrane fouling are closely related to characteristics of membrane. At present, however, little is known about the details of carbohydrates that cause membrane fouling in MBRs. Information on characteristics of the organic matters causing severe membrane fouling in relation to the types of membrane should be very useful for establishing a strategy to mitigate membrane fouling in MBRs (Miyoshi et al., 2010; Yamamura et al., 2008). In this study, affinity chromatography was applied with a variety of lectins to assess the fouling potential of some specific carbohydrates contained in mixed liquor suspension in an MBR. Lectins are proteins with the ability to bind specific carbohydrates with high selectivity. By changing the types of lectin in affinity chromatography, different types of carbohydrates can be removed from the liquid phase. After that, reductions in the degree of fouling potential associated with the removal of specific carbohydrates by lectins were evaluated. This investigation was carried out for different membrane and the results were compared with each other in order to assess the effects of membrane materials on fouling potential of carbohydrates. In addition, organic matters retained in lectin columns were eluted and were then characterized. Based on the results obtained in this study, factors associated with difference in fouling potential of specific carbohydrates caused by difference in membrane materials will be discussed. Continuous operation of a pilot-scale MBR was conducted at Soseigawa Municipal Wastewater Treatment Center in Sapporo, Japan. A pilot-scale MBR was operated with baffled MBR (BMBR) configuration (Kimura et al., 2008b). Hydraulic retention time (HRT) and solid retention time (SRT) were set
Table 5
Lectins used in assessment of fouling potential
Lectin
Binding specificity
Aleuria aurantia lectin (AAL) Concanavalin A (Con A)
al–6, al–2, al–3 Fuc
Datura stramonium agglutinin (DSA) Lens culinaris agglutinin (LCA) Maackia amurensis lectin (MAM) Ricinus communis agglutinin (RCA) Sambucus sieboldiana agglutinin (SSA) Wheat germ agglutin (WGA)
a-D-Man in type N-glycans hybrid type high Man bianntenary and hexaantennary, All O-types, a-Glu Galbl–4GlcNAc, GlcNAc a-D-Man in di- and tri-complex-type Nglycans with core a-Fuc, a-Ghi NeuAca2–3Gal b-Gal, GalNAc NeuAca2-6Gal/GaINAc b-GlcNAc, sialic acid, GlcNAcbl, 4GlcNAcbl, 4GlcNAc, chitoriose
From Kaku H et al. (2007), Journal of Biochemistry 142: 393–401; Opitz L, et al. (2008) Vaccine 25: 939–947; Greenwel P, et al. (2008) International Journal of Pharasitology 38: 749–756.
around 2.9 h and 35 days, respectively. As a result, MLSS concentration in the reactor was 16.471.4 g l1. Solution containing dissolved organic matter was obtained by centrifugation (4800 rpm; 5 min) followed by filtration using a membrane filter paper with a pore size of 0.45 mm. Figure 58 shows the experimental procedure employed in this study. Carbohydrates with high affinity to the lectin in the column were retained. As a result, filtration resistance should be lowered in the subsequent filtration test if the retained carbohydrates had high fouling potentials. Evaluation of fouling potential was also conducted for the solution without passage through a lectin column as a control test. Commercially available pre-packed lectin–agarose columns (Seikagaku Corporation, Tokyo, Japan) were used in this study. Table 5 lists the lectins used in this study and their binding specificities (Kaku et al., 2007; Opitz et al., 2008; Greenwel et al., 2008). All columns were washed with 60 times of gel volume of phosphate buffer saline pH 7.2 (PBS) and then 50 ml of sample solution containing dissolved organic matter in the mixed liquor of MBR was loaded. The liquid which passed through a lectin column was collected and subsequently applied to dead-end filtration test to evaluate the fouling potential. After passage of sample solution, the column was washed with 10 times of gel volume of PBS to remove nonbinding constituents. Bound constituents were eluted with PBS containing elution reagent and were then analyzed. The elution reagents used for each lectin are
Membrane Filtration in Water and Wastewater Treatment
summarized in Table 6. In this study, characterization of organic matter eluted from lectins in terms of characteristics of sugar (e.g., monosaccharide composition) was not performed. This is because some lectins have monosaccharide or disaccharide as elution reagent (Table 6) which can interfere with the results of sugar analysis. In this study, organic matters eluted from lectins were characterized by means of excitation–emission matrices (EEMs) because this method does not suffer from interference by elution reagents. Dead-end filtration tests were conducted for the assessment of fouling potential of polysaccharides in mixed liquor suspension in an MBR. Flat-sheet membranes were used for the dead-end filtration test. In this study, two different membranes were used in dead-end filtration test. One was made of polypropylene (PP; Kubota, Osaka, Japan) and the other was made of PVDF (Toray, Tokyo, Japan). Nominal pore size values of PP and PVDF membrane were 0.4 and 0.1 mm, respectively. Operating pressures of PP and PVDF membrane were 20 and 30 kPa, respectively. These values were selected to equalize Table 6
Elution reagent for each lectin
Elution reagent
Lectin
L-Fucose
AAL
Methyl-a-D-glucoside Chitooligosaccharide Ethylenediamine Lactose N-Acetyl-D-glucosamine
Con A, LCA DSA MAM RCA, SSA WGA
From Kaku H et al. (2007), Journal of Biochemistry 142: 393–401.
60
59
pure water permeation flux of new membrane for both membranes. Effective surface areas of the membranes were 13.9 cm2. Pure water permeability was measured for the membranes before and after filtration of 5 ml of sample solution (i.e., effluents from the lectin columns), and the difference between them was used to evaluate the fouling potential of carbohydrates. Pure water permeability was evaluated by membrane filtration. Figure 59 shows filtration resistance developed in the filtration of sample solutions. Regardless of the types of membrane, some lectins were effective for reduction in filtration resistance while the others were not. This indicates that the fouling potential differed depending on the types of carbohydrates. Among the tested lectins, wheat germ agglutin (WGA) and concanavalin A (Con A) which were used for visualization of carbohydrate involved in membrane fouling, were not effective for reduction in fouling potential regardless of the types of membrane. The results obtained in this study clearly indicate that carbohydrates which can be visualized by those lectins do not represent all of the membrane fouling. WGA recognizes N-acetyl-glucosamine (Table 5), which is the major component of peptide glycan in the cell wall of bacteria. It was likely that the contribution of the carbohydrates derived from cell debris to membrane fouling was relatively low compared to those with other origins. Aleuria aurantia lectin (AAL) was effective for the reduction of fouling potential for both membranes. Carbohydrates which have higher affinity to this lectin would cause severe membrane fouling at least in the two membranes tested in this study. In contrast, some lectins reduced fouling potential for only one type of membrane. For example, Sambucus sieboldiana agglutinin (SSA) and Maackia amurensis lectin (MAM) were
PVDF membrane
50 40
Filtration resistance (1011 m–1)
30 20 10 0 Control 60
WGA
RCA
LCA
Con A
SSA
AAL
MAM
DSA
RCA
LCA
Con A
SSA
AAL
MAM
DSA
PP membrane
50 40 30 20 10 0 Control
WGA
Figure 59 Filtration resistance development. AAL, Alueria aurantia agglutinin; Con A, concanavalin A; DSA, Datura stramonium agglutinin; LCA, Lens culinaris agglutinin; MAM, Maackia amurensis agglutinin; PP, polypropylene; PVDF, polyvinylidene fluoride; SSA, Sambucus sieboldiana agglutinin; RCA, Ricinus communis agglutinin; WGA, wheat germ agglutinin.
Membrane Filtration in Water and Wastewater Treatment
300
400 350 300
450
Excitation (nm)
350
Excitation (nm)
Excitation (nm)
400 350 300 250
250
400 350 300 250
250 300 350 400 450 500
250 300 350 400 450 500
250 300 350 400 450 500
250 300 350 400 450 500
Emission (nm)
Emission (nm)
Emission (nm)
Emission (nm)
SSA
AAL Excitation (nm)
400 350 300
450
400 350 300 250
250
DSA
MAM
450
450
450
Excitation (nm)
Excitation (nm)
400
Con A
450
450
250
Excitation (nm)
LCA
RCA
WGA 450
Excitation (nm)
60
400 350 300 250
400 350 300 250
250 300 350 400 450 500
250 300 350 400 450 500
250 300 350 400 450 500
250 300 350 400 450 500
Emission (nm)
Emission (nm)
Emission (nm)
Emission (nm)
Figure 60 Excitation–emission matrix (EEM) spectra of organic matter eluted from lectin columns.
effective for the reduction of fouling potential in PP membrane while reductions in fouling potential by those lectins were negligible in the case of PVDF membrane. On the other hand, effectiveness of Ricinus communis agglutinin (RCA), which was slightly effective for the reduction of fouling potential in PVDF membrane, was not recognized in PP membrane. Those results clearly indicate that types of carbohydrates that cause severe membrane fouling were different for two membranes examined in this study. A fouling potential of specific carbohydrate would depend heavily on characteristics of a membrane. Since many carbohydrates found in activated sludge system coexist with other organics, characteristics of the organic matter associated with the carbohydrates are thought to be important for fouling potential of retained carbohydrates. To investigate the characteristics of the organic matter associated with the carbohydrates retained by lectins, EEM fluorescence spectra analysis was applied to the organic matters eluted from lectins. Figure 60 shows EEM fluorescence spectra of the organic matter eluted from lectin columns. Basically, shapes of EEMs obtained for different samples were different indicating that types of organic matter retained by lectin were different depending on the types of lectins. However, the shapes of EEMs obtained for the organic matters eluted from Con A and LCA were similar. Both EEMs have two major peaks located on Ex/ Em ¼ 325 nm/425 nm and 275 nm/425 nm. These two peaks can be attributed to humic acid-like substances. Similarity in the shapes of EEMs can also be seen between the organic matter eluted from RCA and SSA. Those EEMs have one major peak around Ex/Em ¼ 325 nm/400 nm that can be attributed to humic acid-like substances that have features similar to those found in ocean and one minor peak around Ex/ Em ¼ 275 nm/425 nm. As can be seen in Table 6, elution reagent of organic matter for Con A and LCA is the same (methyl-a-D-glucoside). RCA and SSA also have the same elution
reagent (lactose). It can be assumed that lectins which have the same elution reagent have high affinity to similar structure of polysaccharides. As a result, carbohydrates adsorbed onto Con A and LCA or RCA and SSA were likely to overlap to some extent. The similarities in the shapes of EEMs obtained for the organic matter eluted from Con A and LCA or RCA and SSA could partly be explained by those overlapping. However, the similarities in the shapes of EEMs did not coincide with the degree of reductions in fouling potentials by each lectin. For example, LCA was slightly effective for reduction in fouling potential in PVDF membrane but Con A was not. SSA was effective for PP membrane but was not for PVDF membrane. In contrast, effectiveness of RCA on reduction in fouling potential can only be recognized in PVDF membrane. Those results clearly indicate that fouling potentials are not related to characteristics of organic matters associated with carbohydrates retained by lectins. Difference in structures or properties of sugar chain which could not be assessed by EEM analysis would play an important role in determining fouling potentials. Taking into consideration the results of dead-end filtration test that indicated an influence of membrane materials on fouling potential of specific carbohydrate, investigation on the relationship between characteristics of sugar chain in the polysaccharides which have high fouling potentials and characteristics of membrane (e.g., surface morphology, roughness) will be important to elucidate the interactions between carbohydrates and membranes. Further studies regarding this point are needed.
References Greenwel P et al. (2008) International Journal of Pharasitology 38: 749–756. Jang N-Y, Watanabe Y, and Minegishi S (2004) Performance of ultrafiltration membrane process combined with coagulation/sedimentation. Water Science and Technology 51(6–7): 209--219.
Membrane Filtration in Water and Wastewater Treatment
Kaku H et al. (2007) Journal of Biochemistry 142: 393–401. Kimura K, Nishisako R, Miyoshi T, Shimada R, and Watanabe Y (2008a) Baffled membrane bioreactor (BMBR) for efficient nutrient removal from municipal wastewater. Water Research 42: 625–632. Kimura K, Miyoshi T, Naruse T, and Watanabe Y (2008b) The difference in characteristics of foulants in submerged MBRs caused by the difference in membrane flux. Desalination 231(1–3): 268--275. Kimura K, Yamato N, Yamamura H, and Watanabe Y (2005) Membrane fouling in pilotscale membrane bioreactors (MBRs) treating municipal wastewater. Environmental Science and Technology 39: 6293--6299. Lee S, Jang N, and Watanabe Y (2004) Effect of residual ozone on membrane fouling reduction in ozone resisting microfiltration (MF) membrane system. Water Science and Technology 50(12): 287--292. Miyoshi T, Tsuyuhara T, Aizawa E, Kimura K, and Watanabe Y (2010) Fouling potentials of polysaccharides in MBRs assessed by lectin affinity chromatography. Water Science and Technology 62. Opitz L et al. (2008) Vaccine 25: 939–947. Suzuki T, Watanabe Y, Ozawa G, and Ikeda S (1998) Removal of soluble organics and manganese by a hybrid MF hollow fibre membrane system. Desalination 117(1–3): 119--129.
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Yamamura H, Chae S-R, Kimura K, and Watanabe Y (2007a) Transition in fouling mechanism in microfiltration of a surface water. Water Research 41: 3812--3822. Yamamura H, Kimura K, Okajima T, Tokumoto H, and Watanabe Y (2008) Affinity of functional groups for membrane surfaces: Implications for physically irreversible fouling. Environmental Science and Technology 42(14): 5310--5315. Yamamura H, Kimura K, and Watanabe Y (2007b) Mechanism involved in the evolution of physically irreversible fouling in microfiltration membranes used for drinking water treatment. Environmental Science and Technology 41(19): 6789--6794. Yamato N, Kimura K, Miyoshi T, and Watanabe Y (2006) Difference in membrane fouling in membrane bioreactors (MBRs) caused by membrane polymer materials. Journal of Membrane Science 280(1–2): 911--919. Yonekawa H, Tomita Y, and Watanabe Y (2004) Behavior of micro-particles in monolith ceramic membrane with pre-coagulation. Water Science and Technology 50(12): 317--325. Watanabe Y and Yonekawa H (2008) Flocculation and its inclusion into membrane filtration. Lecture note of Academic Summer School on particle separation in water and wastewater treatment, organized by IWA Specialist Group on Particle Separation (in press as a book published by IWA publishing).
4.03 Wastewater Reclamation and Reuse System HL Leverenz and T Asano, University of California at Davis, Davis, CA, USA & 2011 Elsevier B.V. All rights reserved.
4.03.1 4.03.2 4.03.3 4.03.4 4.03.5 4.03.6 4.03.6.1 4.03.6.2 4.03.6.3 4.03.6.4 4.03.6.5 4.03.6.6 4.03.7 4.03.7.1 4.03.7.2 4.03.7.3 4.03.8 References
Foundation of Water Reuse Water Reuse Terminology and Definitions Reclaimed Water Applications Water-Quality Considerations Treatment Technology Infrastructure for Water Reuse Storage Facilities Distribution Systems Centralized Systems Decentralized Systems Satellite Systems Point-of-Use Treatment Source Control Salinity and Toxic Constituents Source Separation Graywater Future Directions for Water Reuse
63 63 63 65 65 67 67 67 68 69 69 69 70 70 70 70 70 71
4.03.1 Foundation of Water Reuse
4.03.2 Water Reuse Terminology and Definitions
Inadequate water supplies and water-quality deterioration represent serious contemporary concerns for many municipalities, industries, agriculture, and the environment in various parts of the world. Several factors have contributed to these problems such as continued population growth in urban areas, contamination of surface water and groundwater, uneven distribution of water resources, and frequent droughts caused by extreme global weather patterns. Water reclamation and reuse accomplishes two fundamental functions: (1) the treated effluent is used as a water resource for beneficial purposes, thereby reducing potable water demands and, (2) where effluent is returned to the environment, improves overall water quality in the receiving water, which is often used subsequently as potable water supply and habitat. The foundation of water reuse is built upon three principles: (1) providing reliable treatment of wastewater to meet strict water-quality requirements for the intended reuse applications, (2) protecting public health, and (3) gaining public acceptance. Whether water reuse is appropriate for a specific locale depends upon careful economic considerations, potential uses for the reclaimed water, and the relative stringency of waste discharge requirements. Public policies can be implemented that promote water conservation and reuse rather than the costly development of additional water resources with considerable environmental expenditures. Through integrated water resources planning, the use of reclaimed water may provide sufficient flexibility to allow a water agency to respond to short-term needs as well as increase the reliability of long-term water supplies (Asano and Levine, 1995).
Early developments in the field of water reuse are synonymous with the historical practice of land application for the disposal of wastewater. With the advent of sewerage systems in the nineteenth century, domestic wastewater was used at sewage farms and by 1900 there were numerous sewage farms in Europe and in the United States. While these sewage farms were used primarily for waste disposal, incidental use was made of the water for crop production or other beneficial uses. During the past century, the growing need for reliable water has resulted in the development of a number of water reclamation and reuse projects. Wastewater reclamation is the treatment or processing of wastewater to make it reusable, and water reuse is the use of treated wastewater for beneficial purposes such as agricultural irrigation and industrial cooling. Reclaimed water is a treated effluent suitable for an intended water reuse application. In addition, direct water reuse requires the existence of pipes or other conveyance facilities for delivering reclaimed water. Indirect reuse, through discharge of an effluent to receiving water for assimilation and withdrawals downstream, is recognized to be important but does not constitute planned direct water reuse. In contrast to direct water reuse, water recycling normally involves only one use or user and the effluent from the user is captured and redirected back into that use scheme. In this context, water recycling is predominantly practiced in industry. To facilitate communication among different groups associated with water reuse, it is important to understand the terminology used in the arena of water reclamation and reuse. Water reclamation and reuse definitions commonly used are summarized in Table 1.
63
64
Wastewater Reclamation and Reuse System
Table 1
Water reuse terminology and definitions
Term
Definition
Agricultural water use Aquifer Beneficial uses
Water used for soil cultivation, crop production, and livestock uses. Geological formations that contain and transmit groundwater. Many ways water can be used, either directly by people, or for their overall benefit. Examples include municipal water supply, agricultural and industrial applications, navigation, fish and wildlife, and water contact recreation. 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; also referred to as water consumed. Incorporation of reclaimed water directly into a potable water supply system. Domestic water use includes water for normal household purposes, such as drinking, food preparation, bathing, washing clothes and dishes, flushing toilets, and watering lawns and gardens. A collective term that includes loss of water from the soil by evaporation and by transpiration from plants. The infiltration or injection of natural waters or reclaimed waters into an aquifer, providing replenishment of the groundwater resource or preventing seawater intrusion. Planned potable reuse indirectly allowing mixing and assimilation by discharge into an impoundment or natural body of water, such as in domestic water supply reservoir or groundwater. Water in industry is used for cooling, transportation, as a solvent, and as an ingredient of the finished products. The principal water users in industry are thermal and atomic power generation. Artificial application of water on lands to assist in the growing of crops and pastures or to maintain vegetative growth in recreational lands such as parks and golf courses. Turf and landscape irrigation systems in ways that enable the efficient and safe application of reclaimed water in such places as golf courses, public parks, playgrounds, school yards, and athletic fields. The water withdrawals made by the populations of cities, towns, and housing estates, and domestic and public services and enterprises. Also includes water used to directly provide for the needs of urban populations, which consume high-quality water from city water supply systems. All water reuse applications that do not involve either indirect or direct potable reuse. Water suitable for human consumption without deleterious health risks. The term, drinking water is a preferable term better understood by the community at large. An augmentation of drinking (potable) water directly or indirectly by highly treated reclaimed water. Municipal wastewater that has gone through various treatment processes to meet specific water-quality criteria with the intent of being used in a beneficial manner (e.g., irrigation). The term recycled water is used synonymously with reclaimed water, particularly in California. The process of tapping into a sewer main and extracting wastewater locally, which can then be treated in a satellite or decentralized treatment plant and reused for beneficial purposes. State of California regulations for how treated and recycled water is used and discharged that is listed in Title 22 of the California Administrative Code. The state-wide Water Recycling Criteria are developed by the Department of Health Services and enforced by the nine State Regional Water Quality Control Boards. Used water discharged from homes, business, cities, industries, and agriculture. Various synonymous uses such as municipal wastewater (sewage), industrial wastewater, and storm water. Treatment or processing of wastewater to make it reusable with definable treatment reliability and meeting appropriate water-quality criteria. The use of wastewater which is captured and redirected back into the same water use scheme such as in industry. However, the term ‘water recycling’ is often used synonymously with water reclamation. The use of treated wastewater for a beneficial use, such as agricultural irrigation and industrial cooling.
Consumptive use Direct potable reuse Domestic water use Evapo-transpiration Groundwater recharge Indirect potable reuse Industrial water use Irrigation water use Landscape irrigation Municipal water use
Nonpotable reuse Potable water Potable reuse Reclaimed water (also, recycled water) Sewer mining Title 22 regulations
Wastewater Water reclamation Water recycling Water reuse
4.03.3 Reclaimed Water Applications In the planning and implementation of water reclamation and reuse, the reclaimed water application generally governs the type of wastewater treatment needed to protect public health and the environment, and the degree of reliability required for each sequence of treatment processes and operations. In principle, wastewater or any marginal quality waters can be used for any purpose as long as adequate treatment is provided to meet the water-quality requirements for the intended use. The dominant applications for the use of reclaimed water include: agricultural irrigation, landscape irrigation, industrial recycling and reuse, and groundwater recharge. Among them, agricultural and landscape irrigation are widely practiced throughout the world with well-established health protection guidelines and agronomic practices.
From a global perspective, water reuse applications have been developed to replace or augment water resources for specific applications, depending on local water use patterns. In general, water reuse applications fall under one of seven categories: (1) agricultural irrigation, (2) landscape irrigation, (3) industrial reuse, (4) groundwater recharge, (5) environmental and recreational uses, (6) nonpotable urban uses, or (7) indirect or direct potable reuse. The relative amount of water used in each category varies locally and regionally due to differences in specific water use requirements and geopolitical constraints. Notable aspects of each of these water reuse applications are given in the following:
•
Agricultural irrigation represents the largest current use of reclaimed water throughout the world. This reuse
Wastewater Reclamation and Reuse System
•
•
•
•
•
•
category offers significant future opportunities for water reuse in both industrialized countries and developing countries. Landscape irrigation is the second largest user of reclaimed water in industrialized countries and it includes the irrigation of parks; playgrounds; golf courses; freeway medians; landscaped areas around commercial, office, and industrial developments; and landscaped areas around residences. Many landscape irrigation projects involve dual distribution systems, which consist of one distribution network for potable water and a separate pipeline to transport reclaimed water. The reclaimed water pipelines are normally color-coded with purple color in the United States. Industrial activities represent the third major use of reclaimed water, primarily for cooling and process needs. Cooling water creates the single largest industrial demand for water and as such is the predominant industrial water reuse either for cooling towers or for cooling ponds. Industrial uses vary greatly and water-quality requirements tend to be industry specific. To provide adequate water quality, supplemental treatment may be required beyond conventional secondary wastewater treatment. Groundwater recharge is the fourth largest application for water reuse, either via spreading basins or via direct injection to groundwater aquifers. Groundwater recharge includes groundwater replenishment by assimilation and storage of reclaimed water in groundwater aquifers, or establishing hydraulic barriers against salt-water intrusion in coastal areas. Recreational and environmental uses constitute the fifth largest use of reclaimed water in industrialized countries and involve nonpotable uses related to land-based water features such as the development of recreational lakes, marsh enhancement, and stream flow augmentation. Reclaimed water impoundments can be incorporated into urban landscape developments. Man-made lakes, golf course storage ponds, and water traps can be supplied with reclaimed water. Reclaimed water has been applied to wetlands for a variety of reasons, including habitat creation, restoration and/or enhancement, provision for additional treatment prior to discharge to receiving water, and provision for a wet weather disposal alternative for reclaimed water. Nonpotable urban uses include fire protection, air conditioning, toilet flushing, construction water, and flushing of sanitary sewers. Typically, for economic reasons, these uses are incidental and depend on the proximity of the wastewater reclamation plant to the point of use. In addition, the economic advantages of urban uses can be enhanced by coupling with other ongoing reuse applications such as landscape irrigation. Potable reuse is another water reuse opportunity, which could occur either by blending in water-supply storage reservoirs or, in the extreme, by direct input of highly treated wastewater into the water distribution system. Although the likelihood of implementing this option in the most locations is remote, a successful example includes the City of Windhoek, Namibia (Asano et al., 2007).
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4.03.4 Water-Quality Considerations The acceptability of reclaimed water for a given water reuse application is dependent on the physical, chemical, and microbiological quality of the water. The effects of physical parameters (such as pH, color, temperature, and particulate matter) and chemical constituents (such as chlorides, sodium, heavy metals, and trace organics) on vegetation, soil, and groundwater are well known, and recommended limits have been established for these constituents. In contrast to the agronomic considerations associated with chemical constituents that may be present in wastewater, pathogenic constituents may present health considerations for the distribution and use of reclaimed water. The recommended guidelines and regulations related to reclaimed water quality are found in State of California (2000), World Health Organization (2006), and Asano et al. (2007). Source control programs can limit the input of chemical and microbiological constituents that may present health, environmental, or irrigation concerns or that may adversely affect treatment processes and subsequent acceptability of the reclaimed water for specific uses. In some arid and semi-arid regions, the level of total dissolved solids is a major quality concern, and source control measures are considered for domestic water users such as restrictions on water softener use. Assurance of treatment reliability is an obvious, yet sometimes overlooked, quality control measure. Water-quality considerations in water reuse applications are extremely important especially where health and environmental issues are of concern. Unless the product water is of sufficient quality to meet the required criteria and regulations for the intended reuse application, acceptance by the potential users or beneficiaries will not occur. By the same token, over-treatment that is excessive for its intended use is a waste of resources in terms of energy, labor, equipment, and money.
4.03.5 Treatment Technology Treatment technologies used for the production of reclaimed water typically follow conventional secondary treatment. These technologies include depth and surface filters, membranes, carbon adsorption, disinfection, and advanced oxidation. The type of treatment processes that are selected to produce reclaimed water will depend on several factors, including the quantity and quality of reclaimed water required and the life cycle costs of the reclaimed water system. However, in terms of water quality, virtually any quality of reclaimed water that is desired can be produced using currently available technology. A summary of these technologies and the specific constituents that are removed is presented in Table 2. Additional details on the technologies described in Table 2 can be found in Asano et al. (2007) and Tchobanoglous et al. (2003). With further refinement and development, the cost and robustness of these technologies is improving. Membranes represent the most significant development as several new products are now available for a number of water and wastewater treatment and water reuse applications.
Table 2
Unit operations and processes used for the removal of classes of constituents found in wastewater for reuse applications
Unit operation or process
Constituent class Suspended solids
Secondary treatment Secondary with nutrient removal Depth filtration Surface filtration Microfiltration Ultrafiltration Dissolved air flotation Nanofiltration Reverse osmosis Electrodialysis Carbon adsorption Ion exchange Advanced oxidation Disinfection Natural processes Source control
Colloidal solids
Organic matter (particulate)
x x x x x x x
Dissolved organic matter x x
x x x
x x x x x x
Nitrogen
Phosphorus
x
x
Trace constituents
x x x x x x
x
x
x
x x x x x
x
x
x
x
Bacteria
x x x x
x
Protozoan cysts and oocysts
Viruses
Energy needs
þ þþ
x
x
Total dissolved solids
x x
x x x
x x
x x x x x x x
x x x x
x x
x x
x x
x
x
Modified from Asano T, Burton FL, Leverenz H, Tsuchihashi R, and Tchobanoglous G (2007) Water Reuse: Issues, Technologies, and Applications. New York: McGraw-Hill and Tchobanoglous et al. (2003).
þ þ þ þ þ þ þ þ þ þ þ þ
þ þ þ þþ þþ
þþ
Wastewater Reclamation and Reuse System
Membranes had been limited previously to water softening and desalination, but are now being used increasingly for wastewater applications to produce high-quality reclaimed water suitable for reuse. Treatment trains that incorporate membrane filtration are capable of producing several grades of product water that can serve a range of water reuse applications. Desalination of reclaimed water is also being done by means of reverse osmosis and electrodialysis. Increased levels of contaminant removal not only enhance the product water for reuse, but also lessen health risks. Further, the cost of producing high-quality reclaimed water has decreased considerably, largely due to the development of low-pressure membranes and the entrance of a number of suppliers in the competitive marketplace. Chlorination remains as the most widely used disinfection technology and its effectiveness is enhanced by improved reclaimed water quality. Increased removal of particulate matter and the development of ultraviolet disinfection technology also improve the applicability of reclaimed water for many more applications. Advanced oxidation is also an important technology for reducing or removing trace constituents and emerging contaminants to safe levels, especially for indirect potable water reuse applications. While not specifically identified in Table 2, natural treatment systems, including oxidation ponds, constructed wetlands, sand filtration, bank filtration, soil/vadose zone filtration, and anaerobic processes, are also used commonly in water reuse. In addition, practices such as urine segregation, graywater systems, and source control of specific constituents can have a significant impact on overall water quality and the type of treatment system that is required. The specific application of these technologies is site specific and, therefore, subject to local constraints. Natural treatment processes often experience a higher level of variability due to the reduced level of control that can be applied. However, in terms of energy usage, these processes can be used to accomplish water reuse with significantly lower power output. Thus, natural processes should be considered and implemented where feasible to reduce the overall carbon footprint of the water reuse system.
4.03.6 Infrastructure for Water Reuse In areas where reclaimed water is distributed widely, for example, in urban lawn irrigation projects, a substantial redistribution system is usually required. An alternative reclaimed water project that minimizes the installation of pipelines is where reclaimed water is added directly to the potable water reservoir. However, this type of reuse, indirect potable reuse, requires a higher level of treatment compared to water used for lawn irrigation but takes advantage of the preexisting infrastructure used to distribute potable water. Thus, the type of reclaimed water system must be evaluated carefully, depending on the level of treatment that is attainable reliably and the expense to install pipelines for the reclaimed water. The primary factors governing distribution and storage facilities include the location of the reclaimed water treatment plant and the location and demand requirements of the reclaimed water users. The principal facilities needed for the
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delivery of reclaimed water are storage tanks, pumping stations, and transmission and distribution pipelines and depend on the overall type of reclaimed water. Important issues and factors, typical to most reclaimed water distribution and storage projects, are planning and implementation issues, planning and conceptual design of distribution and storage facilities, design of pipelines, design of pumping facilities, operation and maintenance of pipelines and pumping stations, design of storage facilities, and operational issues in reclaimed water storage. Planning and implementation issues that must be addressed when considering storage and distribution facilities for a reclaimed water project include: 1. the type, size, and location of physical facilities; 2. the interrelationship between the potable and reclaimed water systems, that is, is the reclaimed water system being installed in an area where an existing potable water system exists or is a dual distribution system (for potable and reclaimed water) needed?; and 3. the involvement of the public during the planning and implementation process; the public may be affected directly by facilities siting and construction. A brief description of key infrastructure used for water reuse systems is presented in the following sections.
4.03.6.1 Storage Facilities Storage of reclaimed water is required in situations where there is a difference in the production and utilization of reclaimed water, such as in cases where water is stored for nighttime irrigation and where there is a seasonal use of reclaimed water such as agricultural irrigation. Elevated storage tanks are used to regulate the system pressure. The size of storage required should be determined from a flow analysis. Storage reservoirs can be above- or below-ground tanks, open reservoirs, or an aquifer. The type of storage to be used is determined by the site constraints and requirements of the application. Open reservoirs are subject to contamination from recreational and wildlife activities, thus requiring treatment prior to distribution for reuse. However, the open reservoir can also be viewed as an environmental buffer, which is used in cases of indirect potable reuse. The facilities to store and distribute the reclaimed water to potential users can be planned and designed once the source of reclaimed water and the location and nature of the water reuse areas and demands are known. In most respects, facilities for the storage and distribution of reclaimed water are similar to those for potable water. Because of the characteristics of reclaimed water and the potential changes in water quality that may occur over time, care must be taken during the planning, design, and operation of distribution and storage facilities to prevent or mitigate any effects.
4.03.6.2 Distribution Systems A reclaimed water system may be planned, designed, and installed as a system totally separate from the potable water system or planned as part of a dual distribution system that provides both reclaimed and potable water to the service area (see Figure 1). The integrated planning, design, and
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Wastewater Reclamation and Reuse System Remote community with independent collection and treatment system Remote community or development serviced by independent satellite reclamation plant connected to centralized collection system To local reuse Waste solids
Satellite reclamation plant
Trunk sewer
High-rise building (typical)
(c)
Satellite reclamation plant
(d)
Decentralized reclamation system
(e)
Centralized treatment facility
To local reuse Receiving water body Dual distribution system
Central collection system
Nonpotable in-building reuse
Screenings In-building reclaimed water distribution system
Waste solids Satellite reclamation plant
Effluent recycle (at low flow, if needed) To local reuse
To outdoor reclaimed water system Flow equalization tank
Central collection system
(b)
Waste to centralized system
Reclaimed water storage tank Onsite treatment process for water reclamation and reuse
(a)
Figure 1 Definition sketch for various types of water reuse systems: (a) interception-type satellite, (b) extraction-type satellite, (c) upstream-type satellite, (d) decentralized, and (e) centralized. Modified from Asano T, Burton FL, Leverenz H, Tsuchihashi R, and Tchobanoglous G (2007) Water Reuse: Issues, Technologies, and Applications. New York: McGraw-Hill.
construction of a dual system offers advantages in both water resource management and cost savings, as discussed in AWWA (1994) and Okun (2005). Substituting reclaimed water for potable water is one of the primary purposes of dual distribution systems. The use of reclaimed water for nonpotable purposes serves to conserve the potable water supply for use where drinking water quality is needed. In the planning of a dual distribution system, if the reclaimed water is used for fire fighting in lieu of potable water, the potable water pipelines and storage can be sized for delivery of domestic flows and not fire flows. Potable
water-quality benefits accrue because pipeline and storage sizes are reduced which in turn reduces the residence time in the potable water system. Long residence times can result in the loss of disinfectant residual and may promote the regrowth of microorganisms, which can affect bacterial quality, and tastes and odors. The distribution system for reclaimed water can be designed to provide unrestricted, on-demand service, or the reclaimed water can be provided in restricted hours. Because the principal use of reclaimed water in urban areas is for landscape irrigation, which is applied generally during the
Wastewater Reclamation and Reuse System
nighttime hours to minimize human contact and evaporation loss, unrestricted service may result in a high peak flow demand. The peak demand may be several times higher than the daily average flow rate available for producing reclaimed water. Storage reservoirs are, therefore, needed to meet maximum hourly demands. Where reclaimed water is used for fire fighting, emergency storage can serve as a backup for the distribution system when pumping stations or pipelines are out of service for maintenance or repair.
4.03.6.3 Centralized Systems The use of centralized or regional wastewater collection and treatment facilities for the production of reclaimed water is practiced extensively in developed urban regions and other densely populated areas (see Figure 1(e)). For some reuse applications, such as indirect potable reuse through reservoir or groundwater augmentation, centralized facilities are well justified. However, when a centralized collection system is not available, or it is desirable to have independent treatment facilities, decentralized and satellite wastewater systems may be an option. Decentralized wastewater reclamation systems have been used widely for landscape irrigation in suburban areas, thereby reducing demand on potable supplies in addition to other benefits. In areas located adjacent to a centralized collection system, satellite facilities may also be used to meet some of the reclaimed water demand. While satellite facilities share some common characteristics with the decentralized systems, satellite systems are differentiated because they have a direct connection to a centralized wastewater collection system and therefore do not have to manage or store solids on site.
4.03.6.4 Decentralized Systems Decentralized wastewater management (DWM) systems are used most commonly in semi-urban, rural, and remote areas, where installation of a centralized sewer system is not feasible (see Figure 1(d)). However, in some areas decentralized systems are used instead of centralized sewers to limit and control the type of development in a given area. Decentralized treatment systems present a significant challenge for the design engineer due to the need for high-quality reliable performance in light of a number of constraints, including long periods of time between maintenance activities, lack of redundant systems, high variability in flow rate and constituent concentrations, and site-specific factors. Decentralized systems are an integral component of smartgrowth community design initiatives in unsewered areas and an element of sustainable development because of the potential for low-impact wastewater management and other advantages presented below. Further, due to practical and economic limitations, it is recognized that it is not possible or desirable to install centralized sewers to service all areas in the United States. Therefore, DWM systems are necessary for the protection of public health and environment and for the development of long-term strategies for the management of water resources. DWM is defined as the collection, treatment, and reuse of wastewater at or near the point of waste generation (Crites and Tchobanoglous, 1998). Decentralized facilities may be used
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for wastewater management from individual homes, clusters of homes, subdivisions, and isolated commercial, industrial, and agricultural facilities. The wastewater flow rate, quality, and flow distribution will depend on the types of activities taking place as well as the scale of the application.
4.03.6.5 Satellite Systems In most collection and treatment systems, wastewater is transported through the collection system to a centralized treatment plant located at the downstream end of the collection system near the point of disposal. Oftentimes, opportunities for instituting water reuse applications, especially for agricultural and landscape irrigation or groundwater recharge, are limited as the points of use are located remotely from the wastewater-treatment facilities. The infrastructure costs for storing and transporting reclaimed water to the points of use are often prohibitive, thus making reuse uneconomic. An alternative to the conventional approach of transporting reclaimed water from a central treatment plant is the concept of satellite treatment at upstream locations with localized reuse. Residuals generated by satellite treatment process are discharged to the collection system for processing downstream at the central treatment plant. Satellite treatment systems generally fall into three categories: (1) interception type, (2) extraction type, and (3) upstream type. Each of these types of satellite systems is described further below. The distinction between satellite types is made because the characteristics of the wastewater to be treated, the treatment technologies that will be used, and the infrastructure needed to implement them are somewhat different, and, in some cases, quite different. Interception type. In the interception type, as illustrated in Figure 1(a), the wastewater to be reclaimed is intercepted before it reaches the collection system. Typical applications for this type of satellite system are for reuse in high-rise commercial and residential buildings. The quantity of flow to be intercepted and reclaimed will depend on the local and seasonal water reuse requirements. Typically, all of the flow from an individual building will be intercepted for reuse. In some cases, it may be necessary to supplement the intercepted flow with potable water. Should excess flow occur, it would be discharged to the collection system. Extraction type. In the extraction type, as illustrated in Figure 1(b), the wastewater to be reclaimed is extracted (mined) from a collection system main, trunk, or interceptor sewer. Typical applications for this type of satellite system are for reuse in landscape, park, and greenbelt irrigation; for reuse in nearby high-rise commercial and residential buildings; and for commercial and industrial cooling tower applications. The quantity of flow to be extracted and reclaimed will depend on the local and seasonal water reuse requirements, especially so for landscape irrigation applications. Upstream type. In upstream type, as illustrated in Figure 1(c), the wastewater reclamation facilities are used to reclaim water from developments located at the extremities of a centralized collection system and where opportunities for water reuse (e.g., golf course and median strip irrigation) are available and the capacity of the collection system is limited. Typical applications for this type of satellite system are for new
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Wastewater Reclamation and Reuse System
housing developments and remote commercial centers and research parks. The quantity of flow to be intercepted and reclaimed in upstream satellite systems will depend on the local and seasonal water reuse requirements. In general, all of the flow from a housing development will be intercepted for reuse. In some cases, however, it may be necessary to divert some of the flow directly to the centralized collection system, before or after treatment.
wastewater-treatment processes include strong disinfectants, fabric softeners, chemical sanitizers for holding tanks, chemotherapy medications, high amounts of oils or grease, and brine from water softeners. In larger systems a degree of anonymity exists that makes it difficult to identify the particular source of an offending discharge and there is less individual responsibility for performance and operational matters, as these activities become the responsibility of the municipality.
4.03.6.6 Point-of-Use Treatment For multiple water reuse applications, the economic question that must be addressed is whether it is more cost effective to (1) produce multiple grades of reclaimed water to meet the quality criteria for all users, (2) produce reclaimed water of a single quality that meets all criteria, or (3) produce a single grade of quality that meet most criteria and provide treatment at or near the point of use in special circumstances. Typically in water reuse system that involves multiple uses and a single quality of product water, reclaimed water-quality requirements are determined by a major user that requires the highest quality. For example, if a reclaimed water distribution system is to provide water for landscape irrigation, high-rise building toilet and urinal flushing, and industrial cooling towers, the microbial requirements for toilet and urinal flushing will be critical, whereas industrial cooling tower usage may require nitrogen and phosphorus removal to control biological growth, scaling, and corrosion. Thus, polishing treatment as required for a given application can be applied at the point of use may prove more economical depending on site-specific circumstances.
4.03.7 Source Control The spectrum of household products used on a daily basis will increase the overall salinity of the resulting wastewater. Other chemicals or compounds discharged with wastewater may be toxic to treatment organisms or plants irrigated with the treated effluent. Because the removal of salts and toxic constituents is beyond the scope of most small wastewater-treatment applications, source control or dilution may be required for some irrigation-type reuse applications.
4.03.7.1 Salinity and Toxic Constituents Ions commonly added to wastewater from domestic water use that contribute to salinity include the cationic species (such as sodium, calcium, magnesium, and potassium), and anionic species (such as bicarbonate, carbonate, chloride, fluoride, and sulfate). A potential advantage of decentralized treatment systems is that individuals who use the system have direct control of the problematic constituents entering the wastewater stream. While the concentration of salts in the water is typically low enough not to be of concern for most applications, if the discharge of brine from regenerating water softeners or the use of toxic chemicals is not compatible with a particular process or reuse application, these issues can be discussed with the system users, who also have an interest in proper operation of the system. Examples of substances which have been implicated in negative impacts to
4.03.7.2 Source Separation Source separating systems include facilities that are used to separate solid and liquid wastes without commingling with the bulk wastewater stream. Human waste can be segregated, with or without the use of water, with composting systems and waste incinerating systems. Collection and processing of human waste (and food waste in the case of in-sink food waste grinders) with a composting toilet or separate wet-composting system can reduce the size of downstream wastewater management systems and produce a compost material that can be used for landscaping purposes (Del Porto and Steinfeld, 1999). Source separation can also be used for liquid wastes, including urine diversion and graywater separation. Because of the high nutrient value of urine, toilets have been developed that divert urine to a separate holding tank for reuse in agriculture (Ecosan, 2003). Similarly, graywater is often considered for reuse due to the reduced presence of pathogens and organic matter. The level of maintenance and user participation required for source separating systems should be considered carefully when selecting these systems, as many of these processes have failed to work adequately in the field. However, in some areas, where limiting conditions exist, source separating systems may be a preferred alternative.
4.03.7.3 Graywater The water from bathing, hand washing, and clothes washing (not including soiled diapers), collectively known as graywater, is sometimes managed separately from human waste because it is relatively free of pathogens, organic matter, and trace constituents. When graywater is separated, wastewater from kitchen sinks, automatic dishwashers, and food waste grinders is discharged typically with toilet flushing water, collectively known as blackwater (note that drainage from kitchen sinks is included in household graywater in Australia). Separated graywater may be treated and reused more easily than combined graywater and blackwater. Some system designs incorporate direct drainage of graywater to mulch basins for tree irrigation, therefore not requiring treatment or storage and greatly reducing the system cost and maintenance needs (Ludwig, 2000). Separated blackwater may be treated separately or discharged to a collection system. Graywater systems are usually expensive to retrofit into a building, and therefore should be included, if possible, during building planning and construction. In some areas, the use of graywater for irrigation and toilet flushing is recommended during periods of water shortages. Management of graywater systems may present challenges if there is insufficient planning.
Wastewater Reclamation and Reuse System
4.03.8 Future Directions for Water Reuse In many parts of the world, agricultural irrigation using reclaimed water has been practiced for many centuries. Landscape irrigation such as irrigation of golf courses, parks, and playgrounds has been successfully implemented in many urban areas for over 30 years. However, salt management in irrigated croplands and landscapes may require special attention in many arid and semi-arid regions. Beyond irrigation and nonpotable urban reuse, indirect or direct potable reuse needs careful evaluation and closes public scrutiny. It is obvious from public health and acceptance standpoints that nonpotable water reuse options must be exhaustively explored prior to any notion of indirect or direct potable reuse, although modern technology is beginning to obviate these criteria. Reservoir augmentation as well as groundwater recharge with reclaimed water and direct potable water reuse shares many of the public health concerns encountered in drinking water withdrawn from polluted rivers and reservoirs. Three classes of constituents are of special concern where reclaimed water is used in such applications: (1) enteric viruses and other emerging pathogens; (2) organic constituents including industrial and pharmaceutical chemicals, residual home cleaning and personal care products and other persistent pollutants; and (3) salts and heavy metals. The ramifications of many of these constituents in trace quantities are not well understood with respect to long-term health effects. For example, there are concerns about exposure to chemicals that may function as endocrine disruptors; also, the potential for development of antibiotic resistance is of concern. As a result, regulatory agencies are proceeding with extreme caution in permitting water reuse applications that affect potable water supplies. In each case in the United States where potable water reuse has been contemplated, alternative sources of water have been developed in the ensuing years and the need to adopt direct potable water reuse has been avoided. As the proportional quantities of treated wastewater discharged into the receiving water increases, much of the research which addresses indirect potable reuse via reservoir augmentation and groundwater recharge and direct potable water reuse is becoming of equal relevance to unplanned indirect potable reuse such as municipal water intakes located downstream from
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wastewater discharges or from increasingly polluted rivers and reservoirs. Examples include New Orleans, Louisiana on the Mississippi River, and the Rhine Valley communities along the Rhine River in Germany and The Netherlands. Reclaimed water is a locally controllable water resource that exists right at the doorstep of the urban environment, where water is needed the most and priced the highest. Closing the loop of the water cycle not only is technically feasible in industries and municipalities but also makes economic sense. While direct potable reuse of reclaimed water is more or less a possibility, reservoir augmentation and groundwater recharge with advanced wastewater-treatment technologies are a viable option backed by the decades of experience in Arizona, California, Florida, New York, and Texas as well as in Australia, Israel, Germany, The Netherlands, and the United Kingdom. Water reuse has become an essential element of future water resources development in integrated water resources management in many parts of the world.
References Asano T (ed.) (1998) Wastewater Reclamation and Reuse, Water Quality Management Library, vol. 10. Boca Raton, FL: CRC. Asano T, Burton FL, Leverenz H, Tsuchihashi R, and Tchobanoglous G (2007) Water Reuse: Issues, Technologies, and Applications. New York: McGraw-Hill. Asano T and Levine AD (1995) Wastewater reuse: A valuable link in water resources management. Water Quality International 4: 20--24. AWWA (1994) Dual Distribution Systems, AWWA Manual M24, 2nd edn. Denver, CO: American Water Works Association. Del Porto D and Steinfeld C (1999) The Composting Toilet System Book: A Practical Guide To Choosing, Planning, and Maintaining Composting Toilet Systems, an Alternative to Sewer and Septic Systems. Concord, MA: The Center for Ecological Pollution Prevention. Ecosan (2003) Ecosan – closing the loop. In: Proceedings of the 2nd International Symposium on Ecological Sanitation. Lubeck, Germany: GTZ. Ludwig A (2000) Create an Oasis with Greywater, 4th edn. Santa Barbara, CA: Oasis Design. Okun DA (2005) Dual systems to conserve water while improving drinking water quality. In: 20th Annual WateReuse Symposium. Denver, CO. State of California (2000) Water recycling criteria. In: Title 22 Code of Regulations, Division 4, Sections 60301 et Seq., 2 December 2000. World Health Organization (2006) WHO Guidelines for the Safe Use of Wastewater, Excreta and Greywater. Volume II: Wastewater Use in Agriculture. Geneva, Switzerland: WHO.
4.04 Seawater Use and Desalination Technology S Gray, Victoria University, Melbourne, VIC, Australia R Semiat, Grand Water Research Institute, Technion, Israel M Duke, Victoria University, Melbourne, VIC, Australia A Rahardianto and Y Cohen, University of California, Los Angeles, CA, USA & 2011 Elsevier B.V. All rights reserved.
4.04.1 4.04.2 4.04.2.1 4.04.2.2 4.04.2.2.1 4.04.2.2.2 4.04.2.2.3 4.04.2.2.4 4.04.2.3 4.04.2.3.1 4.04.2.3.2 4.04.2.4 4.04.2.5 4.04.2.5.1 4.04.2.6 4.04.2.7 4.04.3 4.04.3.1 4.04.3.2 4.04.3.3 4.04.3.3.1 4.04.3.3.2 4.04.3.4 4.04.3.4.1 4.04.3.4.2 4.04.3.4.3 4.04.3.4.4 4.04.3.4.5 4.04.3.4.6 4.04.3.4.7 4.04.3.5 4.04.4 4.04.4.1 4.04.4.2 4.04.4.3 4.04.4.4 4.04.4.5 4.04.5 4.04.5.1 4.04.5.1.1 4.04.5.1.2 4.04.5.1.3 4.04.5.2 4.04.5.2.1 4.04.5.2.2 4.04.5.3 References
Introduction Seawater Water Quality Evaporative Techniques Pretreatment Multi-stage flash Multi-effect distillation Vapor compression Membrane Processes Pretreatment Reverse osmosis Desalination Process Costs Quality of Water Produced Increase in water hardness/water stabilization Environmental Aspects Energy Issues Brackish Water Brackish Water Desalination Applications Brackish Water Desalination Technologies Common Process Configuration RO/NF process configuration ED/EDR process configuration Major Challenges Concentration polarization and membrane mineral scaling Mitigation of membrane mineral scaling Managing the impact of feedwater-quality variation Enhancing water recovery Concentrate disposal Specific contaminant removal Cost of brackish-water desalination Future Developments Desalination of Wastewater for Reuse Water Quality Pretreatment RO Processes Final Water Quality Concentrate Disposal Alternative Technologies Membrane Distillation Brief history Membrane distillation configuration The Memstill project Forward Osmosis Background Recent developments Capacitive Deionization
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4.04.1 Introduction Water, water, every where,
And all the boards did shrink;
Water, water, every where,
Nor any drop to drink
The above lines from Samuel Taylor Coleridge’s The Rime of the Ancient Mariner are often quoted to highlight the abundance of seawater but our inability to use it because of its high salt content. People require water of low salinity or freshwater for consumption, and typical values of less than 500 mg l1 total dissolved solids (TDS) are often used by health regulators to specify salinity requirements for human consumption (National Water Quality Management Strategy, 2004). It is, therefore, not surprising that many believe that distillation processes have been used to produce freshwater since the fourth-century BC, although the first documented case appears to be from the early seventeenth century when Japanese sailors boiled seawater in pots and collected the condensate in bamboo tubes. The world’s first industrial desalination plant is considered to have been commissioned in 1881 on the island of Malta, while Saudi Arabia had its first desalination plant installed by the Ottoman Turks in Jeddah, 1907. The application of desalination technology grew throughout the twentieth century, with many applications in the Middle East and on water-scarce islands. However, at the end of the twentieth and start of the twenty-first century, there has been a rapid increase in the desalination capacity, and Figure 1 shows the rapid growth in installed global desalination capacity between 1980 and 2009 (Global Water Intelligence, 2009). This rapid growth has been driven by population growth and
changing climatic conditions that have lead to lower rainfall or altered rainfall patterns, and has resulted in many communities becoming water stressed. Figure 2 is an estimate of regions that will be experiencing various degrees of water scarcity by 2025, and the World Health Organization (WHO; World Health Organisation, 2010) estimates that one in three people in the world are affected by water scarcity (Seckler, 1998). In an effort to combat water-scarcity issues, communities are treating poorer-quality water sources, and desalination of seawater, brackish groundwater, and salty wastewater is increasingly practiced. Coupled with the need for increased use of desalination technology there has been a dramatic decrease in the cost of desalination, with approximate costs of 20–35 cm3 for brackish water, 30–40 cm3 for wastewater, and 50–100 cm3 for seawater. Figure 3 shows the dramatic decrease in unit costs for seawater desalination over the period 1990–2003. The increased affordability and need for desalination have resulted in many communities becoming increasingly dependent upon desalination technologies. Initially, largescale desalination was predominantly confined to areas with severe water limitations such as the Middle East or island communities; however, desalination is increasingly practiced by communities that have traditionally relied upon surface water sources, such as London, Singapore, Chennai, and Sydney. The removal of salt is thermodynamically more difficult than the removal of solid particles or large-molecular-weight molecules, as the osmotic pressure of the salt solution must be overcome (see Chapter 4.11 Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis). Therefore, the energy required for desalination is generally greater than that for other treatment processes, although the energy required is a strong function of salt concentration. The thermodynamic minimum amount of energy required for desalination of seawater is 0.79 kW h1 m3 if water is taken from an infinite salt solution (Semiat, 2008).
New desalination capacity 1980−2009 8 Commissioned 7 Capacity (million m3 d−1)
Contracted 6 5 4 3 2
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Figure 1 Global desalination capacity. From Global Water Intelligence (2009) New desalination capacity 1980–2009-chart. http:// www.globalwaterintel.com/archive/10/10/analysis/new-desalination-capacity-1980-2009-chart.html (accessed April 2010).
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Projected water scarcity in 2025
Physcial water scarcity
Little or no water scarcity
Economic water scarcity
Not estimated
Figure 2 Global Watering. From Seckler D, Amarasinghe U, Molden D, de Silva R, and Barker R (1998) World water demand and supply, 1990 to 2025: Scenarios and Issues, International Water Management Institute, Research Report 19, http://iwmi.cgiar.org/Publications/ IWMI_Research_Reports/PDF/PUB019/REPORT19.PDF
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Figure 3 Cost of seawater desalination (US$ m3) from new plants against year. Data from Adham S. (1997) Desalination: Applications, cost trends, and future issues of RO technology. In: AWA Membranes Specialty Conference II, 21–23 February. Melbourne, VIC: Australian Water Association.
If the water recovery is increased to 50%, then the minimum energy requirement is 1.09 kW h1 m3, because water is extracted from incrementally higher salt concentrations as the water recovery increases. These thermodynamic values are the absolute minimum amount of energy required, and actual desalination must use more than this. Typically, commercial seawater-desalination plants use between 4 and 10 kW h1 m3 depending upon the type of plant installed. These energy requirements compare to values of o1 kW h1 m3 for alternative water supplies such as dams, storm
water, and recycled water (Leslie and Myraed, 2009). The greater use of desalinated water has coincided with climatechange concerns arising from greenhouse-gas emissions, and subsequently communities have sought to limit the greenhouse-gas emissions from seawater desalination by the use of energy from renewable sources (Crisp, 2009; Voutchkov, 2009). Although the energy used for operating an entire water and wastewater-treatment supply might only be 15% of the energy used for heating water in some Western communities (Kenway et al., 2008), the focus on energy efficiency that
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pervades modern communities dealing with climate change implies that energy use will remain an issue for seawater desalination and be a key focus for new desalination technologies that are being developed. This chapter outlines the common desalination processes used for treatment of seawater, brackish water, and wastewater, and comments on operating issues and performance of these processes. It focuses on reverse osmosis (RO) membrane systems for treatment of the three types of waters considered and thermal desalination systems for seawater desalination, as these two processes are the dominant commercial processes. Discussion of the desalination of wastewater focuses on fouling chemistry, as another chapter (see Chapter 4.03 Wastewater Reclamation and Reuse System) explains wastewater reclamation and reuse systems. Additionally, a brief outline of alternative desalination processes is provided as many such processes are being developed.
4.04.2 Seawater 4.04.2.1 Water Quality About 97% of the water on Earth is found in the seas and oceans that cover approximately 70% of the Earth’s surface. Salt content and concentration vary slightly from place to place. Open oceans contain approximately 3.5% weight salt, while smaller closed seas may contain higher concentrations. For instance, the Mediterranean Sea contains close to 4% salt, while the Red Sea and the Persian Gulf contain 4.2% dissolved salts. Lower concentrations can be found in other closed seas, such as the Baltic Sea where salt concentration changes during the year from 0.5% to 1.5%, while the salinity of the Black Sea salinity is below 2%. The typical salt composition of seawater is shown in Table 1 (Water Chemistry, 2010), while Figure 4 shows a global distribution of ocean salt concentrations. Sodium and chloride represent the most abundant cations and anions in the sea, and the concentration of magnesium is significantly greater than that of calcium. Besides the spatial variation in salt concentrations across the globe, the amount of CaCO3 also varies with depth. CaCO3 is saturated in the surface layer of seawater and below saturation concentration at lower levels (Le Gouellec et al., 2006). This is of significance for RO membrane systems, since Table 1
Seawater salt composition
Component
Concentration (%)
Calcium Magnesium Sodium Potassium Bicarbonate Sulfate Chloride Bromide Total dissolved solids
0.042 0.13 1.07 0.04 0.015 0.27 1.94 0.007 3.5
From Water Chemistry (2010) Nitto-Denko – Hydranautics. http://www.membranes. com/docs/papers/04_ro_water_chemistry.pdf (accessed April 2010).
the concentration of salts increases along the RO membrane during the process and precipitation of salts fouls the membrane. The bromide and boron concentrations in seawater are low, but their concentrations are significant as they can adversely affect the treated water quality, particularly for RO treatment systems. While bromide rejection is high through RO membranes (90–95%), the permeate still contains sufficient bromide to cause bromate issues should ozonation be used as a means of disinfection. Where bromide contributes to taste and odor issues in the system, such as in Perth, Australia, low bromide concentration in the final treated water has been specified (0.1 mg l1; Crisp, 2009). Boron is present as boric acid, in a concentration of approximately 5 ppm. Boric acid is a small molecule that can penetrate through RO membranes so the product may contain around 1 ppm of boron. This is important due to the sensitivity of many crops to boron content in irrigation. New RO membranes can reject up to 90% of the boron compared to 30–70% rejection for standard RO membranes. High boronrejecting membranes have the potential to minimize problems due to boron, although currently there is little long-term operational experience with these membranes. Seawater also contains organic contaminants that come from living marine creatures. These contaminants include all types of small to large molecules, colloids and viruses, bacteria, algae, and larger living or nonliving suspended matter. These contaminants also need to be removed along with salt to attain drinking-water standards (Morse et al., 1979) and to prevent fouling of membrane and thermal desalination systems. It is important that particulate foulants are removed prior to treating the water in RO systems as particles around 1 mm in size can block spacers in RO modules, while small particles o1 mm in size can lead to particle fouling of RO membranes. Particle fouling is less of an issue for thermal desalination processes. Algal blooms or red tides can cause temporary but significant increases in turbidity and total organic carbon (TOC), leading to rapid fouling of membranes. The high TOC concentrations not only lead to greater organic fouling of membranes, but also provide food for microorganisms to grow on the membranes. Pseudomonas, Bacillus, Arthrobacter, and Corynebacterium are the bacteria usually associated with biofouling in seawater RO systems (Voutchkov, 2008). In areas prone to red tides or oil spills, sensors are often used to detect their presence and the feed stopped if algae or oil is present. The concentration of organic molecules typically varies from o0.2 mg l1 in ocean waters not affected by freshwater sources (rivers, stormwater, and wastewater) or algal blooms to as high as 8 mg l1 or more for waters impacted by freshwater sources (Voutchkov, 2008). Additionally, the composition of the organic compounds can vary, with the lowmolecular-weight compounds being typically between 20% and 50% of the TOC, low-molecular-weight acids and neutrals between 18% and 25%, humic substance between 26% and 52%, and polysaccharides between 1% and 14%. Such variations in composition can have a significant effect on biofouling, as the presence of easily biodegradable compounds, such as polysaccharides, can increase the level of biological activity in the desalination process, thereby increasing the
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Figure 4 Global surface seawater salinity levels. From Antonov JI, Locarnini RA, Boyer TP, Mishonov AV, and Garcia HE (2006) In: Levitus S (ed.) World Ocean Atlas 2005, Volume 2: Salinity. NOAA Atlas NESDIS 62, 182pp. (CD-ROM). Washington, DC: US Government Printing Office. Permission from NOAA’s National Oceanographic Data Center. http://serc.carleton.edu/eslabs/corals/4c.html (accessed May 2010).
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degree of biofouling. It has been observed in a number of plants that temporal increases in TOC over a 1–2- week period is linked to increases in biofouling (Voutchkov, 2008).
4.04.2.2 Evaporative Techniques 4.04.2.2.1 Pretreatment Thermal processes are not very sensitive to the initial concentration of seawater and are also less sensitive to suspended particles than membrane-based systems. A simple straining filtration technique to remove coarse particles is usually suitable. De-aeration is needed to remove oxygen and to reduce the possibility of noncondensing gases accumulating, as these can cause corrosion within the thermal desalination process. Simple ejector-condensers are often used for this purpose. Thermal processes are more sensitive to possible precipitation of calcium salts than membrane processes, mainly gypsum on heat-transfer surfaces.
4.04.2.2.2 Multi-stage flash Multi-stage flash (MSF) distillation is still considered as the most common and simple technique in use. It has been operated commercially for more than 40 years (Awerbuch, 1997b). The technique is based on condensing low-pressure steam to produce heat for evaporation of seawater. A schematic presentation of an MSF desalination plant is shown in Figure 5. The process is based on slightly pressurizing the seawater feed and passing it through long closed pipes and a series of flash chambers. Condensing vapor, generated in the flash chambers is used to heat the feedwater in the pipes. Energy is added to the system to heat the feedwater to the initial high temperature of approximately 120 1C, and lowpressure steam is commonly used as the heating source. The low-pressure steam is usually extracted from a power station. The heated seawater feed is introduced into a series of flash chambers where it is allowed to flash along the bottom of the chambers. The pressure is reduced along the chambers so that water continues to flash in each chamber. This generates vapor
Steam
that passes through mist eliminators before condensing on the seawater feed pipes. The condensate is collected from the pipes and pumped out as the plant product. Part of the concentrated brine is recycled and mixed with the feed to increase the recovery ratio, and the rest is pumped out and rejected into the sea. Energy is transferred to heat the seawater feed as the vapor condenses on the condenser pipes containing the seawater feed. The sensible heat of condensation is recovered to produce vapor. Based on an enthalpy balance, the water-recovery ratio is low and recirculation of the brine is required in order to increase water recovery. The energy consumption in this technique is high, and is associated with heating of the feedwater, low sensible-heat recovery, and pumping of feed and brine recirculation. The energy efficiency, the size of the plant, and the cost involved are affected by design parameters such as (1) the number of stages from the high-temperature feedentrance point to the brine exit, (2) the recirculation ratio, (3) the temperature of the preheated feed seawater, (4) heattransfer quality of the condensing vapor, (5) improved utilization of the heat rejected with the product and the rejected brine, and (6) controlling and preventing scale formation and prevention of accumulated noncondensable gases. Working on extracted steam at the end of a power station, at a temperature of about 120 1C, the typical energy consumption is estimated to be 7–9 kW h1 m3, depending on the gain output ratio (GOR) (Semiat, 2008). Corrosion is of concern in MSF systems, as water of very high purity is corrosive by nature. Corrosion is related to the operational temperature, the existence of dissolved oxygen in the water, and the choice of materials for the heat-transfer surfaces used in the heated seawater environment. Stainless steels are used for the pipe work and epoxy-coated wrought iron for the vessels. De-gassing is achieved in the pretreatment by the use of strippers or vacuum systems and the addition of hydrazine is also practiced. Typical temperatures vary between 110 and 120 1C at the hot end, down to seawater temperature at the cold end.
Vent ejector
Brine heater
Condensate
Seawater
Brine
Product water
Figure 5 Schematic presentation of a multi-stage flash (MSF) desalination plant, MSF Sidem Design. Adapted from http://www.acwasasakura.com/ images/msf.gif (accessed 24 April 2010).
Seawater Use and Desalination Technology
An important advantage of this process, when compared with other distillation processes, is that scale does not precipitate on heat-transfer surfaces, but in the flash chambers. This enables the heat-transfer surfaces to remain clean, and cleaning of the system is rarely required. The quality of the product from this technique, as in other evaporation techniques, is extremely good. Product water usually contains less than 50 ppm of TDS. TDS is carried into the product via small drops that pass through the mist eliminators, rather than via the vapor phase. Better quality of water, down to 10 ppm TDS can be produced for industrial purposes if greater efforts are made to reduce carryover of mist. Due to lack of salts, the product water is aggressive and can cause corrosion. It is usually passed through a bed of lime to increase the calcium-carbonate concentration, or it is mixed with another source of water to stabilize the water and prevent corrosion. MSF processes are usually associated with large-scale cogeneration plants, where waste heat from power plants is available. The simple design and ability to be unaffected by scale make MSF a robust process that is easy to operate. However, the trade-off is higher energy consumption compared to alternative thermal and membrane-based systems and relatively high production costs. Information regarding the cost of this process may be found in Jassim and Ismail (2004).
4.04.2.2.3 Multi-effect distillation Multi-effect distillation (MED) is considered a more sophisticated and more energy-efficient evaporation technique than MFS systems (Awerbuch, 1997b). Multi-stage evaporation has been used for many years for the purpose of solution concentration, crystallization, solution purification, etc., and has also been used for seawater desalination, for the last 45 years. The method is based on a low-temperature source of energy. The main source of energy is spent steam emerging at the exit of a steam-operated power station, but alternatives also include low-level steam or hot fluid from other sources. Figure 6 describes the schematics of a horizontal tube MED unit. The steam enters the plant and is used to evaporate
79
heated seawater. The secondary vapor produced is used to generate tertiary steam at a lower pressure. This operation continues along the plant from stage to stage. The primary steam condensate is returned to the boiler of the power station. The technique is based on double-film heat transfer, where latent heat is transferred in each stage from condensing of steam through the heat-transfer piping to the evaporated falling film of seawater. The process is repeated up to 16 times in existing plants, between the upper possible temperature and the lower possible cooling water, which depends on seawater temperature. Condensate accumulates from stage to stage as product water. A vacuum pump removes the accumulated noncondensable gases, together with the remaining water vapor, after the last condensation stage in order to maintain the gradual pressure gradient inside the vessel. The pressure gradient is dictated by the saturation pressure of the feedstream and the saturation pressure of the condensing steam leaving the last stage and condensed by cooling with seawater. Typical pressure gradients of 5–50 kPa across the system (o5 kPa per stage) are typical. Steam condenses in common MED installations inside horizontal pipes where seawater evaporates on the other side as it falls down the tube bundle. Heat transfer in double-film condenser–evaporators is a very efficient mechanism that controls the process and can operate with a low-temperature driving force across the tube walls. The heat transfer is bounded by the increasing boiling-point elevation along the plant as the salt concentration increases with removal of water on one hand, and prevention of surface boiling by keeping the on-temperature difference so that boiling does not start, on the other. The performance ratio, or the GOR, which refers to the number of tons of water produced per ton of initial steam, is considered high. The ratio in MED can be up to 15 compared to a maximum of 10 for MSF. Therefore, the energy or thermal efficiency is essentially higher for MED than it is for MSF (Ophir and Weinberg, 1997). Energy utilization is increased with the number of stages. If low-cost heat at lower temperature is available, optimization of operation conditions may lead to a lower number of stages. Additionally, lower-temperature operation allows the use of low-cost heat-transfer surfaces without the problems of severe
MED and TVC process schematic Heat-recovery evaporator
Heat-rejection condenser Heat-pressure steam
Recycle vapor
Heat-pressure steam
NCG Seawater
High-pressure steam (for an TVC plant) Low pressure steam (for an MED plant)
Freshwater Brine
Condensate tank Condensate return pump
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Polyphosophate Feed pump
Figure 6 Schematics of a horizontal tubes multi-effect distillation (MED) plant, IDE Design. From http://www.ide-tech.com/ (accessed May 2010).
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corrosion and reduces the likelihood of CaSO4 precipitation on the tubes, and hence improves plant reliability. Low temperature differences and good wetting of the surfaces prevent scaling. Therefore, if a plant is operated below 70 1C, it is possible to use aluminum pipes, reduce corrosion, and operate at sub-saturation conditions for gypsum up to 60% recovery. Additionally, there is no need to remove oxygen below 70 1C, as corrosion rates are very low and cleaning is less frequent. The capacity of an MED plant is usually less than 30 000 m3 d1, but the modular design enables several trains to be built adjacent to each other to enable larger overall plant capacity. MED system designs can vary, with vertical or horizontal tubes, or flat-sheet heat exchangers, arrangement of the stages horizontally or vertically, and co-current or countercurrent flow of seawater against the direction of produced steam. Such variations in design affect the ease of cleaning heat exchangers, the pumping of water flows, and energy losses in the system, and sometimes, specific process designs are developed based on the site conditions. Designs also differ with regard to scaling potential, as the path of the circulating brine in connection to calcium sulfate hydrate saturation may vary. Co-current operation is advantageous in this respect, since the calcium sulfate hydrate saturation level increases when the water temperature reduces. In co-current operation, the highest calcium sulfate hydrate concentrations occur at the lowest temperatures, where higher saturation levels must be reached before precipitation occurs. In countercurrent operation, however, the opposite is true and the highest calcium sulfate hydrate concentrations occur at the highest temperature where the saturation levels before precipitation occur are lower. This is an important design consideration for scale control and for water circulation and pumping expenses. Good water distribution, evenly distributed on the heat-transfer tubes is essential to reduce scaling. Co-current operation takes place in the MED-Metropolitan Water District (MWD) tower design, where the highest temperature is obtained at the lowest calcium sulfate concentration. In this design, the brine temperature–concentration curve along the tower is closed, almost parallel to the CaSO4 saturation–temperature curve. More information on the cost of MED may be found in Ophir and Lokiec (2004).
4.04.2.2.4 Vapor compression The vapor compression (VC) technique is similar in operation to the MED process, as condensation of vapor from each stage is used to generate vapor from brine in the next stage. Heat transfer usually takes place, as in MED, in the form of a double-falling film, which is an effective heat-transfer mechanism. Seawater is preheated against the brine discharge and the product water leaving the system. This is a heat-pump process, where the latent heat of the condensing vapor is used to make more vapor on the other side of the heat-transfer surface. However, vapor generated during the last stage of the production process is compressed to higher pressure, followed by an increase in temperature, after which it is recycled to the first stage of the process where it condenses and the heat of condensation is used to evaporate feedwater. In contrast,
vapor from the last stage of an MED or MSF process is condensed and mixed with the product water. Therefore, VC allows the sensible heat associated with the last stage to be used in the evaporation process and hence is more energy efficient than MED and MSF. The main need for energy is, therefore, for elevating the pressure of vapor from the last stage to provide the driving force for heat transfer in the first stage. The process usually comprises one to six stages. The operating temperature may be chosen for the best optimization of the process, as the upper temperature is determined by the number of stages and the lower temperature by the flow rates and the properties of the vapor. Compressing of lowtemperature gases is expensive due to their density and/or specific volume. A part of the brine recirculates to increase water recovery. A common approach to recycling of the vapor is the use of a mechanical compressor that operates at relatively low pressure and high specific vapor volume, but thermal compression is also practiced by mixing with higherpressure steam. Figure 7 presents a schematic view of a mechanical VC unit. Mechanical VC benefits from the fact that it requires only an electrical source of energy from the grid or from a diesel generator, and energy consumption is between 7 and 8 kW h1 m3. A source of steam and a source of electricity for water evaporation and pumping are needed for the thermal compression process. The process can operate close to a power station where steam and electricity are readily available, or it may use hot gases from different sources to generate lowpressure steam, for example, from a gas turbine or a diesel generator. VC operates mainly at small scale, in small installations such as hotels or refineries. The maximum reported capacity is of the order of 5 000 m3 d1 using two mechanical compressors. Higher capacities, up to 10 000 m3 d1, may be achieved with thermal compression. The ability of VC to operate at low temperatures makes it possible to use simple metals such as aluminum, with almost no corrosion attack and prevention of scale formation. The use of electricity makes the technique compatible for use in parallel with other desalination techniques, as in hybrid operation for optimization of energy consumption. A modern compressor presents efficiency of up to 80%. The quality of the product is similar to that of other evaporation techniques. The technique may also be used for part removal of salts that are at saturation level, in cases of low-boiling-point elevation. More information on VC can be found in El Dessouky (2004).
4.04.2.3 Membrane Processes Desalination with RO membranes is based on applied dynamic pressure to overcome the osmotic pressure of the salt solution in feedwater. The term osmotic pressure represents a property of a solution, containing dissolved matter, salts in water, starch, or sugar, such as that found in the roots of most plants. The relatively high concentration of this solution allows transfer of water from the land surrounding the root through the membrane skin of the root. Applying a pressure to the concentrated solution on one side of the membrane will stop the flow of water. This pressure is defined as the osmotic pressure of the solution. Higher applied pressures on the
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MVC process schematic Evaporative condenser
Heat transfer tubes
Vapor compression system
Decoder seperator
NCC removal auxiliary condenser
Seawater supply pump Feed dosage pump Product storage
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Brine
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Figure 7 Schematic presentation of a horizontal tube, single-stage vapor compression (VC) desalination unit, IDE Design. From http://www.idetech.com/ (accessed May 2010).
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Pre-treatment Energy-recovery unit To concentrate disposal
Product water
Figure 8 Schematic presentation of reverse osmosis (RO) desalination plant. From Semiat R and Hasson D (2009) Sea-water and brackish water desalination with membrane operations. In: Drioli E and Giorno L (eds.) Membrane Operations. Innovative Separations and Transformations. Weinheim: Wiley-VCH Verlag GmbH & CO. KGaA. ISBN: 978-3-527-32038-7.
concentrated solution side, well above the osmotic pressure, will overcome the solution properties and transfer water from the concentrated solution through a membrane to produce freshwater. This is the basis of the RO process, which allows water-selective permeation through the membrane from the saline side to the freshwater side (Faller, 1999). Salts rejected by the membrane stay in the concentrate stream and are removed from the membrane by the flow of fresh salt solution along the membrane. Removal of permeate, the freshwater product, occurs via a permeate tube on the lower-pressure side of the membrane (see Chapter 4.11 Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis for more details).
Figure 8 depicts a schematic flow sheet of a typical RO desalination plant. Feed pretreatment for the removal of suspended material, bacteria, and organics is performed by either media filtration or now increasingly by UF or MF modules in modern plants (Semiat and Hasson, 2009). If residual chlorine is present, it is removed by active carbon filters or sodium metabisulfite. The feedwater is pumped into the RO module where water selectively passes through the membrane to produce freshwater. The high-pressure pump used to feed the RO-membrane module may be connected on a single shaft with the motor and a turbine, as is the case in the Eilat seawater plant, in order to recover the energy content of the pressurized concentrate. Energy-recovery devices such as
82
Seawater Use and Desalination Technology
independent turbines for secondary stages, pressure exchangers, and other techniques may also be used. The concentrate is disposed back into the ocean. Each of these unit processes are discussed in turn, along with the water-quality issues that affect process design.
4.04.2.3.1 Pretreatment Membrane-based desalination systems are reliant on thin, semipermeable membranes to separate water from brine, and these membranes are sensitive to contaminants in water. Therefore, extensive pretreatment is required to provide highquality water to RO membranes. This is a general requirement for high-pressure membranes and applies equally to seawater, brackish water, and wastewater feeds. Indeed, many performance issues can be traced back to poor pretreatment of the feedwater and subsequent fouling or scaling of the membrane surface. The standard industry test for the suitability of feedwater quality with respect to particle load is the silt density index (SDI) (ASTM, 2002). This test is based on measuring the time taken to filter the initial 500 ml of water through a 0.45 mm membrane at a constant pressure of 207 kPa, and the time taken to filter a second 500 ml of water after 15 min of filtration. The time taken for these amounts of water to pass reflects characteristics of the filter cake that has developed, and an equation is used to provide a single SDI number. Low values of SDI represent high-quality feedwater and high numbers poor-quality feed. Generally, SDI values o3 are considered suitable for RO feedwater, values between 4 and 5 represent adequate feedwater quality for slow flux decline, and values above 5 reflect poor quality. While SDI is used as the industry standard test, it is commonly known that it is not suitable for all waters, that it is sensitive to the test conditions and results may vary with slight changes in technique (Mosset et al., 2008). There have been several attempts to develop an improved testing method to determine the suitability of feedwater quality, such as the modified fouling index (MFI) (Boerlage et al., 2003), but currently these have not been adopted by the industry. The common approach to removal of particles is screening, to remove coarse particles and debris, followed by coagulation, sand filtration, membrane filtration, or dissolved air filtration (DAF). A screen with bar racks (75–100 mm) is used to remove large debris from seawater. This is followed by fine bar screens (3–10 mm) to remove finer material and protect downstream processing units. Further screen or grit removal chambers may also be used prior to filtration, particularly if microfiltration (MF) or ultrafiltration (UF) systems are installed. Coagulation of seawater prior to filtration uses ferric salts such as ferric sulfate or ferric chloride. Aluminum salts are not preferred, as it is difficult to maintain low dissolved aluminum levels and this can cause subsequent membrane-fouling problems. Similarly, overdosing of ferric salts can increase the dissolved iron concentration, which leads to membrane fouling downstream and regular jar tests are required to maintain the correct coagulant dose. Coagulation tanks (approximately 30 min coagulation time) or in-line static mixers may be used; however, static mixers are not recommended
when there are large flow variations due to their inability to provide adequate mixing under these circumstances. Dual-media pressure filtration with anthracite and sand is commonly used, with typical filtration rates between 15 and 20 m3 m2 h1. Gravity filters are also used when algae are likely to be present, but lower filtration rates of 10 m3 m2 h1 are typical. If the turbidity of the feedwater is regularly high (430 NTU), then sedimentation prior to sand filtration may also be used. The addition of coagulant allows sand filters to remove particles as small as 0.5 mm as well as some natural organic matter. Backwash of the sand bed is performed in order to remove the accumulated particles that are either sent to the sea or, where environmental regulations are stricter, to landfill. Small desalination plants sometimes use beach wells during the pretreatment stage. Seawater passage through the soil layers around the wells serves as a sand-filtration unit. This operation is limited due to the low possible capacity of a single well. DAF uses small air bubbles to capture contaminants and float them to the surface of the tank where they are removed by skimming. Oils, algae, and suspended matter attach to air bubbles and rise to the surface where they accumulate before being skimmed out for disposal. The water is often passed through a sand filter following DAF treatment to ensure that water with low suspended solids is fed to the desalination stage, and at the Tuas seawater desalination plant, the filtration units are incorporated into the DAF. DAF has the advantage of removing dissolved and suspended organic matter and oils that cannot be removed using sand filtration, and can treat waters with turbidity up to 50 NTU. This makes DAF suitable for handling waters prone to algal outbreak. MF and UF systems are now available for filtration of seawater and are increasingly being used, with seawater desalination plants in Yu-Huan (China), Fukuoka (Japan), Saudi Arabia, and Turkey using UF pretreatment. The advantage of using MF or UF pretreatment is that these systems can remove particles and colloids as small as 0.2 mm for MF and 0.02 mm for UF, and have more than 4 log removal of bacteria. This leads to consistently high treated water quality, with turbidity consistently less than 0.1 NTU and SDI less than 3. Iron coagulation is generally used to remove organic compounds and to enable fluxes between 50 and 100 l m2 h1 to be achieved. Both pressure and vacuum systems are used, with vacuum systems requiring less coagulant prior to the membrane and pressure systems being less sensitive to source-water temperatures. For cold waters less than 15 1C, pressure systems are more economically attractive. Finer screening is required before MF/UF systems than for dual-media filters, and screening of particles down to 120 mm or less in size is required. This is because the hollow-fiber MF/ UF systems may be cut or punctured by shells or sand particles. Furthermore, embryonic barnacles need to be removed to prevent colonization within the MF/UF systems and a 120-mm screen is able to achieve this. MF and UF systems can also be prone to fouling by organic compounds of biological origin and do not treat algae-laden waters efficiently. Chemically enhanced backwashing with chlorine solutions of 25– 100 mg l1 chlorine is used to control this fouling, and it may be used 1 or 2 times a day.
Seawater Use and Desalination Technology
Waste streams from media filtration systems are generally only half the volume of waste streams generated by MF/UF systems. Typically between 2% and 4% of feedwater is rejected in the waste stream in media-filtration systems, while between 5% and 8% of the feed might be rejected in MF/UF filtration. The higher waste associated with MF/UF systems is associated with their more frequent backwashing and the additional waste streams are associated with chemically enhanced backwashing and chemical cleaning of the membranes. Ion exchange is also used in some cases, such as in Palmachim, Israel. The ion-exchange resins are regenerated by the brine concentrate, and the calcium and magnesium released from the ion-exchange resin stabilize the product water. Nanofiltration (NF) membranes were suggested as a means to remove Ca/Mg and SO4 ions, but the cost for such an operation is high (Wang et al., 2009; Hassan et al., 1998; Hilal et al., 2007). Disinfection of the feedwater by chlorine compounds is done in some instances, in an effort to reduce biofouling of the RO membranes. The sensitivity of the RO membranes to chlorine implies that feedwater needs to be dechlorinated prior to treatment by RO membranes, and activated carbon is needed for this purpose. Unfortunately, in many cases, activated carbon allows bacteria to grow after the dechlorination stage, and, in some cases, it increases the impact of biofouling at the RO stage. Therefore, disinfection of the feedwater is no longer done in modern plants. The final stage of pretreatment is often cartridge filtration, particularly for systems that use media filters rather than MF/ UF systems. The cartridge filters are included to protect the RO membranes by removing contaminants that have made it through the pretreatment stage and to filter out particles associated with the breakthrough of media filters. The filters have nominal pore sizes between 1 and 25 mm and may need to be changed every 6 months in systems with good pretreatment or every 6 weeks if the source water is more challenging. Calcium carbonate is saturated on the surface layer of seawater and is below saturation concentration at lower levels. This is of significance for RO-membrane systems, since the concentration of salts increases along the RO membrane during the process, along with the level of supersaturation. This salt may, therefore, precipitate on the membrane and cause clogging and reduce the flux of freshwater. Acidification and addition of anti-scalants are often practiced, but it is now considered unnecessary because it is believed that the high concentration of brine assists in preventing precipitation. Ba and Sr salts are also in supersaturation in seawater; yet, their concentration is low and causes minimal damage to the membranes. CaSO4 has three different forms of salts, yet they are all below supersaturation levels in the brine leaving the desalination plant. More detailed information on pretreatment before membrane-seawater-desalination systems can be found in Voutchkov (2008) and Voutchkov and Semiat (2008).
4.04.2.3.2 Reverse osmosis The RO process relies on semipermeable membranes to selectively pass water from salt solutions. The building
83
materials for RO membranes are usually polymers such as cellulose acetates, polyamides, or polyimides. The membranes are semipermeable and made of thin layers about 200 nm thick that are adhered on to a thicker support layer. A few types exist, such as symmetric, asymmetric, and thin-film composite membranes. The membranes are usually built as long sheets, separated by spacers, and spirally wound around the product tube. In some cases, tubular or capillary membranes are used or even hollow fibers. Sidney Loeb developed the first modern RO membrane based on cellulose acetate (Loeb and Sourirajan, 1963; Loeb, 1981). Modern membranes are made of polyamides and polyimides that have better rejection properties, longer life, and require less energy. RO membranes are usually sensitive to changes in pH and the recommended pH range for polyamide membranes is between pH 2 and 11, but they can work for a short time (30 min) at extreme conditions of pH, such as 1 or 13, for cleaning purposes. Small concentrations of oxidizing substances such as chlorine, chlorine oxides, and ozone can severely damage the membrane skin. So also can a wide range of organic materials and biological organisms such as algae and bacteria, and this is one reason why high-quality pretreatment is required. The process takes place at ambient temperature, and RO membranes are usually stable only up to 35–40 1C. However, variations in water temperature within this range still influence membrane performance. The flux through a membrane increases with rising water temperature as the viscosity of water decreases. Higher temperatures increase both the water flux and salt flux, and lower rejections are obtained. Using hot seawater flowing from the cooling system of a large power plant, as is sometimes practiced, increases the efficiency as the membranes are able to operate at higher flux. A schematic presentation of an RO desalination process is shown in Figure 8. Seawater is pumped through the pretreatment stage before high-pressure pumps feed the water to the RO modules. Water then penetrates the membrane and freshwater is recovered in the permeate stream. The highpressure purged concentrate contains energy that may be recovered using turbines or pressure-exchange devices. The osmotic pressure of seawater, as an example, varies between 24 bars to twice as much for the concentrate at 50% recovery. Operating pressures therefore, vary between 60 and 80 bars for seawater desalination in order to allow sufficient permeation of water through the membrane at the relatively high concentrations of the brine along the pressure vessel. Water recoveries of 35–50% are usual, with lower water recovery obtained in closed seas, such as the Red Sea or the Persian Gulf, due to higher salt concentration. The requirement of a two-pass process during RO and the occurrence of membrane fouling are discussed in the following. The need for a two-pass process. RO membranes are able to selectively pass water, but the separation is not 100% and small amounts of salt and organic compounds are also able to pass through the membrane into the permeate. The level of salt passage increases as the salt concentration of the water being desalinated increases, and therefore, more salt passes through at the end of the RO process where the salt concentration is higher than at the beginning.
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Seawater Use and Desalination Technology
The quality of water produced, therefore, depends on membrane rejection properties together with the degree of water recovery and the system design. Relatively small, uncharged molecules such as carbon dioxide, silica, and boric acid may penetrate the membrane and reduce water quality. Silica and CO2 concentrations in the permeate usually present no issues for downstream use where the presence of CO2 assists in dissolving CaCO3 in the stabilization process. The presence of silica can be a problem for specialized industrial uses such as the fabrication of microelectronic components, where ultrapure water is required. However, salt and boron permeation through RO membranes remains problematic. Seawater contains approximately 5 ppm boron, which reduces to a value slightly greater than 1 ppm of boron after treatment through a RO membrane. Boron is an important component for plant growth; however, high concentrations are harmful and cause significant reduction in yield for many crops. Boron may be removed from water by ion exchange (Nadav, 1999), by using secondary or higher RO stages (Redondo et al., 2003; Glueckstern and Priel, 2003; see Figure 9; Faigon and Liberman (2003)), by increasing the pH of the water on the feed side of the membrane (RodriguezPastor et al., 2001; Prats et al., 2000), and by using electrodialysis reversal (EDR) applied on the product. A combination of these techniques is also suitable (Sagiv and Semiat, 2004). The current demand in Israel is to produce water containing less than 0.4 ppm (of boron) in Ashkelon and less than 0.3 ppm in Hadera – a plant that has been operating since early January 2010. The reason for this is related to the recovery of wastewater that remains after the use of desalinated water. Boron reaches the wastewater from different sources, and this may damage crops irrigated by treated wastewater. This requirement for low boron concentrations in the treated water has resulted in Ashkelon using up to three permeate stages of RO membranes to remove boron to concentrations below 0.4 ppm. This also results in a significant reduction of the dissolved salt concentration. Reduction of salt content in the product obtained after seawater reverse osmosis (SWRO) is a by-product of boron removal by a second RO stage applied on the product. If boron removal is performed using another technique, a secondary RO stage is still needed to improve the TDS quality, as concentrations of above 600 ppm are still found after the first pass. The types of membranes used in each pass differ, with SWRO membranes used in the first pass where the concentration of salt is very high and brackish water reverse osmosis (BWRO) membranes used in the second pass where the salt concentration is lower. SWRO membranes have higher rejection of salt, while BWRO membranes operate at higher flux enabling cost reductions to be achieved. Typical operating fluxes are 13–17 l m2 h1 in the first SWRO stage and 30–40 l m2 h1 in the second BWRO stage. The improvement in membrane rejection of boron and salts has reduced this problem; however, there is still a need for a BWRO second stage to partially treat the product water in order to produce water of a satisfactory quality. The high-pressure driving force for SWRO is reduced along the module in a single-pass system. The mode of operation using six to eight membranes in a module, at a constant pressure (minus the friction along the membranes) reduces
the driving force applied to desalinate the water. The difference between the operating pressure and the osmotic pressure at the first membrane may be 40–45 bars, while at the exit, on the last membrane it is of the order of 15–20 bars. It is obvious that the flux declines along the module to less than 30% of the starting flux of the first membrane. The flux depends on the operating pressure, the operating temperature, as well as on the salt concentration. Different membranes are available in the market; yet the order of magnitude of flux is up to 40 l m2 h1. The difference in driving force causes uneven load on the different membranes, that is, higher pressure drop at the entrance causes significant concentration polarization (CP) close to the membrane surfaces at the entrance to the membranes modules, and hence increased fouling. Operating an increased number of passes allows changes in the operating pressure to get a better distribution of the driving force. Theoretically, using a large number of brine stages while maintaining low operating pressure, just above the osmotic pressure, may save some energy in the process, down to the theoretical minimal thermodynamic energy for separation, but at the expense of the equipment, which would have to bear a high number of membranes and pressure vessels. Between the two extremes, it is possible to use two or three passes to gain some energy reduction and improved operational conditions. A two-stage operation can be used with two pumps that feed the two stages in a series; the first pump elevates the feed to 35–45 bars, while the second pump takes the concentrate of the first stage and increases the pressure for the second stage. The second stage may be operated by a turbine that is based on the brine of the first stage to increase the pressure to the second stage (e.g., Tuas SWRO Plant, Singapore). Membrane fouling and cleaning. Despite all precautions taken to improve the quality of water entering the membranes, membrane fouling occurs and frequent cleaning is required. While salt precipitation mainly occurs in brackish-water membranes, corrosion products may also accumulate, usually close to the spacer between the membranes. Suspended matter that was not removed during the pretreatment process may also accumulate on membrane surfaces, along with dissolved organic matter that concentrates in the thin layer close to the membrane skin. The latter may enhance biofilm growth on membranes if bacteria also reach the membrane surface. Accumulation of a fouling layer on the membrane is a process that enhances itself, with the rate of fouling increasing as fouling progresses. Clogging of a membrane reduces the flux through the membrane and the overall performance of the plant. All this calls for frequent membrane cleaning. The exact timing is dictated by the need to control fouling to a low level, while reducing the frequency of cleaning to maintain production capacity and minimize the consumption of chemicals. Initial acidification of the seawater feed may release CO2 from the water and may reduce the possibility of CaCO3 precipitation. The accumulation of organic matter may be removed partially by using sodium hydroxide (saponification). Membranes are also frequently cleaned with acid that releases some contaminants from membrane surfaces, particularly inorganic foulants. The cleaning process needs to be performed while the solutions are flowing along the membrane in order to remove fouling species that are released
Recovery 45% SWRO
Recovery 45% SWRO
Recovery 45% SWRO
2nd RO
Front product
2nd RO
Front product
2nd RO
Front product
3rd RO
4th RO
Boron select. IX
Weak acid IX
3rd RO
4 Stage boron removal system
25%
50%
25%
34%
36%
30%
10%
Boron <0.4 ppm Chloride ~ 120 ppm
2nd RO recovery – 80–85% Overall recovery – 41.4%
2 Stage boron removal system + Selective IX
Boron <0.4 ppm Chloride ~ 40 ppm
2nd RO recovery – 50% 3rd RO recovery – 90–95% Overall recovery – 44.1%
3 Stage boron removal system + WAIX
Boron <0.4 ppm Chloride ~ 35 ppm
2nd RO recovery – 80–85% 3rd RO recovery – 90–95% 65% 4th RO recovery – 90–95% Overall recovery – 44.5%
25%
Figure 9 Boron-removal processes. From Redondo J, Busch M, and De-Witte JP (2003) Boron removal from seawater using FILMTECTM high rejection SWRO membranes. Desalination 156: 229–238. Adapted from Faigon M and Liberman B (2003) Pressure center and boron removal in Ashkelon desalination plant. IDA World Congress on Desalination and Water Reuse (BAH03-181), International Desalination Association, Paradise Island, Bahamas.
Scheme C
Scheme B
Scheme A
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Seawater Use and Desalination Technology
during the cleaning process. Excessive suspended solids in this stage may accumulate in the spacer between the membranes and clog the flow path. Unlike other filters and even MF/UF membranes, NF and RO membranes are not subjected to a backwashing process. The main reason is the fear that the delicate layer that determines the separation properties of the membrane may be removed during the backwash process, causing severe damage to the membrane. However, it was found recently, that controlled osmotic backwashing, applied without damaging pressure, may be useful in removing contaminants from the membrane. The technique may be applied in different ways:
•
•
•
•
Shut down the feedwater occasionally for a short time. This will allow immediate osmotic backwash of the membrane. Water will penetrate the membrane at fluxes that are a function of the local salt concentration along the membrane. More water will penetrate the high-concentration locations that are more prone to scale deposition and might dissolve small, precipitated scale on the membrane (Sagiv and Semiat, 2005; Sagiv et al., 2008). Reduce the operating pressure of the concentrate side of the membrane to a pressure below the osmotic pressure in the system. Water will penetrate the membrane but the backwash flow rate will be reduced. Allow a wave of highly concentrated solution to pass through the feed channel without changing the operating condition. Backwash flow will increase but it is necessary to maintain a stock of a highly concentrated solution for this task (Liberman, 2004a, 2004b). Increase the permeate pressure to a level that allows back flow. This pressure should be below the concentrate side pressure on the feed side. Masaaki and Toshiyuki (2001)issued patents on a similar backwash using air pressure from the permeate side. However, this requires high-pressure piping on the permeate side of the membrane as well, and this will increase equipment cost significantly.
4.04.2.4 Desalination Process Costs Most of the evaporation processes were built along the coast of the Persian Gulf and some were also built on remote islands. Usually, they are connected to power stations. The real cost of these systems is not clear due to different calculations based on the real cost of energy. Estimations for SWRO are about US$0.55–1 m3 of freshwater produced. There are promising signs for reducing desalination costs by analyzing the cost components. Table 2 presents an estimated cost breakdown of desalinated water produced in a modern plant, taking into account partially, the changing trend of energy costs. The main constituent is, of course, the capital and financial cost. This is composed of the cost of the main items of the equipment: feed tanks, pretreatment filtration units, pumps, turbines and piping, controls, membranes, and membranes housing post-treatment and product tanks. One way to reduce desalination costs is to seek possibilities for cost reduction in each of the above items of equipment. Some items of equipment are restricted to the desalination industry and their cost may simply go down due to market forces. Investment in sophisticated automation and
Table 2
Cost estimation – modern project
Item
Cost (%)
Capital Energy Chemicals Manpower Replacement parts Membranes replacement Insurance Overhead
34 38 5 3 9 5 1 5
Table 3
Power usage in RO sweater plant with partial second stage
Item
Cost (%)
High-pressure pump High-pressure pump, 2nd unit Product transfer pumps Seawater supply Pretreatment system Miscellaneous
80.6 3.8 6.7 4.5 2.6 1.8
From Wilf M (2004) Fundamentals and cost of RO–NF technology. In: Proceedings of the International Conference on Desalination Costing, pp. 18–31. Limassol, Cyprus.
control equipment can result in lower water costs by maintaining stable high throughputs and savings in manpower costs. As can be seen from Table 2, manpower costs are no longer significant, since modern desalination plants may operate largely unattended. Energy is the main cost component to consider. The energy cost can be reduced by the use of a dedicated gas-turbine power station. A dedicated power station is more efficient than the grid because it is insensitive to the familiar sine-curve power consumption, due to fluctuations between day versus night and summer versus winter electricity demand. Modern devices for energy conservation also act to reduce energy cost though at the expense of increased capital cost. Wilf (2004) presents in Table 3 all the energy-demand components in a two- pass RO desalination plant. Others may claim that the relative low-pressure pumping costs are higher due to the distances to and from the plant. More information on RO costing history can be found in Glueckstern (2004). Costing information on the Ashkelon plant, which has been the largest RO desalination plant in the world for the last 4 years may be found in Kronenberg (2004) and Velter (2004). Figure 10 shows a picture of the Ashkelon plant, which has an operating capacity of 108 million m3 yr1. Compliance with proper operational procedures and implementation of a careful maintenance program can also reduce desalination costs by minimizing replacements of damaged membranes, lessening the use of cleaning chemicals, and reducing the inventory of membranes and spare parts. Membranes are often replaced after 5 years, but good pretreatment and cleaning practices can extend membrane life to 10 or more years (Montgomery et al., 2006). The design of a desalination plant in which it is envisaged that the operators are insufficiently trained, will invariably be based on exaggerated safety factors. Well-trained and
Seawater Use and Desalination Technology
87
Figure 10 The Ashkelon plant – 100 million m3 yr1 during 2005. From IDE Technologies Ltd. http://www.ide-tech.com/projects/ashkelon-israel-botswro-plant (accessed April 2010).
experienced operators can extract a higher production capacity from a desalination plant by identifying and debugging bottlenecks. Electrodialysis (ED), or EDR, is operated by applying a direct current (DC) electrical field across membrane stacks. Ions are transferred through semipermeable membranes into concentrated streams, leaving behind dilute salt solution. This was considered to be a promising technique, mainly attributed to the relative insensitivity of the membranes for fouling, and due to the thermodynamic transfer properties of this technique. Unfortunately, the technique did not succeed in taking the naturally expected position among other processes. Currently, the technique is used mainly for brackish-water desalination, and water purification (Thampy et al., 1999). EDR membranes are also used to remove special salts such as nitrates from slightly polluted waters. Strathmann (2004) reports the costing of the ED process.
includes an increase in the pH level, addition of Ca (up to the level of about 100 ppm as CaCO3), and alkalinity, namely HCO 3 (also to a level up to 100 ppm as CaCO3), according to local water regulations. This is often achieved by carbon dioxide addition or sulfuric acid addition to limestone. The need for addition of magnesium salt is currently under review by the WHO, as magnesium helps prevent heart disease. Additionally, trace amounts of sulfur are required to ensure good plant growth. Desalinated seawater contains bromide, and when disinfected with chlorine, brominated by-products such as bromochloroacetonitrile may form. In most applications, the presence of these compounds is sufficiently low so as to not compromise the final water quality (Agus and Sedlak, 2009), but it should be considered when using seawater desalination plants.
4.04.2.6 Environmental Aspects 4.04.2.5 Quality of Water Produced Thermal processes typically produce water containing between 5 and 50 ppm of TDS, with the relative concentration ratios of mineral ions similar to that in the feed seawater. Therefore, the problem of high boron concentrations in the product water does not exist. Feedwater containing dissolved volatile organic compounds, however will generate, unless special care is taken, water contaminated with the same components. This is true for both RO and evaporation techniques.
4.04.2.5.1 Increase in water hardness/water stabilization The product water of both thermal and membrane-desalination process is aggressive, tends to corrode iron pipes, and dissolves protective layers containing calcium and other salts on the inner sides of the mains. This may result in the phenomenon called ‘red water’, which is a release of corrosion products by water that dissolves the pipes’ protective layer of CaCO3. Water needs, therefore, post treatment that usually
Desalination processes may be characterized by their effluent discharges to the environment, the air, the nearby land, and to the seas. Desalination is dependent on energy and usually uses energy derived from fossil-fuel sources. Air pollution is associated with energy production, that is, emission of NOx, SO2, volatile compounds, particulate, CO2, and water, etc., either by using electricity produced by a conventional power station or by using a dedicated power station. Operation of dedicated gas turbines at high-efficiency levels will reduce the amount of contaminants to the atmosphere. Several desalination plants in Australia have offset their energy use with renewable energy, such as the Perth desalination plant (Crisp, 2009). The Perth plant included the construction of an 82-MW wind farm. While this produces sufficient power to drive the desalination plant, the intermittent nature of the power source means that base-load power sources are still required. This increases the cost of water as electricity is a significant cost in desalination operations, but it is
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Seawater Use and Desalination Technology
used to offset the carbon dioxide emissions associated with desalination. Effluents of desalination plants contain relatively highly concentrated water, which depend on the water recovery from the feed brine. In case of seawater desalination, rejected brine is concentrated almost to twice the concentration of the original seawater solution. The concentrate also contains chemicals used in the pretreatment of the feedwater. The latter may contain low concentrations of anti-scalants, surfactants, and acid added to the feedwater that reduces the pH. To this may be added occasional washing solutions or rejected backwash slurries from feedwater pretreatment. At small-scale operation, the problem is mild and no serious damage to marine life is likely. At large scales of water production, the problem is more serious; however, dilution and spreading of effluents usually solves the problem. Natural chemicals that do not harm the environment will probably replace the added chemicals, in the future. The current trends in concentrate disposal is to transfer it into the deep sea, sending jets upward in a few directions, at an angle to the horizon, enabling the concentrate to be diluted very close to the concentration of the sweater. Another approach is to mix the concentrate with the cooling water leaving a large power station before discharge. These techniques minimize the influence of the concentrate on the sea environment.
4.04.2.7 Energy Issues Desalination, as a separation process, needs energy. The current specific energy for RO desalination was reduced significantly during the last decade and it is now not far from the limiting theoretical thermodynamic minimum. This has been achieved by the development of large pumps with efficiencies as high as 92%, and modern efficient turbines and energyrecovery devices. The newer devices are the turbocharger, pressure exchanger, or work exchangers – the names adopted by different producers represent efficient ways to recover the energy content of the high-pressure concentrate. More details on energy-recovery devices may be found in Voutchkov and Semiat (2008). Turbines are used to turn the concentrate pressure into the velocity of jets that spin a wheel. This is used either to reduce the power consumption of the motor that drives the pump, or in conjunction with the turbine pump, to boost the pressure of the feed to a second stage. There were other methods that were used to exchange the pressure of the brine concentrate by simple devices that transfer the pressure to the seawater feed. Using these new techniques, significant reductions in power consumption have been achieved. For example, processing of Mediterranean seawater at a recovery of 50% needs only 2.7 kW h1 m3 produced by recovering the concentrate energy with turbines. Pressure exchangers can go even lower to 2.2 kW h1 m3 water produced. Since more energy is consumed for the feed and concentrate pumping and for the pretreatment stages, the overall energy needs are below 3.7 kW h1 m3 produced from seawater. The energy cost of an optimized desalination plant is around 30–40% of the total cost of water. The cost is based on optimization of the operation and the exchange between energy and equipment. This optimization is made during the
design of the plant; yet, the cost of energy may vary significantly during the lifetime of the project. During the course of writing this chapter, the cost of natural oil increased significantly in comparison with the cost at the design stage of, say, the Ashkelon plant. It is difficult to change the optimal design of the plant after it is built. However, using RO membranes while considering possible changes of energy cost, it is possible to minimize the losses by designing for lower possible energy consumption at the expense of equipment costs. A 100-million m3 RO based seawater desalination plant demands an electrical energy supply of less than 40 MW. A dedicated power station can work at much higher efficiency than a regular power station for this purpose, since it is operated constantly without the known sine wave that represents day–night and summer–winter changes in consumption. Higher efficiency is expected for gas turbines, since the high temperature of the gases may also be used. Therefore, the real energy needed is lower than that for other common uses. Environmentalists are often heard criticizing the levels of energy consumption for water desalination. Water is needed for the many people on Earth and for meeting their basic needs. This is of higher priority for the use of energy compared to its use in air-conditioning and/or large, high-energy consuming cars. More on energy consumption and comparison with other forms of energy usage may be found in Semiat (2008). Environmental concern arising from the CO2 green-house effect associated with the use of hydrocarbon fuel, has led to the goal of supplying desalination energy from renewable energy sources. While writing this chapter, the cost of a barrel of oil is higher than US$70. With this trend, renewable energy sources may be soon compatible and economic for general electricity production. At this stage, they will also be suitable for desalination purposes. No doubt, more effort should be directed toward the use of renewable sources of energy. The real test, however, for any new source of energy is its acceptance as a common source of energy. The savings on CO2 emissions need to replace other forms of energy use and not be used for the very delicate issue of desalination for freshwater production. Using nuclear energy, which is currently more expensive than fossil-fuel-derived energy, is dangerous in areas where political instabilities exist. It is also problematic where the technology is not available locally and the technology, expertise, and skilled manpower need to be imported. A possible way for efficient use of energy in a sufficiently large desalination plant is to design a hybrid plant consisting of a membrane unit and/or a VC unit (Awerbuch, 1997b), using electrical energy, and a multi-effect evaporation plant, using heat. Such an operation is common in the chemical industry. The energy costs are minimized by coupling the desalination plant with a dedicated power plant generating electricity and waste heat at optimal economic conditions. One of the benefits that can be claimed from the day– night, summer–winter electricity-production cycle is that it can produce desalinated water during the night when lower power is consumed. The main disadvantage is that the desalination equipment will not be in use for a high percentage of the time. This is unlikely to be economical, since, as in any modern plant, the production cost is more expensive if the equipment is not in full use. In other words, an efficient
Seawater Use and Desalination Technology
desalination plant needs to be operated round the clock, 24 h a day, 365 days a year, with exceptions for maintenance only. During this time, it needs the full supply of energy, at the lowest cost. Since energy is so important for desalination, a few comments on possible energy use that may significantly reduce desalination costs are made. With respect to the use of spent energy from large steampower plants, it is well known that steam cycle power stations purge large amounts of energy at the steam-condensing stage at the turbine exit. This source of heat may be combined with a thermal desalination technique in order to supply the primary steam, as in MSF and MED (Awerbuch, 1997a, 2004). A modern, efficient power station releases the exhausted steam at around 35–40 1C, which is too low for the proper operation of the desalination plant. It is required, therefore, to release steam at elevated pressure and temperature using backpressure turbines that fit in with the desalination-plant needs. This, of course, will cause some loss of production at the power-generation plant; therefore, there is a need to integrate and optimize the two processes together. Such designs are extremely difficult to perform if two different authorities are involved – power and water production (El-Nashar, 1997). This type of hybridization was successfully employed in the Persian Gulf countries when the same authority controlled the two industries. However, there is always a difference between the demand for electricity and the demand for water, and this is the main reason it is not implemented in other places.
4.04.3 Brackish Water 4.04.3.1 Brackish Water Desalination Applications Brackish water, generally defined as water with TDS content between that of freshwater (r500 mg l1 TDS) and seawater (33 000–48 000 mg l1 TDS), can occur naturally as brackish groundwater in subsurface saline aquifers, as surface water due to natural erosion, or as a result of seawater mixing with river water (in estuaries) or groundwater (in coastal aquifers). Natural brackish water, particularly brackish groundwater, exists in most continents in quantities almost equal to or more than fresh groundwater and surface waters combined (Shiklomanov, 1993). Human activities can also cause fresh surface water and groundwater resources to become brackish through consumptive use and increase in their salt loading. For example, excessive groundwater pumping from coastal aquifers can cause salt-water intrusion that extends the brackish water zone of mixing inland, while saline return flows from irrigated agricultural lands can increase the salt loading of surface waterways. Some agricultural irrigation practices such as subsurface tile drainage often generate agricultural drainage waters that are highly brackish. Similarly, mining activities can also generate mine-drainage waters that are brackish and contaminated with heavy metals. In the past, brackish-water desalination applications have been limited to small-scale municipal and industrial applications. It is especially popular in the USA as it accounts for the majority (B77%) of the nation’s total online-desalination capacity (Committee on Advancing Desalination Technology, 2008). One of the early large-scale inland brackish-water
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desalting plant, the Yuma Desalting Plant in Arizona (USA) (Lohman, 2003), was completed in 1993 for the purpose of supplementing Colorado River water deliveries to Mexico (via desalting brackish agricultural water return flows; B2500 mg l1 TDS). Although its production capacity remains the largest in the world for a brackish-water desalting plant (272 500 m3 d1 or 72 million gallons a day (MGD)), the plant was operational for only two occasions in 1994 and has remained offline (although well maintained). With dwindling freshwater supplies and maturing RO/NF and ED/EDR process technologies, desalting of under-utilized brackish groundwater and surface-water resources has attracted significant interest. Some recent (2000–10) large-scale brackish-water desalting installations include the Aigu¨es Ter-Llobregat’s (ATLL) Plant (Spain; 220 000 m3 d1), the Al Wasia Plant (Saudi Arabia; 200 000 m3 d1), the El Atabal Plant (Spain; 165 000 m3 d1), the Wadi Ma’in Plant (Jordan; 135 000 m3 d1), and the K. B. Hutchison Plant (USA; 104 000 m3 d1). All of these are RO plants, except for the ATLL plant, which is an EDR plant. Feedwater quality plays a major role in both the design and operation of brackish-water desalination processes. Brackish waters vary greatly in ionic composition and content, both temporally and geographically, depending on hydrogeologic conditions and related human activities (i.e., industrial or agricultural). For example, in California’s San Joaquin Valley, one of the most productive agricultural regions in the US, tile drainage of irrigated agricultural lands generates brackish waters with a wide salinity range (3000–30 000 mg l1 TDS). In Texas (USA), subsurface aquifers with salinity ranging from 1000 to 10 000 mg l1 TDS have been estimated to hold as much as 3 trillion m3 of brackish groundwater. Major solutes in brackish waters, such as sodium, chloride, calcium, sulfate, and bicarbonate ions, typically originate from water reactions with minerals such as halite, gypsum, anhydrite, calcite, and dolomite. Other common minor solutes include silicates, iron, strontium, barium, fluoride, selenium, and boron. Some examples of brackish-water composition are listed in Table 4.
4.04.3.2 Brackish Water Desalination Technologies The choice of brackish-water desalination technologies and process configuration is a site-specific combination of many factors, including source-water quality, target productivity, product-water quality requirements, brine-disposal options, and other local conditions and regulations (e.g., permits). Pressure-driven membrane processes of RO and NF are predominant brackish-water desalting technologies, while electrochemically driven membrane processes of ED and EDR still maintain important niche applications. In applications where high-purity product is of importance, electrodeionization (EDI) processes have been integrated for product water polishing. Cross-flow RO/NF operation relies on applying sufficient pressure at the feed side of ion-rejecting RO/NF membranes to overcome the hydraulic resistance of water permeation, as well as the osmotic pressure difference between the feed and permeate side of the membranes. The world’s first municipal RO plant was commissioned in 1965 in Coalinga (CA, USA) for desalting brackish water, just shortly after the first cellulose acetate RO membrane (with practical hydraulic resistance and
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Table 4
Composition of brackish water from a variety of sources (USA)
Analyte (mg l1)
El Paso Water Utilities, TX Airport Wellsa
Panoche Water District (San Joaquin Valley, CA), DP-25 Wellb
Indian Wells Valley Water District, CAa
Colorado River Water, Yuma, AZc
TDS Arsenic Barium Bicarbonate Boron Bromide Calcium Chloride Fluoride Magnesium Nitrate Potassium Selenium Silica Sodium Strontium Sulfate
3170 75 0.05 176 1370 0.61 38.4 0.11 15.9 29.4 745 301
8500 274 23.5 492 1190 255 337 4.3 0.47 31.4 1810 78 4080
1630 0.0052 370 1.74 164 236 1 49 72 6.1 0.059 45 333 1.55 570
941 0.1 212 95 164 34.5 11.6 165.5 1.24 322
a
Data on ElPaso Water Utilities, TX and Indian Valley Water District, CA, form Committee on Advancing Desalination Technology (2008) Desalination a National Perspective, Science and Technology Board, National Research Council. Washington, DC: National Academies Press. b Data on Panoche Water District, CA, from Cohen Y and Christofides P (2010) Reverse Osmosis Field Study, Final Report, DWR-WRCD Agreement 46000534-03, Task Order No. 22, California Department of Water Resources, 16 June 2010. c Data on Colorado River Water, Yuma, AZ, from Rahardianto A, Gao J, Gabliech CJ, Williams MD, and Cohen Y (2007) High recovery membrane desalting of low-salinity brackish water: Integration of accelerated precipitation softening with membrane RO. Journal of Membrane Science 289: 123–137.
salt rejection) was invented at the University of California, Los Angeles (UCLA) (Loeb, 1984). Over the past two to three decades, significant advances in polyamide thin-film-composite (TFC) membranes have resulted in a new class of RO and NF membranes with high water permeability (i.e., low hydraulic resistance) and salt rejection, enabling cross-flow RO/NF to operate at applied pressure levels approaching the limit imposed by thermodynamics (i.e., the osmotic pressure difference at the concentrate end). With larger pores than the ones in RO membranes, NF membranes can selectively reject divalent ions and molecular solutes over monovalent ions. Although fouling-resistant RO/NF membranes remain elusive, improvements in TFC membrane designs (i.e., lower surface roughness and near-neutral surface charge) have made present membranes less prone to fouling by suspended particulates/ colloids. Present commercial RO/NF membranes are packaged as modular spiral-wound membrane elements of standardized dimensions, with fluid channels formed by feed and permeate spacers. An RO element can typically operate up to 15% water recovery with a nominal salt rejection of about 98–99.7%. Operating pressures are normally in the range of 10–41 bars (150–600 pound force per square inch (psi)) for brackish water RO membrane elements, depending on feedwater salinity (i.e., osmotic pressure). Due to its relative simplicity and ease of operation and maintenance, RO has become the primary workhorse in brackish-water desalting. NF elements usually operate at much lower pressures (B7 bar/100 psi), with divalent ion rejection of 50–98% and monovalent ion rejection of 20–75%. As a stand-alone or a feedwaterpretreatment system, NF is increasingly being applied for the removal of hardness ions (e.g., calcium and magnesium),
organics, and specific contaminants (iron, nitrates, pesticides, herbicides, etc.). As alternatives to RO and NF, electrodialysis (ED, EDR, or EDI) relies on electric field to transport ions from diluate (feed) compartments to concentrate compartments across flat-sheet ion-exchange membranes. In an array of alternating cation and anion membranes, separated by alternating diluate and concentrate compartments, anions/cations migrate from diluate compartments toward anode/cathode plates, passing through anion-exchange/cation-exchange membranes, and become trapped in concentrate compartments due to rejection by adjacent cation-exchange/anion-exchange membranes. With the development of the first ion-exchange membranes in 1948, the first commercial ED unit was built by Ionics and deployed at an oil-field campsite in Saudi Arabia in 1953 (Reahl, 2006). The first ED plant in the USA was erected in the city of Coalinga (1958) (Reahl, 2006) less than a decade before the operation of the world’s first RO plant at the same city (1965) (Loeb, 1984). In the mid-1970s, Ionics introduced the EDR process as an improvement on its ED process, whereby DC flow through the ED membrane stack is periodically reversed, along with simultaneous interchange of the product and brine stream flows (Reahl, 2006). The periodic reversal is touted to minimize the formation of mineral scale on ion-exchange membranes. Furthermore, uncharged materials (e.g., some forms of silica) do not accumulate near membrane surfaces. As ion removal is readily controlled by the applied voltage, ED/EDR processes are also relatively more flexible for producing a wider range of product-water purity than RO. ED/EDR processes, however, become less efficient with increasing product-water purity (i.e., due to reduced
Seawater Use and Desalination Technology
solution conductivity); ion removal is therefore typically in the range of 50–95%. ED and EDR processes remain important in niche applications, especially for treating low-salinity brackish-water sources (o3500–4000 mg l1 TDS) where a cost advantage over RO/NF may be present. EDR is sometimes chosen over RO in desalting challenging feedwaters with high membrane fouling/scaling tendency (e.g., waters with high silica content) at high water-recovery levels. Another important development in electrodialysis was in the late 1980s, when EDI was first commercialized (Grebenyuk and Grebenyuk, 2002). EDI incorporates ion-exchange resins within ED compartments to enhance ion transport and provide a substrate for electrochemical reactions. Specifically, target ions displace Hþ and OH ion sites at ion-exchange resins within the feedwater (diluate) compartments before migrating through the anion-exchange/cation-exchange membranes into the concentrate compartment. In addition to driving the migration of ions, the applied electric field also drives water-splitting reactions that continuously regenerate ion-exchange resins in situ (with Hþ and OH ions), eliminating the need to regenerate ion-exchange resins using additional chemicals. While EDI cannot be reliably used to directly desalt brackish water, it has important applications in RO permeate water polishing with respect to specific contaminants – a less-chemical intensive alternative to conventional mixed-bed ion-exchange process.
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number of pressure vessels decreases from one stage to another in order to compensate for the decline in the retentate stream cross-flow velocity in the axial downstream direction (i.e., due to water recovery). In order to cope with increasing osmotic pressure in the downstream direction, inter-stage booster pumps in multistage processes can be employed to operate RO/NF stages at increasingly higher pressure ranges. Such use of inter-stage booster pumps, from a thermodynamic viewpoint, allows RO/NF to operate closer to the reversible process and thus consume less energy for desalting (Zhu et al., 2008). A two-stage system with a 2:1 array (i.e., the first stage has twice as many pressure vessels than the second stage) is common for brackish water desalting at water-recovery levels in the range of 60–80%, while three stages (3:2:1 array) are needed for higher water-recovery levels (Figure 11). If an inter-stage booster pump in a two-stage systems is used, permeate production at equal water recoveries in the first and second stage would correspond to the energy-optimal operation (Zhu et al., 2008). It is noted that, a two-stage configuration with a booster pump may require a larger totalmembrane area than without a booster pump in order to achieve the same product water recovery. However, when high water recovery is targeted in brackish water desalting, the additional membrane cost is usually offset by the reduction in energy cost associated with using an inter-stage booster pump (Zhu et al., 2008).
4.04.3.3 Common Process Configuration The present RO/NF and ED/EDR systems consist of modular building blocks that can be configured to meet productivity and product-quality requirements. These configurations are briefly discussed next.
4.04.3.3.1 RO/NF process configuration A typical RO/NF system for brackish water desalting consists of two or more stages, with each stage consisting of pressure vessels arranged in parallel (Figure 11). Each pressure vessel can usually accommodate up to six to seven spiral-wound RO/ NF elements. The system is typically designed so that the
4.04.3.3.2 ED/EDR process configuration A typical ED/EDR process employs plate-and-frame stacks of membrane cell pairs, with each cell pair consisting of an anion-exchange and cation-exchange membrane that are separated by concentrate and diluate stream spacers. These plateand-frame membrane stacks can be arranged as a series of one or more hydraulic stages (Figure 12). Electrical staging can also be done to improve system performance and flexibility (Figure 12). The number of stacks in series (and thus membrane surface area), in addition to current density, determines the product (the final diluate stream) quality and thus the
Single stage Two-stage (2:1 array) C
F Pump
C
F Pump
P P Three stage with booster pump (3:2:1 array) F
Booster pump
C P
Figure 11 Typical arrangements of pressure vessels for RO/NF membrane elements. F, Feed; C, concentrate; P, Permeate.
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Seawater Use and Desalination Technology
+ HS-1
+ +
ES-1
HS-1 HS-2
HS-1 HS-2
−− − −
HS-3 ES-2
+
Figure 12 Examples of hydraulic staging (HS) and electrical staging (ES) of ED/EDR membrane stacks. Adapted from US Bureau of Reclamation (USBR) (2003) Desalting Handbook for Planners, 3rd edn. Washington, DC: USBR.
water recovery. Each membrane stack can typically reduce salinity by up to about 60% (Reahl, 2006). Single- and twostage systems are most common, but three- and four-stage systems can also be competitive (compared to RO/NF; Reahl, 2006).
4.04.3.4 Major Challenges Achieving and maintaining optimal product-water-recovery levels (with respect to overall desalination operating costs) are perhaps the most important challenges in the application of membrane technology (RO, NF, ED, and EDR) for brackishwater desalting. Product-water-recovery level dictates the volumetric rate of desalted water production relative to that of residual concentrate waste generation. With increasing product water recovery, the volume of residual concentrate waste is reduced, increasing the available options for residual concentrate management (i.e., treatment and disposal). Optimal product-water-recovery levels in brackish-water desalting are highly dependent on feedwater quality, target production capacity, and locally available methods of concentrate disposal. As the costs associated with managing residual desalination concentrate is typically high, especially at inland locations, high levels of product water recovery (85–95%) are often required for optimal desalting operations. Effective feedwater pretreatment for preventing membrane fouling (i.e., colloidal/particulate deposition on membranes and feed-spacer blockage) is an operational prerequisite in membrane-based desalting (i.e., RO/NF/ED/EDR) of brackish water. Typical foulants in brackish water include suspended and colloidal particulates/organic matter, as well as dissolved organics and biological entities (that may contribute to organic fouling and biofilm formation). Particulates/organic
matter removal via conventional coagulation, flocculation, and sedimentation processes followed by media filtration are common in brackish-water desalting applications. MF and UF processes, however, are increasingly being applied due to their superiority in providing stable influent quality to membranedesalting operations with respect to particulates, colloids, as well as bacteria. In-line coagulation and media filtration are sometimes used in conjunction with MF/UF in order to enhance contaminant removal, minimize MF/UF pore plugging, and reduce MF/UF particulate loading (Huang et al., 2009). In some cases, media filtration may be sufficient to remove particulates due to very low concentrations of suspended solids and organic matter (e.g., as in the case of some brackish groundwater). In other cases, pretreatment to remove specific constituents such as dissolved iron, manganese, and sulfides may be necessary as their oxidation may lead to in situ precipitation in membrane systems (US Bureau of Reclamation, 2003). Feedwater disinfection by chlorination/ chloroamination is also used sometimes when biofouling is of a concern (e.g., in brackish wastewater). However, dechlorination may be necessary prior to membrane-desalting operations because, unlike present EDR membranes which have high chlorine resistance, present polyamide TFC RO membranes can only tolerate very low levels of chlorine/ chloroamine residuals (i.e., carryover from pretreatment) due to oxidation of the polyamide active layer. Careful selection of plant equipment and piping materials are also necessary to avoid material leaching that can contaminate feedwater with foulants or membrane-damaging substances. For example, contamination of RO feedwater with phthalate ester from reinforced polyester pipe have been shown to cause fouling and damage to RO membranes (Hasson et al., 1996).
Seawater Use and Desalination Technology
Effective feedwater pretreatment methods for mitigating membrane fouling (particulates/colloidal fouling, as well as biofouling) are available and routinely employed in RO desalination. However, the main bottleneck that remains to achieving high product water recovery in brackish water desalting is membrane mineral scaling – the deposition and crystallization of sparingly soluble mineral salts on membrane surfaces (e.g., gypsum (CaSO4 2H2O), BaSO4, SrSO4, CaCO3, SiO2, etc.). Mineral scaling can occur in pressure-driven (RO/NF) and electrochemically driven (ED/EDR) membrane processes when dissolved mineral-salt concentrations near membrane surfaces are brought above solubility limits with increasing product water recovery. Mineral scale blocks membrane surfaces and thus degrades membrane performance (e.g., permeate flux decline in RO/NF and increase in electrical resistance in ED/EDR). The primary strategy of membrane-scaling mitigation is feedwater conditioning, which involves dosing of chemical additives to alter feedwater chemistry. Common feedwater conditioning methods include feedwater pH adjustment and anti-scalant treatment. Feedwater pretreatment to remove inorganic and organic particulates/colloids is also critical not only to minimize blockage of feed channels and particulate/colloidal fouling, but also to minimize the presence of solid surfaces that can promote the nucleation of mineral-salt crystals. Common feedwater pretreatment/conditioning methods do not remove mineral-scale ionic precursors of mineral scalants. These methods typically allow membrane-desalting operations to concentrate feedwater to limited solution supersaturation levels with respect to mineral scalants (Hydranautics, 2008). The upper limit of solution supersaturation levels, as constrained by the effectiveness of feedwater pretreatment and conditioning methods, impose a limit on product water recovery (i.e., the membrane mineral scaling threshold) at a level that is often suboptimal with respect to overall desalination operating costs. Furthermore, the difficulty in identifying membrane-scaling threshold in real time, coupled by temporal variations in feedwater quality, often leads to operation of RO/NF or ED/EDR processes at waterrecovery levels well below the membrane mineral scaling threshold. In order to achieve and maintain desalting operation at optimal water-recovery levels, several key challenges must be addressed, including managing the impact of feedwater-quality variability, early detection and mitigation of membrane mineral scaling, methods for enhancing water recovery, and management of residual-desalination concentrate (i.e., treatment and disposal).
Feed Q f,C f Pp
4.04.3.4.1 Concentration polarization and membrane mineral scaling One of the primary factors affecting membrane fouling and mineral-scale formation in RO/NF and ED/EDR processes is CP. As separation process takes place at the membrane– solution interface, the concentrations of solutes near the membrane surface are higher relative to the bulk solution. In cross-flow RO/NF processes, pressure-driven convective flux of solution toward the membrane, coupled by ion rejection and water permeation at the membrane–solution interface, leads to the accumulation of solute near the membrane surface, generating a concentration-boundary layer along the axial flow direction (Figure 13). CP enhances the osmotic pressure difference across the membrane, reducing the net pressure driving force for water permeation. In the case of ED/EDR processes, ions are transported from one solution compartment to another under the influence of an applied electric field, passing through or rejected by ion-exchange membranes. The rate of ion transport to ion-exchange membranes, and thus the efficiency of the separation process, is limited by mass transfer in the concentration-boundary layer that develops near the membrane–solution interface. In both RO/NF and ED/EDR processes, local hydrodynamics strongly affect CP, leading to spatial variation of solute concentrations near membrane surfaces. Mitigation of membrane mineral-scale formation must therefore consider not only average CP levels in membrane systems, but most importantly the local CP extremes that may occur, particularly at flow-stagnation points (Lyster et al., 2009; Rahardianto et al., 2006). Advanced numerical methods (e.g., two-dimensional (2D) and 3D computational fluid dynamics) can be used to elucidate the impact of CP on scale formation (Lyster et al., 2009; Lyster and Cohen, 2007). A simple experimental procedure has also been developed for predicting average CP levels in RO membrane elements, based on measurements of permeate-flux decline induced by the osmotic pressure of saline solutions (relative to a salt-free solution) (Sutzkover et al., 2000). With the establishment of solution supersaturation due to CP, the process of membrane mineral-scale formation can take place. Specifically, crystal nucleation and growth may occur in the concentration-boundary layer and directly on the membrane surface (Gilron and Hasson, 1987; Lee et al., 1999). Growth of deposited or surface-nucleated precipitates leads to the formation of impermeable mineral scale that progressively blocks membrane-active area. Mineral scalants commonly encountered in brackish-water desalting include gypsum, calcium carbonate, strontium sulfate, barium sulfate, silicates, U (x)
Cb H/2 Cm(x ) Jw(x)
93
Concentrationboundary layer
Concentrate Qc,Cc RO membrane Permeate Qp,Cp Pp
Figure 13 Cross-flow RO in a membrane channel. Qf, Qc, and Qp refer to volumetric flow rates of feed, concentrate, and permeate streams, respectively; Cf, Cc, and Cp represent solute concentration in feed, concentrate, and permeate streams, respectively; Cm(x) is the solute concentration at the membrane surface; Cb is the solute concentration in the bulk solution; and U(x) is the cross-flow velocity.
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and calcium phosphate. The kinetics and thermodynamics of mineral-scale formation and the resulting scale structure and morphology are influenced by water chemistry, solution composition, solution saturation levels, temperature, pressure, surface properties, and near-surface hydrodynamic conditions (So¨hnel and Garside, 1992), as well as the presence of crystallization retarders (e.g., antiscalants). Coagulant carryover (e.g., alum) from pretreatment coagulation can cause membrane scaling by associating with other components such as silica (e.g., aluminum silicates).
4.04.3.4.2 Mitigation of membrane mineral scaling The most common feedwater conditioning methods for mitigating mineral-scale formation are feedwater pH adjustment and antiscalant treatment (Taylor and Jacobs, 1996). Owing to the acid–base chemistry of carbonic acid, pH adjustment via acid dosing (HCl or H2SO4) can lower the supersaturation level of carbonate minerals by keeping carbonate ions protonated. Antiscalant treatment involves dosing of antiscalant chemicals that kinetically retard scale formation. Antiscalants do not prevent crystallization but delay nucleation and retard growth of mineral salt crystals to an extent that depends primarily on solution supersaturation level and antiscalant type and dose. Most antiscalants are formulations of polyelectrolytes (e.g., polyacrylic acids, carboxylic acids, polymaleic acids, organo-phosphates, polyphosphates, phosphonates, etc.) with molecular weight ranging from 2000 to 10 000 Dalton (Hydranautics, 2008). Dispersants are also commonly included in antiscalants formulations; their function is to keep bulk-formed crystals and colloids in suspension, minimizing their deposition and contribution to scale formation. The upper limits of antiscalant treatment effectiveness are usually specified by manufacturers in terms of a maximumsolution supersaturation level with respect to the specific mineral scalant of concern, quantified as SIx ¼ IAP/Ksp (where IAP and Ksp are the activity and solubility products for ions that form mineral salt x, respectively). Antiscalant manufacturers, for example, typically recommend that SIg (gypsum) and SIb (barium sulfate) be kept below 2.3–4 and 60–80, respectively, in order to ensure effective antiscalant treatment (Hydranautics, 2008). As membrane scaling is often a slow kinetic process that may involve multiple types of mineral scalants, solubility considerations alone may be insufficient for determining the appropriate type and optimal dosing of antiscalants. It has been reported that overdosing can cause certain antiscalants to precipitate out of solution (Hydranautics, 2008), as well as to increase biofouling potential (van der Hoek et al., 2000). Therefore, it is often necessary to resort to experimental methods to determine the specific antiscalant effectiveness for the water source under consideration. Antiscalants, for example, can be ranked based on dosage-induction time relationships for the expected maximum levels of supersaturation in the membrane systems of interest, using solution conditions of interest (e.g., composition, pH, and temperature) (Shih et al., 2004). The effectiveness of various types of antiscalants in preventing membrane mineral scaling has been assessed in membrane systems for the case of calcium carbonate (Hasson et al., 1998; Drak et al., 2000; Lisitsin
et al., 2005; Lisitsin et al., 2009), gypsum (Hasson et al., 2001, 2003), silica (Semiat et al., 2001, 2003a, 2003b), and calcium phosphate (Greenberg et al., 2005). Novel methods have also been developed to assess the effectiveness of antiscalants via direct optical imaging of membrane surfaces in real time during the membrane-separation process (Kim et al., 2009), as well as for rapid off-line membrane analysis (Rahardianto et al., 2006). In Figure 14, for example, the impact of aluminum chlorohydrate (ACH), ferric chloride, and poly-DADMAC (poly-diallyldimethylammonium chloride) flocculants/ coagulants on antiscalant effectiveness (Flocon 260) in retarding gypsum scale formation is compared based on optical images from membrane-scaling runs conducted at the same initial CP level (Kim et al., 2009). In addition to feedwater conditioning (i.e., antiscalant treatment and feedwater pH adjustment), adjustment of operating conditions to alter CP levels can also facilitate the mitigation of membrane mineral scaling. In EDR processes, for example, polarity reversal with simultaneous interchange of feed and concentrate flows periodically renews the CP layer and can therefore reset the crystallization induction time. In RO/NF processes, CP level is typically minimized by maintaining a reasonable level of permeate flux near the membrane-element-concentrate fluid exit and providing a sufficiently high value of cross-flow velocity. Recently, feedflow reversal in RO/NF operation has been demonstrated as an effective approach for mitigating membrane scaling in some brackish-water desalting applications (Uchymiak et al., 2009). Using this approach, periodic changes of the feed-flow direction reverses the axial direction of the concentration– boundary-layer development. In the forward-flow direction, membrane areas near the concentrate fluid exit, which are prone to scale formation (near fluid exit), are exposed to higher solute concentration. By reversing the flow, just before scaling occurs, the same membrane area becomes exposed to lower solute concentration, thereby resetting the crystallization induction times. Feed-flow reversal, however, is most effective when the feedwater is undersaturated with respect to the mineral scalants of concern. In addition to feed-flow reversal, the feasibility and effectiveness of osmotic backwashing of spiral-wound RO membrane elements have also been demonstrated (Sagiv and Semiat, 2005; Sagiv et al., 2008). Unlike pressure-based backwashing in MF and UF systems, osmotic backwashing involves changes in pressureconcentration conditions across the membrane to induce osmotic driving force for periodic reversal in permeate-flow direction. The method has been shown to be effective in minimizing membrane fouling without causing delamination of the polyamide active layer of RO membranes.
4.04.3.4.3 Managing the impact of feedwater-quality variation Feedwater-quality variability has important implications for brackish-water desalting since optimal RO/NF and ED/EDR plant designs are source-water dependent; plants operate best when the feedwater quality is consistent. Feedwater quality governs the required feed pretreatment to prevent membrane fouling and scaling, and the required applied pressure (for RO/NF) or electrochemical driving forces (for ED/EDR) for
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Figure 14 Optical images of membrane surfaces at three different times during membrane scaling tests: (a) without additives, and with the addition of (b) with antiscalant and aluminum chlorohydrate (ACH) H (c) antiscalant and ferric chloride, and (d) antiscalant and poly-DADMAC. Adapted from Kim M-M, Au J, Rahardianto A, et al. (2009) Impact of conventional water treatment coagulants on mineral scaling in RO desalting of brackish water. Industrial and Engineering Chemical Research 48: 3126–3135.
achieving a given water-recovery level. In developing a brackish-water desalting plant, for example, the feedwaterintake location and hydrogeologic conditions (e.g., salinity variation with depth and time for groundwater, etc.) must be investigated carefully given the potential for spatial and temporal variability in source-water quality. The desalting process must then be selected, designed, and configured to cope with site-specific challenges with respect to energy requirements, membrane-scaling mitigation, and residual concentrate waste management. Plant-operating conditions must be selected and adjusted to cope with temporal fluctuations in water-production level that may be imposed due to feedwater-quality variations, which in turn can significantly affect CP, membrane-scaling tendency, and energy consumption. Present RO/NF and ED/EDR plants commonly manage temporal variability in source-water quality by applying process control (automated or manual) for the sole purpose of maintaining process productivity. Process-control systems, however, can be designed to enable adaptive operation of these processes. For RO/NF processes, for example, optimal time-varying operating policy with respect to energy consumption has been
proposed (Zhu et al., 2009). Specifically, recent work has demonstrated that, in order to maintain a constant permeate flow in the presence of feed-salinity fluctuation, feedwaterflow rate and operating pressure can be selected to minimize energy consumption. The work suggests that, by applying a real-time optimization routine in a control system, adaptive operation of RO/NF processes with respect to temporal feedwater-quality variation can keep energy consumption at minimum. Present RO/NF and ED/EDR plants commonly operate at reduced water-recovery levels in order to enable safe operation when feedwater quality can drive the process toward the fouling and/or mineral-scaling thresholds. This conservative process operation is a precaution that must be taken since present traditional measures of plant-performance trends (e.g., primarily permeate-flux decline and salt passage in RO/NF desalting) do not guarantee sufficient early detection of membrane fouling and mineral scaling. Although various methods of scale and fouling detection have been proposed (Chen et al., 2004), it is only recently that real-time early detection of the onset of scale formation has become possible (Uchymiak et al., 2007). For RO/NF processes, for example,
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the use of a high-pressure flat-sheet membrane cell, either with a transparent window (Uchymiak et al., 2007) or a completely transparent RO/NF cell (Rahardianto et al., 2008), allows real-time digital imaging of the membrane surface. For scale detection, the membrane monitor would typically receive a sidestream from a tail element of the RO plant where the retentate concentration is highest. The system can be adjusted such that the level of solution supersaturation at the membrane monitor’s membrane surface is at or higher than that for the last RO membrane module, thereby ensuring that mineral scale would be detected first on the monitor’s membrane surface. An illustration is shown in Figure 15 of early detection of gypsum crystals on the membrane surface prior to the detection of measurable flux decline. This type of scale detection in RO/NF processes can potentially be adapted to ED/EDR processes and can provide an important monitoring capability to enable safe membrane-desalting operations that adaptively vary waterrecovery levels close to temporally varying membrane-scaling threshold levels (resulting from water-quality variation). Extending membrane monitoring for online-biofouling detection is also a challenge that merits pursuit in order to enable the design of effective RO/NF and ED/EDR operational strategies.
4.04.3.4.4 Enhancing water recovery At inland locations, it is desirable to operate brackish-water desalting at high water recovery in order to reduce the volume of generated residual concentrate for locally feasible options of concentrate disposal to become cost effective. To achieve high levels of water recovery, coupling of RO and EDR processes in a series configuration have proven to be effective in some industrial applications (Reahl, 1990, 2006; e.g., Figure 16). In such cases, the use of EDR to desalt RO concentrate is particularly effective when silica is the primary mineral scalant of concern since EDR is not constrained by uncharged species. RO/NF, ED/EDR, and their integrated processes, however, still require scale-mitigation methods; the traditional scale-mitigation methods only retard the onset of mineral scaling (antiscalants, polarity reversal, etc.), but do not remove scale precursors, and thus are constrained to a threshold recovery limit. To overcome this limitation, the integration of intermediate concentrate demineralization (ICD) in a two-step membrane-desalting operation is a promising strategy (e.g., Figure 17). The function of ICD is to remove mineral-scale precursors and thus reduce the membranescaling potential of the concentrate from a primary-membrane desalting step, allowing a secondary RO desalting step to enhance product-water recovery.
(a)
Relative permeate flux (F/F0 )
1 mm
(b)
1 A 0.8
B C D
0.6
0.4 0
5
10
15
20
25
30
35
Time (h) (c)
(d)
Figure 15 Optical images of gypsum scale (a–d) and flux decline (bottom) for the corresponding RO scaling test for which initial gypsum saturation index (GSI) at the membrane (LFC-1) surface was 2.09. The images, a–d were taken at 0, 5, 20, and 30 hs. Adapted from Uchymiak M, Bartman AR, Daltrophe N, et al. (2009) Brackish water reverse osmosis (BWRO) operation in feed flow reversal mode using an ex situ scale observation detector (EXSOD). Journal of Membrane Science 341: 60–66.
MgO
Clarifier (for SiO2reduction)
EDR diluate
RO
−
UF
+
EDR Concentrate waste
Feed Product water
Figure 16 Coupling of RO, EDR, and chemical precipitation for high-recovery desalting. Adapted from Reahl E (2006) Half A Century of Desalination with Electrodialysis, Technical Paper TP1038EN. GE Water and Process Technologies.
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Brine Secondary RO desalting (SRO)
Intermediate concentrate demineralization (ICD):
Antiscalant Soda ash MF
Primary RO desalting (PRO) Solids (CaCO3)
Antiscalant
Feed
PRO concentrate Product water
MF
Figure 17 Schematic of a brine-treatment process for RO concentrate volume reduction from a primary RO desalting process. Brine treatment process consists of intermediate concentrate demineralization (ICD) via chemical precipitation and microfiltration (MF), followed by secondary RO desalting. Adapted from Zhu A, Rahardianto A, Christofides PD, and Cohen Y (2010) Reverse osmosis desalination with high permeability membranes – cost optimization and research needs. Desalination and Water Treatment 15: 256–266.
The two methods of ICD via chemical precipitation and seeded precipitation are discussed in the following. ICD via chemical precipitation. A two-step RO system with ICD (Figure 17) via chemical precipitation of calcium carbonate as a scale-precursor-removal step has been evaluated (Rahardianto et al., 2007) and pilot-tested (Gabelich et al., 2007) for the desalination of Colorado River water (B700– 1000 mg l1 TDS). The precipitation process is analogous to the classical lime-soda or caustic softening processes (i.e., precipitation softening) (AWWA, 1999). Alkaline dosing (e.g., pH adjustment with NaOH, lime, or soda ash) of the primary RO concentrate in a precipitation reactor (e.g., solid contact reactor) induces precipitation of primarily calcium carbonate, depleting the concentration of calcium in the aqueous phase, thus reducing the RO concentrate saturation index with respect to calcium-bearing mineral scalants (e.g., calcium carbonate and gypsum) to well below saturation. As an alternative to alkaline dosing, CO2 stripping has also been shown to be effective for inducing CaCO3 precipitation in brackish waters (Lisitsin et al., 2008). An added benefit of ICD via precipitation softening is the potential co-precipitation of other mineral-scale/fouling precursors, such as of Ba2þ, Sr2þ, silica, and also adsorptive removal of natural organic matter. The permeate production with the above approach demonstrated overall water recovery of up to 95%, provided that good pH control was maintained in the precipitation reactor along with antiscalant makeup to control silica scaling in the secondary RO step. In cases in which silica is the mineral scalant that limits water recovery, the use of EDR instead of RO for the secondary RO desalting step may be beneficial. To improve the efficiency of solid–liquid separation in precipitation softening, fluidized bed reactors and integrated precipitation–filtration systems have been proposed (Graveland et al., 1983; Oren et al., 2001; Sluys et al., 1996). ICD via seeded precipitation. Primary RO (PRO) desalting can potentially generate concentrate that is in meta-stable supersaturation with respect to various mineral salts. Such
behavior is attributed to slow crystal-nucleation kinetics and/ or the application of antiscalant treatment (which retards precipitation). The generated supersaturation level in the PRO concentrate can be utilized to drive precipitation by crystal growth in the intermediate concentrate-demineralization stage, initiated by crystal seeding. In this case, seed-crystal surfaces provide crystal growth sites for precipitation reactions to occur, which would lead to concentrate de-supersaturation (Rautenbach and Linn, 1996; Bremere et al., 1999; Yang et al., 2008). The de-supersaturated PRO concentrate can then be further desalted in a secondary RO step. Thus, unlike demineralization by chemical precipitation, de-supersaturation by seeded precipitation avoids the use of dissolved chemical reagents to generate the precipitation driving force. Antiscalant carryover from the PRO, however, may retard seeded precipitation and significantly reduce the rate of de-supersaturation. Methods have been proposed for deactivating antiscalants prior to crystal seeding, such as the use of coagulants (Yang et al., 2007, 2008) and low-dose lime pretreatment (Rahardianto et al., 2010). The rate of de-supersaturation can also be enhanced by integrating a membrane-concentrator unit to increase solution supersaturation above preexisting levels in the PRO concentrate. In this case, a specially designed membrane-concentrator unit is required in order to avoid membrane fouling and/or scaling. For example, Rautenbach and Linn (1996) described the use of disk-tube NF modules on concentrate primary RO concentrate generated from desalting of dumpsite leachate (Figure 18). High overall RO recovery (495%) was achieved, but periodic NF feed-channel flushing (every 30 s) and alkali cleaning (every 250–300 h) were required as preventive actions against fouling and/or scaling.
4.04.3.4.5 Concentrate disposal An extensive survey in the USA revealed that brackish-water desalting plants employ the following methods for concentrate disposal, in the order of decreasing frequency of use: surface
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Seawater Use and Desalination Technology Crystallizer RO NF Feed water
RO
RO
Purified water
RO
Concentrate reject
Figure 18 Process schematic for high-recovery desalting of dumpsite leachate. Adapted from Rautenbach R and Linn T (1996) High-pressure reverse osmosis and nanofiltration, a ‘‘zero discharge’’ process combination for the treatment of waste water with severe fouling/scaling potential. Desalination 105: 63–70.
water discharge, discharge to sewers, deep-well injection, evaporation ponds, spray irrigation, and zero liquid discharge (ZLD) (Mickley, 2006). Concentrate discharge to surface water and sewers accounts for about B70% of plants surveyed, while deep-well injection accounts for B20% (Mickley, 2006). Most of the plants (B80%) do not treat the concentrate waste before disposal, while the rest apply minimal treatment such as aeration, pH adjustment, degassification, air stripping, or defoaming. Concentrate discharge into surface water/sewers are typically available only in the case of desalting plants that are small or located near coastal areas (e.g., ocean outfall), while deep-well injection systems are typically used in larger inland desalting plants and in remote locations. Nevertheless, concentrate conveyance over long distances for ocean outfall, while costly, has also been practiced in inland desalting. To avoid precipitation of mineral salts (due to supersaturated concentrate) in long-distance concentrate-disposal pipelines, antiscalant treatment, concentrate-stream isolation from the atmosphere (to prevent CO2 release to air), and avoidance of particulate contamination are important (Semiat et al., 2004). Evaporation ponds and spray irrigation have also been used for concentrate disposal, but are land intensive and thus are used less frequently. A system for enhancement of evaporation pond performance known as the wind-aided intensified evaporation (WAIV) process, involves periodic circulation of pond brine over wettable surfaces, designed to increase the effective evaporative surface area (Gilron et al., 2003). This results in enhanced evaporation, which also depends on wind speed and direction in addition to relative humidity. It has been reported that with the above approach, evaporative capacity per area footprint can be increased 450% in a typical Middle East dry climate. Concentrate disposal via ZLD processes typically employ industrial evaporation methods (e.g., single-/multiple-effect evaporators, VC evaporators, evaporative crystallizers, and spray dryers) that can be readily deployed but at high energy and capital costs.
4.04.3.4.6 Specific contaminant removal Brackish-water sources, especially those impacted by human activities, often contain specific contaminants that must be
removed in brackish-water desalting. Brackish water impacted by agricultural activities, for example, may contain elevated concentrations of boron, nitrates, pesticides, and selenium, while those impacted by mining operations may contain elevated concentrations of arsenic and other heavy metals. Most of these contaminants are readily removed in desalting operations by RO/NF or ED/EDR. In some cases, the generated concentrate may require treatment before disposal, but it depends on the concentrate-disposal methods and related permitting/environmental issues. In addition, polishing the permeate (or diluate) water may be required in order to comply with strict water-quality regulations. For example, ion exchange or EDI are common for removal of specific contaminants. In the case of boron, RO/NF and ED/EDR processes are typically ineffective when operated at near-neutral pH as boron primarily exists as uncharged boric acid. The deprotonated form of boron (borate) is highly rejected in RO/NF processes (490–95%), but this requires operation at high pH (pH 410) that may increase the tendency for membrane scaling by carbonate minerals. To allow operation at high pH for enhanced boron removal, a multi-pass RO/NF configuration (e.g., Figure 9) can be utilized in which the permeate from a primary RO/NF is desalted at an elevated pH level in a secondary RO process. It is noted, however, that EDI can also be effective for product-water polishing for boron removal (Wen et al., 2005).
4.04.3.4.7 Cost of brackish-water desalination Major cost items in brackish-water desalination typically include costs of electrical energy, chemical additives (antiscalant and acid for scale mitigation), membrane replacement, concentrate treatment for recovery enhancement, brine disposal, capital depreciation, and financial interests. It is not feasible to arrive at generalization of the cost structure of brackish-water desalination because capital, operating, and financial costs are highly site specific. Nevertheless, one should expect that the energy cost of brackish-water desalting to be much lower than seawater desalting due to lower salinity of brackish water. Therefore, the cost of chemical additives (e.g., antiscalants) can become a significant portion of the total operating cost. Concentrate management (treatment and disposal), however,
Seawater Use and Desalination Technology (a) Single-step RO desalting 58% Water recovery operating cost: US$0.71 m−3
99
(b) Multi-step RO desalting w/ ICD 92% Water recovery operating cost: US$0.45 m−3
Electrical energy
RO membrane Electrical energy
Antiscalant
MF operation Brine disposal
RO membrane
ICD chemicals Antiscalant Brine disposal
MF operation
Figure 19 Operating cost estimates for (a) single-step RO desalting at 58% water recovery and (b) multi-step RO desalting with ICD at 92% water recovery (see Figure 17). Cost estimates are for desalination of San Joaquin Valley agricultural drainage water (8500 mg l1 TDS), assuming deep-well injection for concentrate disposal. Data from Zhu A, Rahardianto A, Christofides PD, and Cohen Y (2010) Reverse osmosis desalination with high permeability membranes – cost optimization and research needs. Desalination and Water Treatment 15: 256–266.
tends to be the primary cost component in brackish-water desalting, especially at inland locations where disposal options are limited. An illustration of the significant impact of brine management (treatment and disposal) on operating costs is illustrated in Figure 19 for the case of agricultural-drainage-water desalting (8500 mg l1 TDS; Zhu et al., 2010). In this example, the effectiveness of antiscalants against gypsum scaling is expected to limit RO water recovery to 58%, leading to a large volume generation of RO concentrate. The operating cost is estimated to be high at US$0.71 m3 product; a large portion of this cost is for concentrate management (81%), given an estimated brine-disposal cost of US$0.8 m3 brine (e.g., via deep-well injection, western San Joaquin Valley; Johnston et al., 1997). Recovery enhancement to 92% recovery via ICD and secondary RO can reduce the total operating cost to BUS$0.45 m3 product. Although the contribution of concentrate-management costs (i.e., ICD chemicals for concentrate treatment and brine disposal via deep-well injection) to the total operating cost is lowered to 64%, this concentraterelated cost remains significant. Total costs of present brackish-water desalting plants vary significantly, representative of the high variability of local resources (e.g., feedwater quality and availability of concentrate-disposal sites). For example, a survey of six brackish groundwater RO desalting plants built during the past decade in Texas (4500–104 000 m3 d1 or 1.2–27.5 MGD production capacity) revealed total production costs between US$0.33 and 0.69 m3 product, with approximately 60–70% attributed to operation and maintenance (O&M) costs and the remainder being financial-interest costs (Arroyo and Shirazi, 2009). Of particular interest in the plants surveyed is the K. B. Hutchison plant (capacity of 104 000 m3 d1; total permeate production cost of US$0.65 m3), in which almost 30% of the capital expenditure was associated with the deep-well injection system for brine disposal (Committee on Advancing Desalination Technology, 2008). Finally, it is noted that
comparative analysis has shown that ED/EDR plants may be more cost effective than RO/NF plants when feedwater salinity is less than about 3700 mg l1 TDS (US Bureau of Reclamation, 2003).
4.04.3.5 Future Developments Significant improvements in the performance and cost effectiveness of pressure- and electrochemical-driven desalting technologies over the past two decades have enabled brackishwater desalting applications to meet anticipated water demands, while moving forward toward water sustainability. To sustain this growth, further technology developments are needed for optimal use of brackish-water resources, considering the spatial distribution and temporal variability of source-water availability and characteristics, energy sources, and concentrate-management options. Concentrate management, in particular, continues to be the primary cost impediment in many brackish-water desalting applications, especially at inland locations. Improvements are needed in the treatment and utilization of desalination concentrate (i.e., brine). Lesschemical intensive methods for scale-precursor removal, for example, would reduce costs associated with water-recovery enhancement in brackish-water desalting. There is also potential for chemical recovery from brine streams, such as acid/ base chemicals and mineral commodities. Significant opportunities exist for the development of smart membrane systems that can operate autonomously and remotely for optimal desalting of brackish-water resources. Such systems could be suitably located at sites that would match local water demands with availability of local brackishwater resources, as well as minimize costs associated with concentrate disposal and chemical and energy use. For example, smart membrane systems that can operate adaptively near fluctuating membrane fouling and scaling threshold levels, due to feedwater-quality variations, could enable one to maximize water recovery in real time and thus reduce
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concentrate-disposal costs. This would require further development of software and devices that can provide high-fidelity sensing of membrane fouling and scaling and timely automatic actuation of preventive actions. Integration of fouling and scaling mitigation methods that are less chemical intensive would also be highly beneficial for smart water systems. Development of high-permeability membranes for RO/NF desalting is not expected to lead to significant reduction in the cost of brackish-water RO desalination, especially since recent RO/NF membranes already allow for cross-flow RO/NF operation near the thermodynamic restriction (Zhu et al., 2008). Further developments of membranes and membrane processes that are fouling/scaling tolerant are needed, as well as membranes and processes that are highly selective for targeted contaminants.
4.04.4 Desalination of Wastewater for Reuse Wastewater reclamation is increasingly practiced around the world to meet industry, domestic, and agricultural needs. When the salinity of the reclaimed wastewater is too high for its intended use, a desalination stage may be required to produce water of appropriate quality. The cost of desalination processes means that desalination of reclaimed wastewater for agricultural purposes is rarely practiced, and the main uses of desalinated reclaimed water are for industrial and domestic uses. RO desalination is the dominant desalination technique used, as it requires far less energy (0.7 kW h1 m3; Leslie and Myraed, 2009) than thermal processes for this application. Therefore, the discussion in this section focuses on membrane-based desalting processes.
4.04.4.1 Water Quality Desalination of reclaimed wastewater and brackish groundwater are similar in some respects, as both have similar TDS concentrations, particularly when compared to seawater. However, the higher concentration of organic compounds and the large variability in their nature means that the operational issues are significantly different, with a high tendency for biofouling existing in reclaimed wastewater systems. Additionally, the higher concentration of phosphates in reclaimed wastewaters may also lead to scaling by calcium phosphates rather than calcium carbonate and gypsum, as is the case for brackish water systems. The quality of water feed to a wastewater reclamation plant depends upon the level of treatment in the preceding wastewater treatment plant, and only secondary or tertiary wastewater effluents would be considered suitable for processing in a reclamation plant. Secondary treated effluent refers to wastewater that has been treated by sedimentation followed by a biological process such as treatment in an activated sludge plant. Tertiary treatment refers to secondary treatment followed by a filtration step, such as media filtration, so that the turbidity and TOC concentrations are generally lower, and if coagulation with metal salts is used, then the phosphate concentration will also be reduced (Henriksen, 1963). Table 5 shows typical water qualities for secondary and tertiary wastewater treatment plant effluents.
Table 5
Typical wastewater treatment effluentsa Secondary treatment
pH Turbidity (NTU) Total suspended solids (mg l1) TOC (mg l1) TDSa Nitrate (mg l1) NH3 – N (mg l1) Phosphate (mg l1)
Tertiary treatment
6.5–8.2 5–25 25–35
6.5–8.2 o3 5–10
10–20 500–1500 20–30 3–10 3–8
4–10 500–1500 a 5–20 a 0.4–5 a 0.5–5
a
Lower limits for biological nutrient removal plants: Note: TDS concentrations lower than 500 mg l would not usually be considered for desalination. From Montgomery Watson Harza (Adham S, Burbano A, Chiu K, and Kumar M) (2006) Development of a reverse osmosis/nanofiltration (RO/NF) knowledge base. Report of the California Energy Commission. Pasadena, CA: California Energy Commission; and Water Corporation (1999) Personal communication, Water quality data from wastewater treatment plants.
4.04.4.2 Pretreatment Water Factory 21 in Orange County, California, USA, was an early successful example of wastewater reclamation. It began operation in 1976 and operated for approximately 30 years. Its pretreatment system initially consisted of chemical clarification and settling, ammonia stripping, re-carbonation, multimedia filtration, and activated carbon adsorption prior to being fed into RO membranes. However, extended operation of this plant revealed that carbon fines blocked the entrance to the spiral wound RO elements, while the ammonia stripper cooled the water leading to lower productivity through the RO membranes. Additionally, residual ammonia is converted to chloramines and assists with biofouling control. The pretreatment process before RO was then altered to lime clarification, re-carbonation, chlorination, and media filtration (Asarno, 1998). Water Factory 21 has now been replaced by the Groundwater Replenishment Scheme, in which the pretreatment process is now MF (Montgomery et al., 2006). The change in pretreatment trends to the use of MF and UF systems is a result of the very high quality of water produced, improvements in the operability of MF and UF systems, and relative reductions in their price. A 2006 survey of desalination systems by Montgomery Watson Harza (2006) showed that of the 14 wastewater-reclamation plants included in the survey, only one had a conventional pretreatment stage (coagulant, sedimentation, and media filtration), whereas nine had only MF or UF, two had both conventional and membrane pretreatment; and the remaining two had only cartridge filtration. The high number of plants with only membrane pretreatment shows that this is the recent trend, and it is increasing in popularity. Water quality from MF or UF systems is usually o0.2 NTU and often consistently lower than 0.1 NTU, and TOC concentrations are reduced from 5–15 mg l1 to 3–14 mg l1. Greater TOC reductions can be achieved if coagulant is used, with ferric chloride or ferric sulfate more commonly used than aluminum-based salts. It is noted that high TOC content may suggest high concentrations of biopolymers, such as
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Cl2 Lime
To RO Wastewater treatment
Lime sludge Lime clarification
Air re-carbonation
Dual-media filtration
Backwash To RO Cl2 Coagulant, pH Cl2, backwash MF/UF
To RO Air
MBR
Figure 20 Various options for pretreatment stages before wastewater-reclamation desalination.
polysaccharides and proteins (Jarusutthirak et al., 2002), which are strong foulants of MF and UF systems and are easily biodegradable, thereby promoting biofilm growth. Flux through the MF/UF membranes is typically 30–40 l1 m2 h1. Either pressure driven or submerged MF or UF membranes may be used in the pretreatment stage (see Figure 20). The decision to choose either pressurized or submerged MF or UF is based on cost and operability. Submerged systems can be easily fitted into a tank, and so are cheaper for large units or when retro-fitting to an existing tank. However, pressurized systems can be packaged in the manufacturer’s plant and have greater capacity to increase throughput during peak-flow periods. Hypochlorite is often added before MF or UF, as it reacts with ammonia in the wastewater to form chloramines that limit biofilm growth on the membranes. If the concentration of ammonia in the feedwater is low, ammonia may be added before the addition of hypochlorite although this is seldom a necessity. Chloramines have been the preferred disinfectant for controlling biofilm growths on MF, UF, and RO membranes because of their lower oxidizing potential relative to chlorine and hypochlorite. Reasonably low but effective chloramine doses (1–4 mg l1) may be applied safely to suppress biofilm growth while minimizing oxidative degradation of membranes. As chloramine tolerance of membranes may vary due to the catalytic effects at high temperature, low pH, or presence of transition metals, optimal chloramine dose should be determined for the specific source water of interest, membrane type, and operating conditions. New membrane materials such as polyvinylidene fluoride (PVDF) and polyethersulfone (PES) systems are chlorine tolerant, and chemically enhanced backwashing (CEB) strategies are in use currently. Chlorine in concentrations between 25
and 100 mg l1 is added to the backwash water and the membranes are soaked in the chlorine solution for 5–10 min. This would typically occur 1–2 times a day, and decreases the requirement for chemical cleaning of the membranes. Chemical cleaning is often required as contaminants can irreversibly foul the membranes with extended filtration times. Chemical cleaning is usually performed with caustic surfactant solutions to remove organic contaminants, while ethylenediaminetetraacetic acid (EDTA) and/or citric acid may be added to remove inorganic foulants. Chemical cleaning would typically take place every 6–8 weeks without chemically enhanced backflush (CEB) and after more than 6 months with CEB. Membrane bioreactors (MBRs) are also being considered for use in pretreatment before RO membranes for wastewaterreclamation plants. Qin et al., (2006) have demonstrated the use of an MBR RO system in place of a conventional wastewater-treatment plant followed by MF or UF pretreatment. They were able to demonstrate that the MBR pretreatment produced water with lower TOC concentrations than the conventional wastewater-treatment plant–MF system, producing TOC in the range of 4.9–5.1 mg l1 compared with 6.8–6.9 mg l1 for the conventional-MF pretreatment. This improved water quality, although only slightly, enabled the RO system to operate at higher fluxes of 22 l m2 h1 compared to 17 l m2 h1 when fed with conventional-MF treated water. The final RO permeate also improved, with TOC values of 24–33 ppb rather than 33–53 ppb. Again, while this was only a slight improvement in water quality, it was significant for the semiconductor industry using the reclaimed water as their requirement was for ultrapure water. Additionally, if being used for potable purposes, the lower TOC value indicates reduced organic components present in the wastewater,
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lessening the potential health risks associated with trace organic compounds, such as endocrine disruptors, in the final product water. Acid and antiscalant addition occurs after the MF/UF or MBR pretreatment stage, to reduce the likelihood of scaling in the RO stage. Addition of acid to lower the pH to around 6.5 is common (Zach-Maor et al., 2008), and the antiscalant type depends upon the likely mineral scalant.
4.04.4.3 RO Processes RO desalting of wastewater can usually achieve 70–90% water recovery using a multi-stage process configuration (e.g., Figure 11). Recovery levels of 70–75% are typically achieved using a two-stage RO desalting process, while recovery levels between 75% and 88% typically require a three-stage RO desalting process. Brackish-water membranes are used for wastewater treatment, as wastewater has salinity levels similar to those of brackish water. The operating pressure varies between 7 and 25 bar, with greater pressures required to achieve high recoveries because of the increase in the final salt concentration. Flux through the membranes varies considerably, with reported values between 14 and 27 l m2 h1 (Montgomery et al., 2006). The recommended design flux is usually quoted as between 15 and 20 l m2 h1. Fouling in wastewater-reclamation plants can vary from that in brackish water, as the higher organic carbon content of the feed increases the potential for biofilm growth, specific organic compounds in the feedwater may foul the membrane, and higher phosphate concentrations may lead to calcium phosphates precipitating rather than calcium carbonate or gypsum. The addition of chloroamine before the MF or UF process enables the disinfectant to control biofilm growth in both the pretreatment and desalination stages. This strategy is generally successful, with cleaning of membranes required between 3 and 6 months. Other approaches to biofilm control, such as chlorine dioxide dosing (Wise et al., 2004) and UV disinfection prior to the RO unit (Lo´pez-Ramı´rez et al., 2003) have been suggested but further studies are required before they can be implemented. Treatment of wastewater from domestic sources does not usually pose any significant problems, but industrial waters contain a wide range of compounds not generally present in domestic sewage and perhaps only present in the particular wastewater catchment. This has caused problems for a number of commercial systems, such as the Wollongong Recycled Water Plant, NSW, Australia (Borse et al., 2009). The Wollongong wastewater contained specific organic compounds, tert butyl phenol, 2-methylthio-benzothiazole, trichlorphenol, and trichlorocresol, that have the potential to foul membranes even when present in small doses and which may not be easily detected by membrane autopsies. Generally, these fouling issues can only be managed by operating at higher pressures or lower fluxes and by developing specific cleaning regimes. Additionally, these specific fouling issues are difficult to detect because of the low concentrations of the specific foulants, the difficulty in detecting them via membrane autopsies, and on occasions due to the intermittent presence of the contaminants. Bench- and pilot-scale field studies are therefore
recommended when treating wastewaters from industrial sources because this is the most reliable approach for detecting such issues. Mineral scaling of membranes is an issue, particularly when operating at high water recoveries. The control of mineral scaling is similar to that for brackish water systems, with acid addition prior to RO treatment to increase the solubility of inorganic mineral scalant, and the use of antiscalants. Calcium phosphate scaling arising from the high concentrations of phosphate in the feedwater, however, is more difficult to control with antiscalants (Greenberg et al., 2005); presently, feedwater pH adjustment appears to be the best available option for controlling calcium phosphate scaling. Alternatively, control of phosphate via coagulant addition in the pretreatment stream is also an option, but higher doses of coagulants are required than that for turbidity removal alone (Henriksen, 1963).
4.04.4.4 Final Water Quality Permeate TOC concentrations are low, around 50 ppb, and TDS concentrations are approximately 20–30 mg l1. Water of this quality is corrosive, and stabilization of permeate by mixing with nondesalination effluent is frequently practiced along with the addition of a residual disinfectant such as chlorine. If the water is to be used for industrial purposes, however, it is sometimes delivered in the high-purity form to be used in cooling towers and boilers. In Singapore, the semiconductor industry requires ultra-high-purity water; therefore, stabilization of recycled water is not in practice. The Luggage Point recycled water facility in Brisbane, Queensland, similarly does not stabilize the water delivered to a petroleum refinery where it is used in cooling towers and a demineralized water plant.
4.04.4.5 Concentrate Disposal The presence of nutrients and synthetic and natural organic compounds in wastewater brine complicates its disposal in comparison to brackish water brines (CH2M HILL, 2009). The treatment of wastewater brines in evaporation ponds is usually unaffected by these contaminants, but discharge to waterways or coastal marine environments may require additional treatment. For instance, the Bundamaba advanced watertreatment plant in Brisbane, Australia, reclaims wastewater for industrial use and indirect potable-water reuse and releases the brine concentrate into the Brisbane River that feeds into Morton Bay. The discharge of nutrients in this region is regulated because of the sensitive ecology, and treatment of the brine concentrate for nutrient removal is required. The pretreatment sludge is thickened with the aid of polymers, centrifuged, and the centrate mixed with the brine concentrate. The combined concentrate is treated via a fixed-film nitrification process following which it enters a denitrification stage that requires methanol addition to assist the process (Davies, 2009). The Bundamba brine-concentrate treatment process uses a conventional nitrogen-removal strategy, but there have also been studies investigating the use of wetlands to remove nutrients and metals from brine concentrate (Kepke et al., 2009).
Seawater Use and Desalination Technology
Sites at Luggage Point, Queensland, Australia and Oxnard, California, USA have pilot-tested the treatment of wastewaterreclamation brine by surface flow and vertical-upflow wetland cells. Initial results demonstrated 70–80% reductions in nitrate, up to 50% reduction in selected metals (B, Cu, Cr, and Mo) under some conditions, and increases in concentrations of other metals (As, Al, Cd, and Mg) and TDS. The increase in load of some metals was due to leaching from the soil, while the TDS increase was the result of plant transpiration reducing the water volume. Additional trials are required before the effectiveness of this approach is fully understood.
4.04.5 Alternative Technologies The large increase in demand for desalination technologies and the relatively higher energy requirements compared to other water-treatment processes have led to the intense search for alternate desalination technologies. The most notable of these processes are discussed in this section.
4.04.5.1 Membrane Distillation Membrane distillation (MD) is a desalination process which brings membranes into thermally based processes such as MED. Therefore, the theoretical approaches to assess MD performance, just as in distillation, stem from the enthalpy of evaporation, and the potential advantages lie in the functions of the membrane, which include
• • •
•
Containment of the evaporating surface (vapor–liquid interface) thus allowing more control of the system in certain applications. Efficient packing of controlled, regular membrane geometry for smaller process footprint. Novel aspects of fluid handling to allow for more functional setups, such as better heat efficiencies, and further footprint reduction. For example, in direct-contact MD (DCMD) mode, higher fluxes are observed due to reduced resistance allowed by the intimate contact of the cooling fluid on the permeate side of the membrane. Moreover, in the air-gap MD (AGMD) mode, a single compact module performs heat recovery simultaneously during desalination, providing improved heat recovery when compared with direct-contact MD. Cheaper materials for constructing the membrane modules (i.e., polymer-based materials) when compared to systems consisting of corrosion-resistant metals.
However, there are also some disadvantages of MD when compared to traditional thermal desalination processes such as MSF, MED, and VC, such as
• • •
Lower heat-transfer coefficients and mass-transfer coefficients compared to traditional thermal desalination processes such as MED, MSF, or VC. The heat efficiency as measured by the GOR is still lower than that of traditional thermal processes. For low temperature VC or MED (70 1C), however, aluminum transfer surfaces are used and corrosion is not an issue, thereby reducing any advantage in the use of construction materials made of polymer in MD.
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So, while there are some potential advantages in the application of MD, it has struggled to find application in the water industry because of the efficiency of traditional thermal processes and RO processes. The recent renewed interest in MD research will need to find specific applications where the advantages of MD can be realized. Applications such as brine treatment, the utilization of available low-grade heat (o70 1C), which traditional thermal processes find difficult to use, or applications requiring a small footprint process appear to be best suited to MD. Such processes have not been economic to consider previously, but with greater pressure on water resources and increasing costs of brine disposal in some locations (lined evaporation ponds), this economic equation may change.
4.04.5.1.1 Brief history The first patents for MD were filed in the 1960s (Bodell, 1963; Weyl, 1967), but the process never progressed into commercial utilization due to the high cost of the membranes and the process at that time. Moreover, RO was being developed, and when weighed up at a time of low-cost electricity, RO was by far superior in producing water at lower-cost per unit volume. Recently, with rising energy costs and awareness of greenhouse-gas emissions associated with electricity production needed for RO, MD has seen a recent resurgence (Curcio and Drioli, 2005). This is coupled with an increase of interest in membrane systems, as there are now numerous industries manufacturing membranes in commercial quantities at competitive prices. It is important to note here that these membrane advancements have been motivated for MF and application in advanced clothing applications (i.e., Gortex-type membranes), and not MD specifically. The range of hydrophobic membranes including polypropylene, PVDF, and polytetrafluoroethylene that have recently emerged make excellent candidates for MD. These improved membranes have led to increased flux and lower fouling, but limitations associated with energy efficiency have not been assisted by these improvements.
4.04.5.1.2 Membrane distillation configuration Generally, an MD system is made up of a hydrophobic porous membrane over which the saline water is passed. As the membrane is hydrophobic, the liquid feed cannot penetrate when the pressure is lower than the liquid entry pressure (LEP) of the membrane. When estimating LEP in MD, it is important to take the largest pore because any breakthrough of liquid has significant consequence to salt rejection. For example, on a 0.5-mm pore-size rated membrane, LEP calculated would be 270 kPa, but in practice, is at least 130 kPa. Such pressures are suitable for most MD setups, but higher LEPs can be achieved with UF membranes in the MPa region as their pore sizes are around 20 nm. These pressures are essential for the design of the system whereby the pressure of the feed must not exceed the LEP. There must also be an allowance factored into the design to allow increasing backpressure as a result of module fouling since feed flow is typically maintained constant for effective MD performance. As a process, there are four MD configurations: DCMD, AGMD, sweep-gas (SGMD), and vacuum (VMD), each of which is shown in Figure 21.
Seawater Use and Desalination Technology
Permeate (a)
(b)
Pore Permeate-vapor
Permeate-vapor (c)
Sweeping gas
Feed
Feed Pore
Pore
Pore
Membrane
Membrane
Vacuum
Cooling plate
Cool feed
Membrane
Permeate
Feed
Membrane
Hot feed
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(d)
Figure 21 Four membrane distillation (MD) setups commonly applied, direct-contact MD (DCMD) (a), air-gap MD (AGMD) (b), vacuum MD (VMD) (c), and sweep-gas MD (SGMD) (d). From Zhang J, Duke M, Ostarcevic E, Dow N, Gray S, and Li J-D (2009) Performance of new generation membrane distillation membranes. Water Science and Technology: Water Supply 9(5): 501–508.
For water treatment, DCMD and AGMD are the most widely researched MD processes for desalination, while SGMD and VMD are generally applied for volatile organic compound (VOC) removal (Curcio and Drioli, 2005). With regard to desalination, AGMD shows a lot of promise due to its potential for high energy recovery resulting from its inherent ability to allow for simultaneous evaporation and condensation (heat recovery) within the module, as shown in Figure 22, which is similar to the MED concept.
4.04.5.1.3 The Memstill project The Memstill project is a major MD development project using the AGMD process for its potential as a lower-energy desalination alternative to RO (Dotremont et al., 2009). It has been claimed that because of its thermal driving force and large surface areas, low-grade heat from waste sources makes it more cost efficient than conventional pressure-driven (i.e., electric) RO. However, operating at MD’s optimal energy efficiency comes at the expense of low flux and high surface areas, and use of waste heat requires a heat source of compatible energy value. Table 6 shows the latest developments of the Memstill process. The first Memstill pilot plant with M28 module began operation at the Senoko Incineration Plant, Singapore, with raw seawater fed from the Straits of Johor. The second and third Memstill pilot studies with M32 and M33 modules were conducted in the Netherlands at the E.ON Benelux power plant (Dotremont et al., 2009). Both were fed with seawater from the Port of Rotterdam. Progress in demonstrating improved efficiency for Memstill has seen heat efficiencies come down to about one-third of the M28 module design. While the processes in MD are similar to the thermal processes presented earlier, the use of membranes instead of metallic heat-transfer surfaces increases the resistance to heat and mass transfer, and low flux across the membranes results. This leads to the need for large surface areas, and high packing densities, as has been achieved for other membrane-based
Figure 22 Concept schematic of the Memstill process. From Dotremont C, Kregersman B, Sih R, Lai KC, Koh K, and Seah H (2009) Seawater desalination with Memstill Technology – a sustainable solution for the industry. Paper presented to IWA Membrane Technology Conference, Paper 190. Beijing, China, 1–3 September.
processes, implying that it could result in a compact device with lower footprint. However, the energy needs are basically similar to those of the thermal processes. Special designs may allow process operation similar to MSF or MED. The thermal processes are fed by low-value heat sources. However, the main question is whether it is possible to obtain high GOR as obtained with the existing thermal processes, particularly
Seawater Use and Desalination Technology Table 6
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Summary of previous Memstill pilot installations
Module used
M28
M32
M33
Location of testing Duration of testing Absolute flux (l m2 h1) Internal heat recovery (%) Heat consumption (kW h1 m3)
Singapore, Straits of Johor March 2006 to June 2007 0.25 30 278–556
The Netherlands, Port of Rotterdam October 2006 to January 2007 2.5 50 111
The Netherlands, Port of Rotterdam April 2008 to October 2008 3 90 97–111
From Dotremont C, Kregersman B, Sih R, Lai KC, Koh K, and Seah H (2009) Seawater desalination with Memstill Technology – a sustainable solution for the industry. Paper presented to IWA Membrane Technology Conference, Paper 190. Beijing, China, 1–3 September.
given the lower temperatures used, which limit the temperature difference and hence the heat recovery. If cooling water is used to condense the vapor via cooling devices, it requires much higher energy than in the regular thermal desalination. This has led to much higher energy consumption (low GOR and high pumping energy) and high cost of equipment (large membranes). Capturing the heat of condensation from the vapor in these processes, however, has the potential to produce high GORs. A certain amount of heat recovery from the permeate is also possible, but the extent of heat recovery is currently much lower than that from the heat of condensation (Table 6).
4.04.5.2 Forward Osmosis 4.04.5.2.1 Background The process of osmosis involves the migration of water from a less-concentrated saline solution across a water-permeable (and salt impermeable) membrane to a more concentrated saline solution. This process is energetically favorable as the system seeks to increase entropy (i.e., mixing). As shown in Figure 23, forward osmosis (FO) works on this principle to harness water, typically from naturally saline sources, such as seawater, by drawing it through the semipermeable membrane into a very different synthetic saline solution at a higher concentration than the original feedwater. The special aspect of FO is that the high-concentration synthetic saline draw solution contains different salts that are more practically separated than the original saline source. Clearly, the energy input to drive FO is in the regeneration and recirculation of the draw solution. According to Bolto et al. (2007), FO has been applied to contaminated waters, drawing water into fluids containing electrolytes and sugars for military applications. In such applications, the draw solution is not regenerated, as the sugars that constitute the draw solution are consumed along with the clean water in the final product. Many uses of FO such as this are practical, and a good comprehensive review on FO was made in 2006 (Cath et al., 2006). However, the use of FO as a continuous desalination process in which the draw solution is recycled has several obstacles even now, with the major process issues being the need to develop compact, stable membrane systems specifically for FO, and to improve the ease of regeneration of the draw solution which must be nontoxic.
4.04.5.2.2 Recent developments RO has also been used to remove the water from a NaCl draw solution as shown in Figure 24. RO brine concentrates from a
Concentrated draw solution recycle Saline feed water
FO membrane unit
Draw-solute separation
Potable water
Brine Diluted draw solution Figure 23 Forward osmosis (FO) concept processes. From McCutcheon JR, McGinnis RL, and Elimelech M (2005) A novel ammonia–carbon dioxide forward (direct) osmosis desalination process. Desalination 174(1): 1–11.
groundwater source up to 17 500 mg l1 were further desalinated by the FO/RO process. The system’s recovery reached 90% but performance was limited by scaling salts. FO may not emerge as an energy-saving process, but due to its operating conditions, may be more suitable in environments which normally foul membranes in conventional RO. This is mostly due to its ability to operate in the feed solution at a much lower pressure and so, for example, in highly turbid water, particles are not forced into the membrane pores leading to blockage (Bolto et al., 2007). The assumption that low-flux FO is economical is yet to be demonstrated. Opportunities may present themselves in harnessing renewable or waste energy for desalination, much as in thermal distillation processes. The draw solution can be regenerated by heat and thus can be adapted to harness solar thermal energy or low-grade waste heat. This is the case for the ammonia and carbon dioxide mixing with the water to form ammonium bicarbonate, which can be removed by heating to around 60 1C (McCutcheon et al., 2005). However, energy is needed not only to distil the ammonium carbonate, but also to pump water from the sea, as in RO processes, and for cooling to remove the heat of adsorption of the ammonium carbonate to regenerate the draw solution. Evaporation of water vapor occurs during the distillation of the ammonium carbonate from the solution, and distillation down to very clean, ammonia-free product is
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Seawater Use and Desalination Technology
Concentrated DS
FO
RO
Feed Draw solution
T C Diluted DS
High pressure pump
Product water
Figure 24 Flow diagram of an FO and RO integrated process to purify water From Martinetti CR, Childress AE, and Cath TY (2009) High recovery of concentrated RO brines using forward osmosis and membrane distillation. Journal of Membrane Science 331(1–2): 31–39.
+ +
+ + +
+
+ +
+
Carbon aerogel
+
+ electrode __ _ _Negative _ _ _ _
_
Treated water Brackish water
+
+
+
+
+
Positive electrode Figure 25 Graphical representation of capacitive deionization (CDI) process. From Gabelich CJ, Xu P, and Cohen Y (2010) Concentrate treatment for inland desalting. In: Escobar IC and Schafer AI (eds.) Sustainable Water For Future Use – Water Recycling Versus Desalination, vol. 2, 1st edn., pp. 295–326. Amsterdam: Elsevier.
necessary. This requires considerable energy and needs to be performed under vacuum conditions, similar to evaporation desalination techniques. The capital cost of VC is high, as is the energy demand. An estimation of the energy consumption for this process has been made by Semiat et al. (2010), who calculated that it would require 13 7 3 kW h1 m3 and that the final permeate would contain 9 mg l1 ammonium. The permeate would require further processing, via a process such as ion exchange, to reduce the ammonium concentration to an acceptable level (o1 mg l1).
4.04.5.3 Capacitive Deionization Capacitive deionization (CDI) is a conceptually simple technique to remove salts from water. It works by passing the saline water over a charged electrode surface which literally causes the ions (e.g., NaCl, CaCO3, and CaSO4) to stick to the electrode. This concept is shown in Figure 25.
Once saturated, the charge must be reversed and the released ions redirected to the discharge brine stream. Clearly, more surface area available to the ions per unit volume of material is desired for economical performance. Hoang et al. (2009) carried out a recent review of the field, identifying key carbon-based materials being investigated to improve efficiency: carbon aerogels, activated carbon cloth with metal oxide nanoparticles, and carbon nanotubes. When treating an artificial brackish water of 2000 mg l1 TDS (Welgemoed and Schutte, 2005), CDI required only 0.59 kW h1 m3 to recover 70% of the water at a permeate concentration of 500 mg l1. An example of such development is reported by Zou et al. (2008), who used high-surface-area activated carbon to effectively reduce the salinity of water in laboratory trials, and also found that modification of titania nanoparticles increased the electrosorption efficiency. There are, however, a number of issues that CDI technology must address before it can be an economic alternative to
Seawater Use and Desalination Technology
RO and ED for the treatment of brackish waters. While sorption capacities of up to 80 mg TDS g1 of aerogel have been claimed, the capacities of carbons in actual application trials has only managed B8 mg TDS g1 of aerogel (Gabelich et al., 2001). This arises from ion adsorption being based on ionic hydrated radius, so that only pores greater than 20 nm in diameter are suitable sorption sites (Gabelich et al., 2002). Additionally, fouling of carbon aerogels by organic compounds readily occurs, which limits the adsorption capacity of the carbon electrodes (Gabelich et al., 2001, 2002; Lee et al., 2008). Commercial systems are now available and are finding application in polishing ultrapure water or treating lowsalinity brackish or wastewaters (Farmer et al., 1996). Limitations of CDI as a concentrate-minimization technology include (1) preference for removal of monovalent ions over divalent, (2) limited sorption capacity of carbon-based electrode materials, and (3) organic fouling to which the electrodes are prone when used on natural waters (Gabelich et al., 2010).
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Relevant Websites http://serc.carleton.edu EarthLabs. http://www.ide-tech.com IDE Technologies.
4.05 Abstraction of Atmospheric Humidity PA Wilderer, Technische Universitaet Muenchen, Institute for Advanced Study, Munich, Germany E Davydova and Y Saveliev, Meteo-Systems, Zug, Switzerland & 2011 Elsevier B.V. All rights reserved.
4.05.1 4.05.2 4.05.3 4.05.3.1 4.05.3.2 4.05.3.3 4.05.3.3.1 4.05.3.3.2 4.05.3.3.3 4.05.3.3.4 4.05.4 4.05.4.1 4.05.4.1.1 4.05.4.1.2 4.05.4.1.3 4.05.4.2 4.05.4.3 4.05.4.4 4.05.4.4.1 4.05.4.4.2 4.05.4.5 4.05.4.5.1 4.05.4.5.2 4.05.4.6 4.05.5 4.05.5.1 4.05.5.2 4.05.5.3 4.05.5.4 4.05.5.5 4.05.6 References
Introduction Volume of Water in the Atmosphere Fundamentals of Rainfall Generation Preliminary Remarks Water in the Atmosphere Atmospheric Processes of Precipitation Formation Thermodynamic processes Unit processes Thermodynamic approach to explaining precipitation events Electrical processes Innovative Abstraction Methods Condensation Technology General remarks Proposed technologies Research needs Fog Collection Generating Clouds with the Aid of Heat Islands Cloud Seeding Development of the technology Evaluations and recommendations Rainfall Enhancement by Cloud-Particle Charging Scientific background Development of the technology Evaluation and Recommendations Rainwater Collection, Purification, and Storage Incentives for Action Rainwater Collection Pollution and Purification of Stormwater Runoff Purification of Stormwater Runoff in Decentralized Treatment Units Large-Scale Storage of the Collected Rainwater Overarching Aspects
4.05.1 Introduction Since the middle of the nineteenth century, the number of people on our planet has been increasing at an unprecedentedly high rate. In 1959, 2.5 billion people lived on the Earth. Within 50 years, the world’s population had risen to about 6.7 billion, with an annual growth rate of about 82 million (Anonymous, 2008). Even more dramatic is the rate at which the population of urban areas has increased. In 2008, about 50% of the world’s population lived in urban areas, half of them in cities of 500 000 inhabitants or less, and the other half in cities of up to even 30 million people. By the year 2035, more than 70% of the global population is expected to live in cities. Moreover, a large proportion of the world’s population today lives in coastal areas, in strips about 100 km wide along various shorelines (Anonymous, 2006). As a consequence of such growth in the population, the demand for water and food, land and infrastructure, and
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commodities and energy has also risen, both globally and locally. Simultaneously, the emission of solid, liquid, and gaseous wastes has proliferated, polluting not only land and air, but also aquifers (groundwater), rivers, lakes, coastal areas, and oceans. These water resources, which are crucial for satisfying the water needs of humans and animals, agriculture, industry, and the planet’s entire ecosystem, are adversely exploited by this pollution. It is universally known that water is the essence of life. In contrast to all other living beings on the Earth, humans require water not only for life-enabling functions, but also for a multitude of other purposes arising from the human desire to create a lifestyle superior to that offered by nature. We use water for showering, recreational bathing, running washing machines and dishwashers, flushing toilets, washing cars, and many other activities. Farmers use water for growing crops and quenching the thirst of domesticated animals. Water is needed to produce paper, steel, and textiles, just to name a few.
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Additionally, water is used for various cooling purposes. All of this equates to a global demand for freshwater that grows overproportionally with respect to the growth of the human population on the Earth. Thus, it is fair to assume that water shortages in many parts of the world are caused not only by changes in climatic conditions and by the enormous growth of the human population, but even more by the insatiability of the human race’s demand for and consumption of water, food, and commodities. The traditional way of satisfying the human water demand is to abstract water from natural resources such as aquifers and surface water bodies (rivers and lakes). Where the water demand has exceeded the capacity of these natural resources, people have invented and implemented a variety of methods to capture rainwater or to melt snow and ice. During rainy seasons, rainwater from roofs and other sealed surfaces is collected and stored in tanks (cisterns). This method is referred to as rainwater harvesting. On a larger scale, holding ponds and dams (reservoirs) are built to ensure water supply to people, small enterprises, industry, power plants, and, importantly, agriculture during prolonged droughts. There are four main reasons why these traditional methods are no longer sufficient to meet the water demands of people, industry, agriculture, and the biota (plants, animals, and bacteria): 1. As mentioned above, the growth of the human population in general and the growth of cities in particular, as well as people and industry’s unquenchable demand for freshwater due to intensification of water usage by all economic sectors have increased the need for abstraction of water from natural and man-made water resources. 2. Extensive use of land for human settlements and industrial complexes has diminished recharge of natural water
resources, groundwater in particular. In many areas, the groundwater table has dropped to a critical point (Mervis, 2009). 3. Excessive use of fertilizer and pesticides, unintended infiltration of leachates from municipal and industrial landfills, intrusion of seawater into aquifers, and discharge of poorly treated wastewater into rivers, lakes, and dams have caused major deterioration of surface water and groundwater quality. 4. In many regions of the world, global warming and the subsequent change of climatic conditions have led to severe irregularities of precipitation (Bates et al., 2008). Regions such as California, the Mediterranean countries, and Australia are reporting unprecedented drought situations. Dams are empty (Figure 1); rivers, fields, and gardens are drying up (Pearce, 2006). To maintain their water supply, many municipalities are now being forced to consider alternative sources of water. The city of Brisbane in Australia, for instance, decided to implement advanced treatment of wastewater, pump it back to and blend it with the water in the Wivenhoe dam, and use it as a source of municipal water (Figure 2). This project has been completed but has not been fully commissioned after rainfall intensity increased in the year 2009, and due to a public backlash over discharging treated wastewater to the water supply of Brisbane City. As early as 1998, Singapore embarked on a program of desalinating treated wastewater for further use in industry (Anonymous, 2002; Tortajada, 2006). The city of Perth in Australia, along with many other large cities in the world, has constructed plants for desalination of seawater. Desalination of brackish water is being considered in some other areas.
Figure 1 Low water of Lake Wivenhoe, Australia, after 6 years of drought. Photo taken by the author.
Abstraction of Atmospheric Humidity
Figure 2 Western corridor recycled water project designed to overcome water shortage-situation in the Brisbane metropolitan area, Australia; wastewater is considered as an alternative source of water. PP, power plant; WWTP, wastewater treatment plant. Adapted from McCann B (2008) Australia’s largest recycled water project. Water 21, Journal of the International Water Association (London) 21: 42–44.
Although advances in membrane technology have made these solutions feasible and affordable, the high costs and high energy consumption limit the broader applicability of such high-tech solutions. There is one other source, which needs to be taken into consideration when looking for solutions to the problem of water shortages – atmospheric humidity. This chapter summarizes potential methods of harvesting atmospheric humidity, and outlines the technology used to abstract this humidity for human consumption.
4.05.2 Volume of Water in the Atmosphere The generic source of freshwater on the Earth is the atmosphere (Figure 3; Rekacewicz, 2002). Water reaches the surface of the Earth as a result of precipitation in the form of rain, snow, graupel, or hail. Some of the rainwater evaporates on its way to the surface and thereafter. Thus, it is transferred back to the atmosphere before it can be used. Some of this water is taken up by plants and animals, and stored for some period of time in plant and animal tissues. Another portion of the rainwater infiltrates geological formations (aquifers) consisting of porous material (gravel and sand), or rock crevices and caverns. The rest flows above ground toward lakes, wetlands, and eventually to the open sea. The quantity of water in the atmosphere is subject to constant and highly dynamic changes. As part of the overall water cycle, atmospheric water is continuously replenished by evaporation from surface waters (sea, lakes, rivers, swamps, etc.), and by evapotranspiration performed by global biota in general, and by plants in particular.
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According to estimates published by Gleick (1996), the total amount of freshwater on the Earth is 34 650 000 km3, of which 13 200 km3 is present in the atmosphere, either in the form of gaseous water (water vapor) or as liquid water droplets forming fog or clouds (Figure 4). In comparison to the overall volume of water on the Earth (1 385 984 000 km3), the water content of the atmosphere appears to be rather low. However, if we were to distribute the water contained in the atmosphere among the 6.8 billion people on the Earth, at any given time each person would receive about 2000 m3. Considering that agriculture requires about 75% of the freshwater being consumed, and only 6% of the overall water consumption relates to domestic usage (drinking, cooking, body care, washing clothes, as well as water consumption by small enterprises), there is still plenty of water available in the atmosphere (about 115 m3 per person at any given time) to satisfy the basic needs of people on the Earth. In conclusion, it is well worth considering the water vapor contained in the atmosphere as a supplementary source of water – not only for human consumption but for agricultural irrigation and industrial purposes as well.
4.05.3 Fundamentals of Rainfall Generation 4.05.3.1 Preliminary Remarks The following basic information relates to the various forms of atmospheric water, the processes leading to fog and cloud formation, and about methods of accessing atmospheric water to mitigate water shortages at local and regional scales. It is not intended to provide an in-depth scientific review. Readers who are interested in digging deeper into the knowledge base of atmospheric physics and meteorology are advised to consult the relevant textbooks and scientific journals (e.g., Rogers and Yan, 1996; Steinfeld and Pandis, 2006).
4.05.3.2 Water in the Atmosphere Water exists in the atmosphere in all three thermodynamic aggregate states, as gas, as a liquid, and in the solid state of ice. The gaseous state of atmospheric water is termed water vapor or simply vapor. Changes of water-aggregate states drive a rich variety of atmospheric processes affecting weather and climate. The change of aggregate state from vapor to liquid is termed ‘condensation’, whereas the reverse process of change from liquid to vapor is termed ‘evaporation’. The change of aggregate state from liquid to solid is termed ‘freezing’, whereas the reverse process of change from solid to liquid is termed ‘melting’. Vapor can also be directly converted into ice, bypassing the liquid-state stage. This change of aggregate state from vapor to solid is termed ‘vapor deposition’ or simply ‘deposition’. Evaporation of ice, the direct change of aggregate state from solid to vapor bypassing the liquid-state stage, is the reverse process. Water vapor is invisible to the naked eye. The amount of vapor in a unit volume of air can be described in terms of water vapor partial pressure, that is, the pressure of vapor contributing to the total pressure of all gases comprising atmospheric air. At a given temperature, partial vapor pressure
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Vapor transport
Precipitation 9000 km3
Precipitation 458 000 km3
Precipitation 110 000 km3
Evapotranspiration 65 200 km3 Evaporation 502 800 km3
Evaporation 9000 km3 Lakes
River runoff 42 600 km3
Infiltration
Ocean Area of internal runoff 119 million km2
Groundwater flow 2200 km3
Area of external runoff 119 million km2
Oceans and seas 361 million km2
Note: The width of the blue and gray arrows are proportional to the volumes of transported water
Figure 3 Graphical representation of the water cycle on the Earth. Reproduced from Rekacewicz P (2002) Vital water graphics. United Nations Environmental Programme/GRID-Arendal. http://maps.grida.no/go/graphic/world_s_water_cycle_schematic_and_residence_time (accessed March 2010), Philippe Rekacewicz, UNEP/GRID-Arendal, with permission.
cannot be increased indefinitely, that is, the amount of vapor which a given volume of air can hold is limited. This limit is reached when the number of molecules evaporating from water or ice surfaces equals the number of molecules condensing into liquid water or depositing onto ice. Such an equilibrium state is termed ‘saturation’. At saturation, vapor is termed saturated, and the partial pressure of vapor is termed ‘saturation vapor pressure’. Saturation vapor pressure increases with air temperature, which means that a volume of cold air can hold a smaller amount of vapor than the same volume of warm air. A common measure of vapor content in the air is relative humidity (RH), defined as the ratio of partial vapor pressure to saturation vapor pressure at a given temperature, usually expressed as a percentage. At saturation, RH is 100%.
The temperature at which the vapor at a given partial pressure is saturated is termed ‘dew point’. If air is cooled, RH may exceed 100% (in this case, the vapor is termed supersaturated), unless surfaces which can be wetted or, at subzero temperatures, surfaces with a structure similar to that of ice, are available allowing the vapor to condense or deposit, respectively. Small airborne particles called aerosols, for example, fine sand, dust, ash, soot, bacteria, and pollen, are almost always present in the atmosphere. Some of those particles with surfaces on which vapor may condense or deposit are called ‘condensation nuclei’ (CNs) and ‘ice nuclei’ (INs), respectively. Typically, the supersaturation of atmospheric vapor does not exceed 1–2% as the latter condenses or deposits on INs
Abstraction of Atmospheric Humidity
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Water in 13 167 km3
− Atmosphere − Surface
125 431 km3
− Permafrost soil
277 197 km3
− Aquifers
10 429 530 km3
− Glaciers
23 804 275 km3 0
10 000 000
20 000 000
km3
Water in 13 167 km3
− Atmosphere − Plants and animals
1 109 km3
− Rivers
2 218 km3
− Soil
16 909 km3
− Wetlands
11 780 km3
− Lakes
93 415 km3 0
40 000
80 000
km3
Figure 4 Graphical presentation of the various categories of freshwater on the Earth. Data from Gleick PH (1996) Water resources. In: Schneider SH (ed.) Encyclopedia of Climate and Weather, vol. 2, pp. 817–882. New York: Oxford University Press.
and/or CNs, which grow into ice crystals and liquid droplets, respectively. Areas laden with airborne cloud particles such as ice crystals and/or liquid droplets, formed as a result of air cooling below the dew point, appear visually as either clouds or fog. The term cloud is used when a clear interface exists between the particle-laden space and the atmosphere below. In contrast, the term fog is used when no such bottom interface exists, and the particle-laden space meets the surface of the Earth (land or a body of water). The efficiency of aerosols acting as CNs and INs varies depending on their chemical composition and surface properties. In particular, cloud formation and development are sensitive to atmospheric pollution (Steinfeld and Pandis, 2006). Due to their hygroscopic properties, sea-salt particles are considered to be excellent natural CNs, causing liquid droplets to grow well beyond their normal size (Biswas and Dennis, 1971). Elementary sulfur, originating from dimethyl sulfide (DMS), a volatile compound generated by marine algae, is another example. In general, the ability to act as CNs varies for different types of aerosols, which makes cloud formation and development sensitive to atmospheric pollution (Steinfeld and Pandis, 2006). Droplets are microscopic in size (typical diameter ranges between 10 and 20 mm). They are light and tend to remain airborne. In contrast, drops are comparably large, and are heavy enough to fall by virtue of gravity.
4.05.3.3 Atmospheric Processes of Precipitation Formation 4.05.3.3.1 Thermodynamic processes Condensation, deposition, and freezing are thermodynamic processes of the aggregate-state change accompanied by the release of latent heat. As a result, the surrounding air becomes warmer. The reverse processes of evaporation and melting result in the cooling of the surrounding air. Evapotranspiration refers to the process of evaporation facilitated by plants and animals.
4.05.3.3.2 Unit processes Clouds are a potential source of precipitating water, but not every cloud delivers precipitation. Rainfall and/or snowfall can only materialize after the cloud particles have reached a threshold weight beyond which gravity takes effect. Although intensive research has been conducted over the past years, knowledge about the governing processes in clouds is still incomplete. In the following, processes which play a major role in the generation of rain and snowfall are briefly described. Figure 5 illustrates the network of interconnected processes in clouds. The capture of supercooled cloud droplets by snow crystals is termed as ‘riming’. When two droplets collide, we term this process as ‘collision’. Coalescence occurs when two droplets fuse.
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Abstraction of Atmospheric Humidity
-The source -
Evaporation
Evaporation Water vapor
Condensation
Melting
Cloud droplets
Ice crystals Coalition coalescence
Riming
Coalition coalescence
Condensation
Evaporation
Melting
Condensation
Condensation
Melting Snow flakes
Rain drops Riming Pr
ec
Lakes
ipi
ip rec
ion
n
tio
ita
tat
P
River runoff 42 600 km3
Infiltration
Ocean -The demand -
Figure 5 Network of processes involved in the development of precipitation. Adapted from Houze RA (1993) Cloud Dynamics, pp. 573–578. San Diego, CA; London: Academic Press; and background picture from Rekacewicz P (2002) Vital water graphics. United Nations Environmental Programme/GRID-Arendal, http://maps.grida.no/go/graphic/world_s_water_cycle_schematic_and_residence_time (accessed March 2010).
The schematic presented in Figure 5 may be somewhat misleading as it provides the impression of a rather universally applicable and static interaction of processes. In reality, however, the concert of processes is heavily affected by site-specific boundary conditions, and by natural and man-made impacts. Moreover, it is subjected to time variations high in frequency and amplitude. As shown in Figure 6, there exists a highly complex system of causes and effects which needs to be understood when attempting to influence weather conditions, and to trigger or enhance precipitation with the aim of mitigating local or regional drought situations.
4.05.3.3.3 Thermodynamic approach to explaining precipitation events If the temperature increases, the air can hold a larger amount of vapor and the corresponding vapor-saturation pressure increases. With rising temperature, the actual partial pressure of vapor falls below the saturation pressure, and RH falls below 100%. The molecule balance at the surface no longer exists and the evaporation of liquid water or ice into the air becomes the dominant process, which, if continued for a limited volume of air, will eventually bring the system back to the equilibrium state of saturation.
Abstraction of Atmospheric Humidity
e.g., - Cosmic rays - Solar radiation - Electrical charge distribution
Altitude
Orography
Geographic latitude
Albedo
e.g.,
e.g., - Air currents - Thermal uplift
- Air pollution - Contrails Vegetation lakes and wetlands ocean
Lakes
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Infiltration
Land use urbanization
e.g., - Volcano outbreak - Forest and bush fire - Man-made emissions
River runoff 3 42 600 km
Lakes
Infiltration
River runoff 3 42 600 km
Ocean
Ocean
(b)
(a)
Distribution stratification Density • Aerosols • Gases • Ice • Nuclei • Water droplets • Water vapor
Processes • Adhesion • Advective transport • Attraction • Coalescence • Coalition • Condensation • Deposition Temperature • Evaporation • Air • Freezing • Droplets • Melting Distribution • Ice • Repulsion scattering • Riming Electric charges • Turbulent mixing • Macroscopic • Microscopic Forces • Air currents • Buoyancy Riv R iivve ver runof runoff off o fff • Gravity 3 Lakes
In Infi IInf nfi n nfiltra filtration ltratition ltra on on
• Factors • Conditions • Processes
Are subjected to time-dependent changes
42 600 42 00 km
Lakes Ocean
(c)
Infiltration
River runoff 42 600 km3 Ocean
(d)
Figure 6 Attempt to visualize the complexity of weather related processes in the atmosphere: (a) geographic boundary conditions; (b) physical and chemical boundary conditions; (c) internal system parameters; and (d) all external and internal factors, conditions, and processes change, often rapidly, with time. Background pictures adapted from Rekacewicz P (2002) Vital water graphics. United Nations Environmental Programme/GRIDArendal. http://maps.grida.no/go/graphic/world_s_water_cycle_schematic_and_residence_time (accessed March 2010).
If the temperature decreases below the dew point, the amount of vapor which can be held by air also decreases, as does the corresponding vapor-saturation pressure. As a result, RH exceeds 100%. In this case, the air is considered to be supersaturated with vapor. This state of supersaturation is unstable as the excess vapor will condense on any available surface (liquid or solid) until the partial vapor pressure is reduced to the saturation pressure, and the system is brought back to equilibrium.
The vapor-saturation pressure over ice is lower than that over liquid water (Bergeron, 1935, 1949). Therefore, ice particles may grow faster than liquid droplets and eventually absorb more vapor by deposition than droplets do by condensation. As the water vapor is consumed by the growing cloud particles, its partial pressure decreases. When the partial pressure of vapor falls below the vapor-saturation pressure, the air becomes undersaturated with respect to liquid water while still being supersaturated with respect to ice. At this point, the
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Abstraction of Atmospheric Humidity
surrounding droplets will evaporate while ice particles continue to grow at the expense of droplets. This mechanism is known as the Bergeron process. Atmospheric aerosols, both of natural and anthropogenic origin, can dramatically affect microphysical processes in clouds. Water droplets are microscopic in size. Most of them grow by condensation to a size of about B105 m in diameter. Due to the surface tension of liquid water, the vapor-saturation pressure over small droplets is higher compared to that of a flat water surface. Therefore, smaller droplets require a higher supersaturation to break the size barrier to condensational growth. This size barrier can be surpassed if a solute such as salt reduces the saturation pressure enough to mitigate the effect of surface tension. Soluble aerosols (e.g., particles of sea salt which are abundant over oceans), when collected by and dissolved in liquid droplets, may alter the effective vaporsaturation pressure over droplets. Aerosols of surface-active substances emitted from industrial areas may also affect the saturation pressure over water droplets. As a result, the droplet spectrum (the distribution of number concentration over droplet size) may be significantly affected. As the collision efficiency of droplet coalescence increases, rainfall may be initiated and enhanced in intensity (Facchini et al., 1999). Pure water droplets can remain in liquid form even down to temperatures near 42 1C (Smith, 1999). Supercooled droplets are typically abundant in cold clouds (i.e., clouds at temperatures below freezing point). When brought into contact with a body of ice or any other substance with similar surface characteristics (e.g., silver iodide particles), the supercooled water may freeze almost instantly. This process is known as ‘contact freezing’ (Sastry, 2005). As a result of contact freezing, ice particles can be instantly formed from a liquid droplet bypassing the relatively slow Bergeron growth of ice-forming nucleus, which involves liquid droplet reprocessing via evaporation. In turn, the produced ice particle may continue to grow either by vapor deposition or by further merging with the next supercooled droplet. The produced particles, when grown large enough, may reach the surface as ice conglomerates, such as snow and graupel, or, if melted as they descend, as rain drops. Rain is also produced by warm clouds (i.e., clouds at a temperature above freezing point) during summer and in tropical areas. Therefore, it has to be assumed that processes other than those related to ice production cause the generation of rainfall. At a certain stage of droplet growth in a warm cloud, droplet merging by collision becomes a growth process which is considered to be even more efficient than condensation. As the force of gravity exerted on a droplet increases more with the size of the latter than the competing force of air viscosity, larger droplets fall faster than smaller droplets. As they descend, the former (in this context termed ‘collectors’) may collide and merge with the latter, thus growing in mass and size and falling faster. In this classic description, such a droplet-merging process is known as ‘collision–coalescence’, or simply ‘coalescence’ (Battan, 2003). The coalescent growth of a descending collector accelerates during its travel through the swarm of small droplets in cloudy air until reaching the cloud base, which may result in the formation of a drop, a liquid water particle, a millimeter in size. On exiting the cloud,
the descending drops start to evaporate as the surrounding air is no longer saturated. Depending on the vertical air-humidity profile and size spectrum of the produced drops, some of the latter may reach the surface of the Earth as rainfall. In turbulent air, droplets can move in different directions with different velocities, generally determined by a combination of the forces of gravity and viscosity in highly variable air motions, causing them to collide as in the case of classic collision–coalescence. Such a turbulent coalescence is sometimes called turbulent coagulation. There is no guarantee, however, that all geometrically possible collisions will result in merging small droplets with the collector, that is, coalescence. Some droplets may be deflected by airflow around the collector surface. Droplets may also coalesce temporarily and then separate, or coalesce temporarily and then separate breaking into a number of smaller droplets. The collision efficiency defines the ratio of actual number of collisions to the number of collisions which are geometrically possible. Moreover, the coalescence efficiency has to be taken into account, defined as the ratio of the number of successful coalescence events to the number of collisions. In addition, it has to be considered that, as in the case of classic collision–coalescence, the presence of large droplets, even in a small number concentration, may significantly enhance turbulent coalescence (Riemer and Wexier, 2005). It was Aitken who discovered the importance of CNs in 1880. Due to their force of attraction to water, sea-salt particles may be considered as prominent CNs frequently available in the atmosphere above the ocean. Elementary sulfur is another candidate for nucleation. It originates, for instance, from DMS generated by marine algae. Dust and fine sand are other media. Dust particles may comprise inorganic or organic matter. The latter may originate from blooming plants (pollen), forest fires, road traffic, or industrial operations. Finally, but importantly, ice crystals are rated as very important CNs and play a major role in the process of cloud formation. Understanding the properties of clouds is still limited by difficulties surrounding the problem of adequately describing the processes of cloud-droplet nucleation and growth. Small changes in droplet population may significantly influence the formation and size of cloud droplets and precipitation (Facchini et al., 1999). Solutes affect the equilibrium and nonequilibrium properties of water. The depression of the ice equilibrium melting point is one such example. Koop et al. (2000) found that homogeneous nucleation of ice from supercooled aqueous solutions is independent of the nature of the solute, but depends only on the ratio between the watervapor pressures of the solution and of pure water under the same conditions. In addition, the authors found that the presence of solutes and the application of pressure have a very similar effect on ice nucleation. A thermodynamic theory for homogeneous ice nucleation was suggested which expresses the nucleation rate coefficient as a function of water activity and pressure. Cloud droplets are so small in size and so light that gravity has little effect on them. The rather uniform size of these particles suggests that the rate of condensation matches fairly well with the rate of evaporation processes. For the particles to grow, gain weight, and eventually descend to the surface of the Earth as rain or snow, additional forces need to take effect.
Abstraction of Atmospheric Humidity
Condensation of water vapor onto the surface of droplets is one of the growth mechanisms to be considered. Ice particles can grow in a similar fashion by the condensation and subsequent freezing of water vapor. For water vapor to condense onto a liquid droplet, the air surrounding the droplet must be supersaturated with respect to water. Likewise, for condensation onto an ice particle, the air must be supersaturated with respect to ice. At sub-freezing temperatures, supersaturation with respect to ice occurs at a lower RH compared to supersaturation with respect to water (Bergeron, 1935, 1949). Therefore, in the same environment, ice particles will grow faster than water droplets, particularly when the ice particles are supercool with respect to the temperature of the air (ambient temperature). As the water vapor is depleted, the environment will become subsaturated with respect to water but will still be supersaturated with respect to ice. At this point, the water droplets will evaporate while the ice particles continue to grow. Thus, ice particles will grow at the expense of water droplets. For the Bergeron process to proceed, the cloud must be positioned at an altitude where the air temperature is well below freezing point (cold cloud). With respect to rainfall generation from warm clouds, the collision–coalescence process described by Battan (2003) may be considered as an alternative to the Bergeron process. As mentioned above, salt particles, because of their hygroscopic properties, attract water when present in the air, causing water droplets to grow well beyond the normal size of cloud droplets (Biswas and Dennis, 1971). As these droplets become heavy, they likely start descending through the swarm of cloud droplets toward the Earth’s surface. On their way, they will inevitably collide with smaller cloud droplets, making some, but not all, merge with the larger droplets unless there are surface-active chemical substances present, which lower the surface tension of water (Facchini et al., 1999). Subsequently, the larger droplets grow in size, and the speed of descent increases. Once the drops exceed 100 mm in diameter, rain drops of 1 mm and larger develop within minutes.
4.05.3.3.4 Electrical processes Cloud processes are also affected by electric charges on cloud particles. Similar to aerosols, electric charges and the resulting electric forces affect cloud properties and the process of precipitation formation via a variety of microphysical mechanisms. The first basic research in this field was conducted by Charles T. R. Wilson, who received the Nobel prize in 1927 for his method of making the paths of electrically charged particles visible by the condensation of vapor. Based on Wilson’s findings, Bernard Vonnegut was the first to develop this concept further (Phelps and Vonnegut, 1970). He and his colleagues discovered that the presence of electric forces enhanced coalescence and the formation of larger drops during collisions. This led the authors to speculate that electrical charges in clouds could aid in the coalescence of droplets and thus initiate rainfall. They argued that negative ions that flow to the positively charged tops of thunderstorm clouds and the point-discharge positive ions that are carried from the Earth toward an electrified cloud base were not necessarily dissipative of cloud electrification. Some of these ions could be
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moved in the convective overturn associated with a growing cloud, resulting in point-discharge ions being carried by updrafts high in the cloud where they attract more negative ions to the cloud. These, in turn, become attached to cloud particles near the cloud boundary and are transported downward by the unfolding of air in and around a growing cloud. Noting that intense rainfall often did not develop in clouds prior to the electric discharge, they proposed an electrostatic precipitation explanation based on the rearrangement of charges around the lightning channels (Moore and Vonnegut, 1973, 1997; Vonnegut 1984, 1995). Harrison and Ambaum (2008, 2009) have added more information to the work of Vonnegut and his colleagues, and endorsed a hypothesis on the role of ions in nucleation of cloud droplets, their growth into raindrops, and the resulting precipitation events. Today, it is known that all clouds – even non-thunderstorm clouds – are electrified to a certain degree. Although complex charge configurations may take place under highly variable atmospheric conditions, cloud particles are predominantly charged positively at the top and negatively at the bottom, thus forming a so-called space charge in those areas. In a cloud, this space charge maintains the associated electric field. Processes of cloud charging and precipitation formation are intrinsically linked. One effect of electric charges on cloud particles is called the electro-hygroscopic effect, that is, the reduction of the size barrier for the growth of water droplets. As with salt and other solutes, condensation is facilitated when droplets are electrically charged. Although the droplet charge required to enhance activation is substantial, Harrison and Ambaum (2008) demonstrated that sufficient charging occurs at the edges of even weakly electrified clouds. Mean droplet charges in the order of 100 elementary charges have been observed near cloud boundaries of non-thunderstorm clouds (Beard et al., 2004). In this context, the term ‘elementary charge’ is to be understood as the positive electric charge carried by a single proton or the negative electric charge carried by a single electron. Cloud droplets with a radius of 3 mm carrying an average 1500 elementary charges have been observed at the base of mountain-top stratocumulus clouds (Twomey, 1956). Charging a haze droplet to 1000 elementary charges can reduce the critical supersaturation to 0.5% (Harrison and Ambaum, 2008). Another effect of electric charges is the enhancement of droplet coalescence by electric forces between charged droplets (Khain et al., 2004). The concept of electrically enhanced coalescence was introduced by Phelps and Vonnegut (1970). Regardless of the relative electrical polarities of the colliding particles, the net electric force between them is always attractive due to electrostatic image forces (Tinsley et al., 2000). For droplets of similar charges, there is a long-range repulsive force, but by virtue of turbulence, droplets may be brought into a range within which image forces take effect. To effectively increase the collision efficiency, a droplet should possess at least a few hundred elementary charges. Not only collision efficiency but also coalescence efficiency is increased in the process of electrically enhanced coalescence. Beard et al. (2002) demonstrated that the coalescence efficiency for collisions between weakly charged cloud droplets with a radius of 55–105 mm is greater than 91%. The efficiency is likely to be greater than 95% even when the droplet charges are
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insufficient for a significant enhancement of collision efficiency. The coalescence of droplets may also be enhanced by their polarization in an external electric field (Schlamp et al., 1976). As in the case of electrically enhanced liquid-droplet coalescence, electric forces may also augment the merging of other cloud particles such as nonwater aerosols and ice crystals. For example, liquid droplets may collect (scavenge) charged salt particles and therefore grow. Electrically scavenged nonsolute aerosols may trigger supercooled droplets to freeze at elevated temperatures (immersion freezing). Contact freezing of supercooled droplets by electroscavenged ice nuclei is a particularly efficient freezing mechanism of ice-particle generation (Tinsley et al., 2000). Another effect is that the electrostatic field of electric charges directly facilitates the freezing of supercooled cloud droplets and thus the production of ice nuclei. Water molecules possess their own electric dipole moment (Figure 7). The oxygen atoms are more negatively charged than the hydrogen atoms, and the molecule has a bent shape. This means that water molecules attract each other electrostatically. Clustering is likely to occur, causing enhancement of the electrical dipole moment. Further enhancement may occur when these clusters are exposed to corona ion emission. If random motion of the molecules is low, the molecules tend to line up in an orderly fashion with the positively charged part of one molecule next to the negatively charged part of another molecule. In an electric field, water molecules will rotate to line up with the field. As Wei et al. (2008) were able to experimentally demonstrate, this is favorable for the freezing of supercooled water at elevated temperatures. The electric field polarizes cloud droplets. Each of the dipoles exerts an attractive force on others above or below, causing them to collide and become larger. This, in turn, triggers the collision–coalescence process. As a result, raindrops develop, which are larger in size compared to those that develop in an electrically neutral environment (Jermacans and Laws, 1999). Cloud and fog formation under an electric field was studied by Teramoto and Ikeya (2000) using a Wilson cloud chamber containing a supercooled atmosphere of ethanol. The electric field necessary to generate dense clouds was about
Slightly positive
H
O
H
Slightly negative Figure 7 Electric dipole characteristic of the water molecule.
4 kV m1. Positive ions produced by ionization condensed the nuclei for the generation of fog and clouds. Plume clouds were generated from a needle electrode. Clouds get electrically charged by external and internal mechanisms. The latter are related to processes which lead to precipitation. The generation of ice is believed to play a major role in internal cloud charging, that is, separation of opposite sign charges into different cloud areas. In thunderclouds, where the internal charging is intense and the positive feedback between charging and precipitation formation is strong, a liquid droplet may grow up to the size of raindrops within seconds. The sudden appearance of heavy rain, often called ‘cloudburst’, during thunderstorms is in agreement with this theory (Moore et al., 1964). Detailed discussions about a number of internal charging mechanisms are outside the scope of this chapter, but can be found in many publications, for example, in a book by McGorman and Rust (1988). External cloud charging relies on the electrical conductivity of the air within the global electric circuit (GEC). The latter is a model for the atmospheric electric system (Wilson, 1929). This system may be described as an electric circuit where electric charges are assumed to be separated by the polar-cap electric potentials generated by the solar wind and global thunderstorm activity mainly in the convective tropic regions. Figure 8 provides an overview of the respective mechanisms. In a simplified form, the charge separation in the GEC by thunderstorms can be described as follows. The negative thundercloud charge is transported to the surface of the Earth by ground flashes. The same process applies to a smaller amount of the positive charge. The rest of the positive charge leaks out of the cloud as the surrounding air becomes slightly conductive due to the presence of atmospheric ions. As the electrical conductivity of air sharply (quasi-exponentially) increases with altitude, most of this leak current is guided to the upper atmosphere, where it is distributed over the globe and maintains the ionosphere at a potential of about 250–300 kV with respect to the ground. The overall system resembles a spherical capacitor. In fairweather regions, the atmospheric ions are driven by the electric field (i.e., gradient of ionosphere-to-Earth potential), thus forming a leakage current with a density of about 1–4 pA m2. Atmospheric ions, which are constantly produced by cosmic rays, and at a lesser rate by surface radioactivity, are the carriers of the leakage current. The current which flows across the vertical column resistance is known as the ambient or fairweather current (Harrison, 2005). A schematic diagram of the GEC is given in Figure 9. Put simply, the essence of charge separation in the above process of external cloud charging is as follows. The initial charge separation on a microscopic scale occurs when pairs of ions of opposite polarities are created by energetic particles, mainly of cosmic origin. Then, the ions of opposite polarities are dragged apart in opposite directions by electric forces in the electric field of the GEC. Positive ions are moved downward, while negative ions are moved upward. Eventually, some of these ions get stuck on cloud particles, thus charging them positive for particles close to the top and negative for particles close to the bottom cloud boundaries. It is important to note that any charge separation requires energy input. The initial energy input to separate ion charges
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Solar wind modulation + + + + + + + ++ + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + +
Ion-neutral Cluster ions chemistry Recombination Attachment +− to aerosols Air ions Reduced − ion mobility Ionization + Enhanced + Corona + − + − space charge Convection Turbulent ions Aerosol − − D current transport − rift − − − −− − − cur nucleation − − − − Radon gas ren Enhanced t + + + + + E-field + − + + + + + − − − − − − − − − + − − − + − − − −− − − − −− − − Cosmic rays
Figure 8 Atmospheric processes relevant to ion–aerosol–cloud interactions. Reproduced from Harrison RG and Carslaw KS (2003) Ion–aerosol– cloud processes in the lower atmosphere. Reviews of Geophysics 41–3: 1012. Copyright 2003 American Geophysical Union. Reproduced by permission of American Geophysical Union.
Cosmic rays
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+ ++
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Charge separation in thunderclouds
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Planetary surface Figure 9 Schematic diagram of the global electrical circuit. Adapted from Harrison RG (2005) The global atmospheric electrical circuit and climate. Survey Geophysics 25: 441–484; inserted photo: Carina Hansen – Fotolia.com
to microscopic distances (i.e., to create an ion pair) is provided by energetic particles. The energy input required to separate ion charges to macroscopic distances is provided by the GEC, which acts as a generator of electric power.
All clouds, layered clouds in particular, get externally charged by the fair-weather current because the electrical conductivity of cloudy air is typically many times less than that of clear air at the same altitude (Zhou and Tinsley, 2007).
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This is mainly due to the attachment of ions to droplets and ice crystals that the cloudy air is laden with. The links between cloud electrification and precipitation have an important aspect. External charging may facilitate precipitation formation by electrical mechanisms in all clouds, including those where the internal charging processes are not yet developed. The latter applies to non-thunderstorm clouds which are sensitive to external charging, especially those of layered structure. Thunderstorm clouds provide an electric remote assistance for precipitation formation in non-thunderstorm clouds by virtue of the initial cloud charging over long distances on a global scale.
cooling agent is passed along, while on the other side there is air containing some humidity. This humidity turns into liquid water which subsequently can be collected and used (Figure 13). The idea has been picked up by a number of companies. In the following part, three examples are presented: 1. The Dutch company, Dutch Rainmaker BV, proposed to use electricity generated by a wind turbine to drive a heat pump, cooling the inner wall of a shaft holding the blades, and the
4.05.4 Innovative Abstraction Methods 4.05.4.1 Condensation Technology 4.05.4.1.1 General remarks As explained in Section 4.05.3.3, the transfer of water from the gaseous to the liquid state, that is, condensation, requires saturation, and even supersaturation of the air and the availability of a surface at which condensation can take place. Saturation is temperature dependent. Condensation begins as soon as the temperature falls below the dew point. When this happens, for instance, on a clear night on a spider web, dew drops form and accumulate on the meshes of the web (Figure 10). Similarly, condensed water accumulates on leaves early in the morning after a clear night (Figure 11). The condensation process continues when the surface upon which condensation occurs is artificially cooled. This process is termed as ‘forced condensation’, and can be visualized by a simple experiment: Figure 12 shows a pitcher filled with iced water. Almost instantaneously, a puddle forms at the bottom of the glass as condensation occurs. Eventually, the liquid water, which accumulates on the surface, starts to flow downward.
Figure 11 Condensation of liquid water at a leaf serving as condensation surface. Photo: makuba – Fotalia.com
4.05.4.1.2 Proposed technologies The process of forced condensation can technically be applied to convert atmospheric humidity into liquid water to be used for various purposes. All that is needed is a surface upon which condensation may proceed. On one side of the surface a
Figure 10 Condensation of liquid water at spider net strings serving as condensation surface. Photo: Dalia Ruckiene – Fotalia.com
Figure 12 Condensation of liquid water at the outside of a pitcher filled with iced water.
Abstraction of Atmospheric Humidity
electricity generator or a heat pump, respectively. Ambient air is blown through the shaft. Water vapor contained in the air condenses on the wall, drips down, and is collected in a storage tank at the bottom of the installation (Figure 14).
At the company’s test facility in Harlingen, a prototype was producing around 0.5 m3 of water per day based on a relative humidity of 45% and a wind speed of 2 m s1. Under the same conditions, the full-scale version is expected to produce between 7000 and 8000 l d1. 2. Another technology has been developed by the UK-based company, Grimshaw. Proposed is a unique large-scale condensation structure for the city of Las Palmas in the Canary Islands (Figure 15). The structure was designed in a bold sculptural form as a backdrop to an outdoor amphitheater. Technically, this apparatus may best be described as a tube-and-fin-type condenser. Cold seawater is pumped through the condenser elements to cool the outer condensation surface. If the atmospheric conditions are right, wind-driven humid air from the ocean passes through the structure. Water vapor condenses on the tube surfaces; the condensed water drops down, is collected at the bottom of the structure, and then transported by gravity to an underground storage tank. From there, the water may be treated by appropriate physical and chemical methods, and used for domestic purposes. In this example, the proposed pipe length was 400 m, and the condenser area was 1000 m3. The temperature of the seawater at the inlet point is typically about 9 1C. The air temperature at Las Palmas varies between 15 and 23 1C in summer, and between 12 and 18 1C in winter. Thus, the water production of the unit was calculated to be 1120 m3 per day in summer and 530 m3 per day in winter (Pawlyn, 2007).
Interfacial Boundary wall layer Temperature
Liquid water film Warm
Cold
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Flux of water vapor < Dew point temperature
Distance
Flow of liquid water Figure 13 Schematic representation of forced condensation of water vapor at a cooled surface.
Wind turbine
Cooling of the innerpart of the shaft
Heat pump
Air filter Air flow Blower
Collection basin
Effluent
Figure 14 Simplified schematic representation of a water abstraction device driven by a wind turbine. Illustration based on a sketch provided by Dutch Rainmaker bv.
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Abstraction of Atmospheric Humidity Condensation structure
Humid air from sea
Amusement theater Return flow
Collection of condensed water
Water for further use Underground storage
Cold seawater intake Figure 15 Condensation of humid air blown from sea through an artistically designed structure – simplified overview based on a drawing of Pawlyn M (2007) An architecture of water purification. In: Huber H, Wilderer PA, and Paris S (eds.) Water Supply and Sanitation for All, pp. 131–136. Berching, Germany: Hans Huber AG. Photo by permission of Grimshaw, London, UK.
3. The third example of condensation technology has been developed by a UK-based consortium comprising Seawater Greenhouse Ltd., Exploration Architecture Ltd., and Max Fordham & Partners LLP. It is called the Sahara Forest Project, and it aims to create a growing environment in hot and dry parts of the world (such as the Sahara), and produce deionized water comparable in quality to rainwater, from seawater using solar energy. The proposed method is based on the so-called seawater greenhouse concept developed by Paton (2001). Its purpose is to provide desalination, cooling, and humidification in an integrated way, using solar energy as the main source of power. Davies et al. (2004) described this concept in some detail. In the Sahara Forest Project, the seawater greenhouse concept is combined with the concentrated solar power (CSP) technology, which is already applied at various locations around the world. Solar radiation is concentrated and focused on a heat-exchanger system to create steam that drives conventional turbines to produce electricity. CSP plants produce large amounts of surplus heat which can be used to evaporate seawater. The vapor may be distributed in the greenhouse structures where it is converted by natural condensation into deionized water, and used to water trees, shrubs, and food crops inside and outside of the greenhouses (Figure 16). Eventually, a forest ecosystem develops outside of the greenhouses which has the potential to change the microclimate of a region. It may even work as a biotic pump, which means that humidity from the ocean is attracted, clouds are formed, and precipitation
occurs in the former desert (Makarieva and Gorshkov, 2007).
4.05.4.1.3 Research needs As sufficient basic knowledge on condensation is already available, exploitation of this knowledge in order to overcome water shortages in arid countries deserves attention. The three examples described above clearly demonstrate the potential of condensation technologies. The field is wide open for more innovative ideas and concepts followed by applied research, technology development, and field trials. Certainly, condensation-based technology will not solve the water crisis at large. Nevertheless, the deployment of many small-scale solutions will assist in the mitigation of local water-scarcity problems.
4.05.4.2 Fog Collection Fog is differentiated from clouds by the fact that it occurs close to the Earth’s surface (land, lakes, or sea), whereas clouds are located high in the lower atmosphere in the form of either horizontally oriented layers (stratus, cirrus) or vertically oriented accumulations (cumulus). Fog, as do clouds, consists of very small droplets of liquid water (see Section 4.05.3). Fog droplets have diameters of about 1–10 mm. Under the influence of gravity alone, these droplets fall very slowly (o1 to about 5 cm s1) and are thus readily subjected to wind force. Horizontal winds of a few meters per second cause even the largest fog droplets to travel horizontally (Schemenauer and Cereceda, 1994a).
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Ocean Forest plantation Greenhouse installation
Concentrated solar power plant Figure 16 Schematic representation of the Sahara Forest Project proposed by Seawater Greenhouse Ltd, Exploration Architecture Ltd, and Max Fordham & Partners LLP. Reproduced from Pawlyn M (2007) An architecture of water purification. In: Huber H, Wilderer PA, and Paris S (eds.) Water Supply and Sanitation for All, pp. 131–136. Berching, Germany: Hans Huber AG.
Fog droplets are readily intercepted when they come into contact with surfaces, for instance, with leaves or plant stems. They accumulate on such surfaces, forming large drops and even films of liquid water. This is similar to (but should not be confused with) the surface accumulation of liquid water caused by condensation. Some of the water accumulated on leaves and stems may be taken up by plants or consumed by small animals. Excess water eventually drips or flows to the ground and contributes to soil moisture. Thus, in many arid areas, fog is the only source of liquid water to support vegetation and animal life. Persistent fog not only provides water, but also controls the natural water management of ecosystems. In the humid tropics, these regions are referred to as cloud forests (Kerfoot, 1968; Stadtmiiller, 1987; Goodman, 1982). Trees are potentially good collectors of fog. Indeed, trees have been used by man as fog collectors for centuries. In the seventeenth and eighteenth centuries, stories about three fogcollecting trees which grew on Hierro Island (Canary Islands) received widespread public attention (Glas, 1764). The trees were discovered in 1565 by Antonio Hernandez who coined them as ‘fountain trees’ (Figure 17) because the water captured on the leaves accumulated in large quantities before dripping down like a shower of rain. The water was collected in cisterns, which were divided in two parts, one for people and the second for cattle and other animals. Today, this concept is depicted on the coat of arms of Hierro Island. The trees belonged to the species of endemic laurel trees (Ocotea foetens). They stood atop cliffs where fog arrived on an almost regular basis from the ocean. Until a huge hurricane uprooted them in 1610, they served as major water sources for the pre-Hispanic population living on the island. In 1945, Don Zo´simo Herna´ndez Martin planted a laurel tree at one of the previous sites. The project was scientifically supervised by Alain Gioda, who in 1993 received the Rolex prize for his groundbreaking work (Gioda et al., 1993). Since 1993, this tree has been providing water obtained from fog in a manner very much like that described by Glas (1764).
Another example of a natural fog-collection system is that found by Schemenauer and Cereceda (1992a) in the Dhofar region of southern Oman. Two small intertwined olive trees stand in a windy environment where fog is almost constantly present. The vertical cross section of a tree can collect water at a rate of about 10 l m2 d1. Over an 83-day observation period in 1990, 580 l d1 of water was harvested (Schemenauer and Cereceda, 1994a). Realizing that fog is a rich source of water, and also due to the rapidly increasing need for water, particularly in rural areas of developing countries, in the early 1980s, Schemenauer and his colleagues commenced extensive research, development, and implementation of projects aimed at mitigating waterscarcity threats around the world (Schemenauer and Joe, 1982; Schemenauer and Cereceda, 1993). The results of these projects have been published in major scientific journals and presented at various international conferences and workshops. Key findings can be summarized as follows:
• •
•
•
•
Capturing and collecting fog is to be understood as a physical–chemical separation process. Fog droplets are separated from the air in which they are distributed. Interception by plant material (leaves, stems, etc.) followed by accumulation and storage can be mimicked by placing in a fog-laden environment, a material which is able to attract water droplets. For the collection of fog droplets, deep, three-dimensional net-like structures made of flat sheets of hydrophobic plastic material proved to be superior to two-dimensional sieve-like structures. Good collection efficiencies were achieved with a doublelayer net made of black polypropylene ribbon (flat ribbon of about 1 mm width and 0.1 mm thickness). The ribbons are woven into a net with a pore size of about 1 cm. Greater amounts of water can be collected when the droplet-laden air (fog) is driven toward and through the net by wind. To achieve optimal droplet-detention time within the net, airflow velocity within the net should not exceed
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Figure 17 Historic drawing and Hierro0 s coat of arms, both depicting the Fountain Tree used to harvest fog on Hierro Island, and the cistern to collect and store the captured water.
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and should not drop below an optimal velocity range. Optimal collection efficiencies were achieved at air-flow velocities between about 3 and 8 m s1 (Schemenauer and Cereceda, 1994b). Due to their small size and mass, fog droplets travel mostly parallel to the surface of the land. Therefore, the fog collectors (racks holding the net) should be oriented in a vertical direction. Fog droplets attached to the surface of the ribbon merge with other droplets by virtue of coalescence. Eventually, the accumulated water loses hold and drips down into a collection gutter. Field measurements of the collection efficiency of the net at the centerline of a large collector gave values of about 66% (wind speed: 3.5–6.5 m s1). This is in good agreement with the theoretical collection efficiency for a single ribbon once the areal coverage of the net is taken into account. At El Tofo, Chile, 50 fog collectors, each consisting of 48 m2 net area, were constructed in 1987. The average water production from this collector system was approximately 3 l m2 d1. This equates to an average production of 11 000 l d1. Production rates varied with the prevailing meteorological conditions from zero on clear days to a maximum of about
•
100 000 l d1, which provided 330 people living in a nearby village with 33 l per capita per day. Due to the remote location of the El Tofo plant, the effluent of the fog collectors was of good quality. It contained some marine salts and soil dust, but little contamination from anthropogenic sources (Schemenauer and Cereceda, 1992b). The measured water quality satisfied the drinking water standards of the Chilean government and the World Health Organization (WHO).
Many fog-collector systems have been installed in South America, Africa, Oman, and Canada. Figure 18 illustrates the typical installation of a fog-collecting system. A simple structure holds nets through which fog passes. The captured water is collected in a half pipe below the net and directed into a storage tank. From the results obtained, it can be concluded that fog collection is a viable and effective low-cost method to abstract water from the atmosphere and use it for domestic purposes, agricultural irrigation, and as a protective measure against forest fires (Walmsley et al., 1999). Foggy days are frequent in many regions of the world, even in some desert areas. In such locations, the collection of fog
Abstraction of Atmospheric Humidity
Figure 18 Typical setup of a fog collection system. Background photo: Fotalia.
for subsequent domestic use appears to be a most attractive method of overcoming water-scarcity situations. Of course, fog-collection technology on its own will not solve the water crisis. However, the deployment of many small-scale solutions will, at least, help mitigate water-scarcity problems. It is therefore worthwhile to further develop fog-collection technologies. Of particular interest here are innovative materials, with special focus on the surface properties of the materials to be used to capture fog droplets. Additionally, more research is required to improve the fog-capturing capacity of the threedimensional structure of collection nets. Finally, the quality of the collected water in relation to the atmospheric conditions at various sites needs to be carefully monitored.
4.05.4.3 Generating Clouds with the Aid of Heat Islands When there are no clouds in the sky, rain can hardly be expected to fall. With this in mind, scientists and engineers have been searching over the last few decades for ways to trigger cloud development. In order to plan and execute cloud-generation technology, it is essential to start with a sound knowledge of how this phenomenon occurs in nature (Battan, 2003). The fundamentals of cloud formation and of processes leading to precipitation have been described earlier in Section 4.05.3. As shown in Figure 3, cloud formation is very often stimulated by thermal uplift. Air containing humidity is lifted up into regions of the atmosphere where the temperature is low enough for conditions of saturation, even supersaturation, and the formation of cloud droplets. Taking this sequence of processes into account, it appears that any method which enhances thermal uplift could favor cloud formation. To trigger a thermal uplift, a relatively large area of land needs to be covered with a material of low albedo that readily absorbs short-wave solar radiation and converts it into longwave radiation (i.e., heat). Subsequently, the temperature of air above such an area would rise, causing the air to ascend. A piece of land which exhibits high-level radiation-absorbing properties is called a heat island. Bare, rocky land of volcanic origin may possess such properties, but so would man-made settings, such as a city where the roads are paved
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with asphalt, and roofs made of red tiles (Black and Tarmy, 1963). Such environments may be classified as ‘unintended’, in contrast to purpose-built heat islands. To purposely create a heat island, it is necessary to pave a large area of land with a material such as asphalt, preferably close to the ocean. On sunny days, small deep-pressure systems may develop above the artificial heat island. Air would be sucked from the ocean or from wetlands where humidity is typically high. Vapor would thus be transported in large quantities toward the heat island and up into the sky leading to the accumulation of cloud droplets and, subsequently, to the development of convective clouds. These clouds may then be transported by the dominant wind some distance away from the heat island depending on wind direction and speed, and also on a variety of other prevailing meteorological factors in the site above the target area (Figure 19). Brenig et al. (2001) developed such a system based on the layout of an artificial heat island. He coined this method Geshem Rain System (Geshem means rain in Hebrew). The sequence of processes characterizing the heat-island concept was first described in the 1950s and 1960s by Malkus and Stern (1953), Malkus (1963), and Black and Tarmy (1963). The theoretical arguments of Malkus and Stern were based on very rough solutions of the hydrodynamic equations governing atmospheric flows. The observational data were derived from a small, flat island in the West Indies that had been thoroughly studied in order to understand the mechanisms which lead to frequent development of cloud rows in an area where clouds were typically not present. The study by Malkus and Stern stimulated some further studies on the potential of urban areas to serve as heat islands. As mentioned above, radiation-absorbing materials are typically used to build roofs and roads. These materials absorb solar radiation better than the vegetation in the surrounding countryside. The larger the area that is covered with such materials, the greater the expected heat flux toward the lower atmosphere. At the same time, small particles emitted by cars (e.g., abrasion of brakes and tires), by industrial activities, and heating devices will be uplifted as well. Rainfall is likely to occur in the downwind region. This phenomenon was studied by Shepherd et al. (2002) for unintended heat islands in areas such as Houston, Texas, by examining data measured by a radar system on board the Tropical Rainfall Measuring Mission (TRMM) satellite of National Aeronautics and Space Administration (NASA). This system measures rainfall rates, droplet size, and latent heat. Two main effects were observed in and near the southern US cities studied: first, an average temperature difference of 3 and 5 1C between the cities and the surrounding area was measured; and second, significantly higher rainfalls and thunderstorms occurred in an area away from the cities in the prevailing wind direction (Burian and Shepherd, 2005; Shepherd et al. 2002). Studies performed by Brenig et al. (1995, 2001, 2005) revealed that there is a lower limit for the size of the artificial heat island and a higher limit for its reflectivity beyond which the buoyancy produced would not be enough to lift the air to condensation altitude. The value of the minimum area for the low-reflectivity surface is a function of its albedo. For lower albedo values, a smaller heat-island surface produces about
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The Geshem rain system Cumulus Rain Wind Rain
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Figure 19 Schematic representation of the Geshem Rain System, the engineered version of the general heat island concept. Reproduced from Brenig L (2007) Making rain on arid regions: The Geshem Rain System. Water and Environmental Exchange, Sevilla (Spain). http://physfsa.ulb.ac.be/IMG/pdf/ brenig07.pdf (accessed August 2010).
the same thermal rising motion for the air as a larger but lessabsorbing black surface. Moreover, these limiting values are also dependent on factors such as wind speed, atmospheric stability, and the thermal parameters of the radiation-absorbing material used. For higher wind speeds over a ground surface of a given albedo and a given size, the buoyancy effect gets weaker since the air flows faster over the hot surface and, consequently, has less time to absorb the heat ascending from the ground. For higher stability of the lower atmospheric layers, the ascending motion of the air will be counteracted by a stability effect. These are some examples of the rather complex set of physical relationships that govern the influence of a solar-absorbing surface at ground level on the local atmospheric circulation. According to Brenig (2001, 2005), for a heat-island system (Geshem Rain System) to function efficiently, the following conditions need to be met:
•
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The material constituting the heat island should reach a minimum temperature excess of 40–50 1C with respect to the surrounding area. In this case, the size of the heat island can be kept as small as 60–90 km2. The general orientation of the heat island should be with its longest axis parallel to the mean wind direction. Ideally, the center of the artificial heat island should be located at a distance between 10 and 30 km from the coast (Yoshikado, 1992, 1994). Closer to the ocean, the clouds generated by heat-island forces may be subject to diurnal
land–sea wind variations, and transported to the ocean rather than to inland areas. One of the major drawbacks of the Geshem Rain System is the enormous amount of space required to build an artificial heat island. In order to reduce this area, research is needed to identify materials exposing a maximum radiative absorption capacity and a low albedo at minimum cost. In collaboration with experts in land management and regional planning, possibilities should be investigated and assessed to deploy artificially built heat islands for multiple purposes, for example, for heating up the air and for electricity generation. In no case, however, should these heat islands compromise the functioning of local ecosystems or the supply of agricultural products. Triggering the development of cumuliform clouds does not necessarily lead to rainfall, as discussed in Section 4.05.3. In order to make optimal use of the water contained in the artificially generated clouds, research should focus on advances in technology which enable the generation and enhancement of rainfall.
4.05.4.4 Cloud Seeding 4.05.4.4.1 Development of the technology It is commonly known that clouds do not always bring precipitation to the surface of the Earth. From the literature review in Section 4.05.3, we have learned that precipitation can only occur when water particles, liquid or ice, gain sufficient
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weight to overcome the prevailing buoyancy in clouds and travel downward to the surface of the Earth. However, on their way, particles are subject to the process of evaporation. Only those drops or snowflakes make it to the grounds which, during their travel time, are not transformed back into water vapor. Taking into account the complexity of processes and boundary conditions shown in Figures 5 and 6, engineered harvesting of clouds appears possible only when, under actual meteorological, time-dependent conditions, the cloud particles (water droplets or ice crystals) are allowed to form, are triggered to merge, grow large, and gain weight. In principle, enhanced growths of cloud droplets could be achieved if the air temperature within clouds could be artificially lowered. In the case of clouds containing supercooled droplets, rapid cooling – for instance, by injecting dry ice (frozen CO2) – is a potential trigger for turning droplets into ice particles. It can be expected that the ice particles will grow since condensation with respect to ice is particularly high. Vincent Schaefer was the first to suggest this idea and investigate its applicability. In the laboratory, he was able to achieve positive results, and on 13 November 1946, he conducted a field test in the vicinity of Mt. Greylock (MA, USA). In this test, an airplane flew across a supercooled stratus cloud and dropped dry ice particles along its flight track. Within minutes, the texture of the clouds significantly changed, and below the cloud, snowflakes were detected. Under the leadership of Irving Langmuir, Nobel prize laureate for chemistry (1932), and in cooperation with Vincent Schaefer and Bernard Vonnegut, experiments continued. The group tried to find a substance which would be as effective as dry ice, but which would work at temperatures closer to the freezing point of water. It was Vonnegut (Battan, 2003) who identified silver iodide (AgI) as a potential seeding material. Silver iodide is highly soluble in water (3 107 g per 100 ml at 20 1C). It is used in photography and as an antiseptic in medicine. With respect to cloud physics, it is worth mentioning that the crystalline structure of AgI is similar to that of ice crystals. Thus, by injecting AgI crystals into a supercooled cloud, the availability of CNs is likely to be improved. Subsequently, supercooled cloud droplets are likely to develop, providing the opportunity for these droplets to instantaneously change into ice particles. Subsequently, the ice particles may grow in size and weight due to condensation processes. For silver iodide to convert into its crystalline structure, it has to be exposed to high temperatures at which the substance vaporizes. Cooling the AgI vapor results in very small crystals, 0.01–0.1 mm in size. They are similar in structure to ice crystals, and thus behave in a manner similar to that of CNs. The problem is, however, that these crystals deteriorate very quickly due to solar radiation. Thus, it is crucial to deliver the crystals to clouds as quickly as possible, by rockets or airplanes. Not only inorganic but also living bacteria have been considered as potential media for ice nucleation. Levin et al. (1987) studied the ice-nucleating properties of a number of Gram-positive and Gram-negative bacteria. They concluded that there is no reason to disqualify active bacteria as cloudseeding agents.
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Of particular interest in this context is the bacterial strain Pseudomonas syringae, which is known to cause ice nucleation at temperatures as low as 8 1C. Drainas et al. (1995) performed studies on the genetic properties of this strain, and explored the reasons why it causes water to freeze at temperatures below 0 1C. On ground, however, these bacteria are known to be responsible for severe surface-frost damages in plants. Tegos et al. (2001) conducted experiments to produce cellfree active INs for biotechnological applications in efforts to avoid detrimental effects once rain comes into contact with plants. A freezing temperature threshold of about 7 1C was observed. It is foreseeable that sooner or later, this type of seeding agent will be used in atmospheric studies as well. The method described above is often referred to as one based on the static-phase hypothesis. In other words, it is based on cloud microphysics. It is assumed that precipitation efficiency can be increased by altering the dynamics or air motion in clouds due to the latent heat release of growing ice particles, redistribution of condensed water, and counteraction against evaporation of cloud droplets, ice crystals, snowflakes, and raindrops. The second basic hypothesis of cloud seeding is the dynamic-phase hypothesis, which is based on the dynamics of clouds. Here, cloud seeding is focused on enhancement of the vertical air currents in clouds. Klatt’s (2000) description of the dynamic phase hypothesis is explained below. Cumuliform clouds are dependent on the presence of a persistent updraft. Air within the updraft experiences adiabatic cooling as it rises, and at some point it will become supersaturated with respect to water. The updraft speed is proportional to cloud buoyancy, the latter being a function of the temperature difference between the cloud and its environment. As water vapor is converted into liquid droplets or ice particles, it releases latent heat to the cloud, thus increasing temperature. This will enhance the updraft and increase watervapor influx. Positive feedback will occur as the increasing quantity of water vapor condenses, deposits, and releases even more latent heat. Particles suspended in the updraft may eventually grow large enough to overcome the upward velocity of the updraft and fall to the ground as precipitation. Precipitation has a very detrimental effect on the cloud. Clouds, which develop in areas where the shear is weak, will have a vertically oriented updraft. In this situation, precipitation which forms will fall straight down through the updraft. The weight of the drops and the drag created as they fall will dissipate the updraft. In addition, the precipitation removes large amounts of water from the cloud, which the updraft can no longer replenish. Once the updraft has ceased, the cloud will quickly evaporate. Observations have shown that seeding does enhance the transformation of cloud particles from liquid to ice (Sax and Keller, 1980; Hallett, 1981). According to the static-phase hypothesis, clouds should be seeded to achieve about 1–10 ice crystals per liter at temperatures warmer than 15 1C. In contrast, the proponents of the dynamic-phase hypothesis suggest seeding clouds such that more than 100–1000 ice crystals per liter may develop. This corresponds to seeding as much as 200–1000 g of silver iodide (Cotton, 1997).
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Dry ice and silver iodide enhance the ice crystal concentrations in clouds by either nucleating new crystals or freezing cloud droplets. They both belong to the category of glaciogenic materials. It should be realized that both substances can only be effective when clouds are exposed to temperatures well below the freezing point. In tropical areas, however, the air temperature remains above the freezing point even at high altitudes. To propagate enhancement of rainfall in such areas, the injection of water drops, dilute saline solutions, or grinded salt has been suggested. The primary objective of introducing these types of materials is to jump-start the coalescence process (Biswas and Dennis, 1971; Murty, 1989; Czys and Bruintjes, 1994). For seeding with salt, the term hygroscopic seeding is often used.
4.05.4.4.2 Evaluations and recommendations Over the past few decades, many attempts have been made to gain a deeper understanding of the underlying physics, to document the responses of cloud seeding, and to validate the causality, that is, the cause-to-effect relationship. The review in Sections 79.3.3.2 and 79.4.4.1 suggests that knowledge about cloud physics and the processes leading to precipitation has reached an advanced level, although more research is needed to consolidate the knowledge base. It is commonly agreed that the success of cloud seeding should be evaluated in terms of the amount of rain or snow that reaches the ground. The development of cloud particles heavy enough to fall is certainly a prerequisite for precipitation, but it is an insufficient criterion from an economic viewpoint. Only rain and snow which reaches the ground can be considered beneficial for humankind, industry, and nature – or detrimental if it exceeds a certain level of volume. Proving that a particular seeding exercise has caused rain or snow has been and still is a matter of controversial discussions. The problem here is that weather conditions are not only complex, but also highly variable in space and time (see Figure 6). Even if the cause-to-effect relationship could be scientifically verified, there is no guarantee that the public will accept this as an unarguable fact. Many people assume that a rain event can only be attributed to natural processes or to a divine power. A farmer whose land receives rain during a drought would most likely be reluctant to pay for this blessing, but would be more inclined to offer up a prayer to the heavens. Beginning with the early trials carried out by Schaefer and Vonnegut, efforts have been made to demonstrate that cloud seeding enhances rainfall or snowfall on the ground. Over time, the level of sophistication has increased – in terms of both test design and the evaluation of results using statistical methods. However, when Battan (2003) first published his book in 1962, he came to the conclusion that despite the evidence that cloud seeding leads to the enhancement of precipitation on the ground, more scientific knowledge was needed to better understand the physical mechanisms involved. He cited a report published in 1955 by the World Meteorological Organization, in which the authors state that a net increase in precipitation had not yet been demonstrated beyond reasonable doubt in any seeding operations. Years later, and after a multitude of field trials, Bruintjes (1999)
provided a critical review of cloud-seeding experiments. His critique is backed up by Garstang et al. (2004), who refer to a report of the US National Research Council (NRC) issued in 2004, which states that the field of atmospheric science is now in a position to answer many of the crucial questions that have impeded or blocked progress in weather modification in the past. However, the authors of the NRC report could still see no convincing scientific proof that cloud seeding works. Boe et al. (2004) countered that there is ample evidence that winter-fog modification, snowpack augmentation, and glaciogenic and hygroscopic seeding enhance rainfall, even though the magnitude of the effects may be difficult to quantify with precision. Qiu and Cressey (2008), referring to the failure of seeding operation during the Olympic games in Beijing, again expressed doubts about the effectiveness of cloud seeding. The question here is how to measure reliably the success of cloud seeding. Statistical analysis has been and still is the method of choice. To execute statistical analysis, it is commonly agreed that cloud-seeding experiments must be randomized. Although this method is well established in science, for fundamental reasons, the results of statistical analysis cannot prove that cloud seeding produces an effect such as rainfall on the ground. Statistical analysis can only provide information about levels of confidence and probability (List, 2005). Apparently, the statistical probability of cloud-seeding operations being successful is rather low. However, this conclusion might be misleading. As some experiments were randomized, the timeline of the three-dimensional meteorological conditions during and after the seeding event was not properly taken into account (see Figure 6). The actual meteorological conditions must be understood, however, as being decisive for success or failure. It is widely accepted that actual meteorological conditions are only occasionally appropriate for cloud seeding. If cloud-seeding operations are executed irrespective of the prevailing meteorological conditions, then it is little wonder that the success rate is rather limited. Nevertheless, according to the current state of scientific knowledge, it is fair to hypothesize that a seeding exercise under the right meteorological conditions will lead to precipitation on the ground (List, 2005). This hypothesis is yet to be proven incorrect. Future research and development efforts should concentrate on the development of a holistic physical hypothesis that incorporates all the major processes governing the generation of cloud droplets, ice particles, and eventually precipitation. Mathematical models need to be further developed to better understand the dynamics of meteorological conditions in time and space, and in response to the peculiarities of the region under consideration. Cloud seeding could eventually evolve into a technology which is predictable in effect, and could be pinpointed to a certain region. It would then become an attractive technology for investors. The proponents of cloud seeding must understand, however, that rainfall generation and enhancement are measures which can help solve regional or global water-supply problems only when treated as a part of an integrated waterresources management (IWRM) approach. Application of the technology has to be governed by national and international laws to ensure that conflicts of interest are avoided and
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sustainable development of regions is secured (for more information see Section 4.05.6). There is another concern not properly shared by the majority of cloud seeders and their economic and political proponents. Chemical substances and bacteria injected into clouds will inevitably end up on the ground and affect the biosphere. Silver, for example, is known to be potentially toxic. Similarly, bacteria such as Pseudomonas syringae are toxic (Drainas et al., 1995). Seeding clouds with substances which are or may be harmful for plants, animals, or humans is not only a shortsighted concept, but ethically and ecologically unacceptable.
4.05.4.5 Rainfall Enhancement by Cloud-Particle Charging 4.05.4.5.1 Scientific background In Section 4.05.3.3.3, we discussed the importance of atmospheric electricity on the evolution of cloud particles, and on subsequent occurrence and intensity of precipitation. It was mentioned that Phelps and Vonnegut were probably the first to consider the role of electricity in the processes leading to precipitation (Vonnegut and Moore, 1958; Phelps and Vonnegut, 1970) and to realize that the intended introduction of electrical charge into existing clouds could trigger or augment precipitation. Since then, numerous attempts have been made to develop weather-modification methods and devices for applications such as fog dissipation and precipitation enhancement by means of cloud-particle charging. As previously mentioned, there are two main microphysical mechanisms of cloud modification by particle charging: (1) electrically enhanced coalescence and (2) electro-freezing of supercooled droplets by contact nucleation. To be effective, each mechanism requires a certain minimum charge per electrically active cloud particle. For example, enhancing coalescence–collision efficiency requires hundreds of elementary (electric) charges on droplets with a radius of 10–20 mm (Khain et al., 2004). Charges per particle of the same order of magnitude are required for effective electro-freezing (Tinsley et al., 2000). Cloud particles charged sufficiently to significantly modify cloud-development processes are referred to hereinafter as supercharged particles. A common method to supercharge cloud or artificial aerosol particles (e.g., water droplets produced with a sprayer) is to deploy a direct current (DC) corona discharge device producing unipolar, that is, predominantly of the same sign, air ions. The particles are then directly charged by ion attachment. Negative ions are preferred as, compared to positive ions, they have a higher mobility and, therefore, a slightly higher particle charge is achievable in the same configuration. In its general form, a DC corona discharge device, suitable for charging airborne particles with the negative sign in the socalled aerosol chargers, comprises two electrodes connected to a high-voltage direct current (HVDC) source: the cathode having one or more surface parts with a high curvature, called an emitter electrode of corona discharge, or simply emitter, and the anode with a smooth surface, called a collector electrode of corona discharge, or simply collector. For example, one or more needles with sharp pins or thin wires can be used as the emitter. In this configuration, negative air ions produced in a strong electric field around a needle, pin, or wire
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surface, drift toward the positively charged collector, thus forming an electric current passing through the air. This current is often referred to as an ionic current. Aerosol particles, carried by an air stream through the ion-drift zone between the emitter and collector, become negatively charged in this zone by direct ion attachment. These particles are then introduced (seeded) into a cloud or fog. A number of designs of aerosol particle chargers have been proposed for cloud and fog modification, for example, those described in the patent applications of Marks (1980) and Khain et al. (2003). Seeding techniques using airborne carriers or ground-based chimney-like conduits have been described by Khain et al. (2003). In practice, however, direct supercharging of cloud particles with aerosol chargers and seeding those particles into a large volume of cloudy air would need to overcome severe engineering difficulties. The average charge on a particle which can be achieved by ion attachment is approximately proportional, among other factors, to the particle size and logarithm of the so-called unipolarity factor, which is the ratio of the number concentration of the dominant-sign ions to the concentration of the opposite-sign ions. In order to supercharge cloud particles, especially small ones, the corresponding unipolarity factor should also be sufficiently high. As ions of the sign opposite to that of corona ions are always present in the air, the required unipolarity can be maintained only within a limited area of the ion-drift zone around the emitter. Therefore, only a fraction of charged aerosols would be supercharged, with the rest of them charged below the supercharging threshold. Another problem is that a strong electric field of the nearby highly charged particles may reduce the ion-production rate of corona discharge (Smith, 1972; Loveland et al., 1972), which requires the prompt removal of the charged aerosols away from the emitter. On the other hand, removal of those particles and the overall charger performance will be limited by the time required for particle supercharging. Due to the sheer size of clouds, a large number of aerosol chargers would probably be required for precipitation enhancement by seeding with charged water particles. Once removed from an ion-drift (charging) zone, particles remain supercharged for only a limited time due to their (nonequilibrium) charge decay, posing the challenging problem of distributing such an unstable seeding medium over a large volume of cloudy air within that time. This would probably require the costly deployment of a number of airborne carriers such as aircrafts or drones and/or chimney-like conduits. In practice, direct supercharging of cloud particles in sufficient amounts with aerosol chargers is difficult to achieve and costly to implement in an application on a reasonable scale, thus making it uncompetitive with conventional (chemical) cloud seeding. Therefore, a practical approach to the problem should be focused on other means of increasing the electric charge on cloud particles. One solution to the problem based on the enhancement of the natural charging processes in non-thunderstorm clouds has been found after rigorous scientific scrutiny of Russian-engineered devices for cloud and fog modification, undertaken by the research team of Meteo Systems AG, Switzerland, and supported by recent research in the field of electrical processes in cloud
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microphysics, for example, links between solar activity and climate. Since non-thunderstorm layered clouds produce a large share of the total precipitation on the Earth, they appear to be attractive targets for modification. As discussed in Section 4.05.3.3.3, the upper and lower boundaries of such clouds are sensitive to electrification by the fair-weather atmospheric electric current. The density of space charge produced by this external cloud charging is proportional to the fair-weather current density (Harrison and Ambaum, 2008). Therefore, electrical processes of precipitation formation in non-thunderstorm clouds are modulated by fair-weather atmospheric electric current and, in principle, can be influenced by engineers, provided the technical means for controlling that current are in place. The density of the (ohmic) fair-weather current is determined by local values of air conductivity and the gradient of the ionosphere-to-Earth potential of the GEC, that is, the intensity of the fair-weather electric field. The conductivity of air depends on local factors such as atmospheric ion-production rate, number concentration, and type of atmospheric aerosols. The intensity of the fair-weather electric field, although strongly correlated with global thunderstorm activity, is subject to variations due to a number of factors. At a given location, the vertical profiles of air conductivity and fair-weather electric field exhibit significant variations, while the density of the fair-weather electric current usually changes very little with altitude. Many observations have provided evidence that weather variables are strongly correlated with the fair-weather electric current. Cyclical and irregular variations in solar activity modulate ionization rates and hence the fair-weather electric current in the lower atmosphere. Recent studies, based on the observed sensitivity of weather variables to variations of solar activity, strongly indicate that the input of cosmic rays is not negligible. Cosmic rays comprise particles with a high-energy potential originating from both solar and nonsolar sources (Tinsley, 2000; Carslaw et al., 2002; Palle et al., 2004; Harrison and Ambaum, 2008). The evidence of a statistical relationship of precipitation with cosmic-ray flux was first presented by Kniveton and Todd (2001). This relationship was examined for the Beijing area by Zhao et al. (2004). Comparative analysis of heavy rainfall correlations with cosmic rays, varying with different locations around the Mediterranean basin, was provided by Mavrakis and Lykoudis (2006). In contrast to the case of a DC corona discharge where the produced ions are unipolar, natural energetic particles produce pairs of air ions of the opposite sign, a process called ‘bipolar ionization’. Due to the motion of ions driven by the fair-weather electric field through clear-to-cloudy air interfaces with high conductivity gradients, the initial microscopic charge separation by bipolar ionization results in a macroscopic charge separation by the accumulation of multiple charges of the same sign on cloud particles, negative at the bottom and positive at the top of clouds, that is, external cloud charging. As this process is scaled with the density of the ohmic fair-weather electric current, increasing charges on cloud particles could be achieved by increasing the number concentration of atmospheric ions, determined by the air ionization rate, or increasing the fair-weather electric field
strength. In theory, augmenting natural bipolar ionization by providing an artificial source of additional bipolar ionization, for example, by means of a laser beam, might be an option. Depending on altitude, about 2–15 ions s1 are naturally produced in 1 cm3 of the troposphere, the layer of the atmosphere where most precipitating clouds form. Taking into account the sheer size of clouds, simple calculations would not engender much hope for achieving an artificial ionization rate comparable to that of natural ionization. Another option is to increase the fair-weather electric field strength at cloud altitudes. This can be achieved locally by the accumulation of negative electric charges below clouds. In this configuration, the electric field of those charges points in the same direction (downward) as the fair-weather electric field, that is, the latter is augmented. The elevation height of such an electric field source above the surface of the Earth should be sufficiently high with respect to the induced image charge at the surface with a finite conductivity, which further reduces the electric field with distance from the source. In terms of electrostatics, the electric field produced by an electric field source and its image charge fades with distance as quickly as the field of electric dipole. In practice, charging atmospheric aerosol particles which will then act as charge carriers appears to be feasible. In contrast to the direct charging of cloud particles, aerosol particles should not necessarily be charged to a supercharging threshold, and this can be achieved by means of ground-based facilities. The produced plume of space charge formed by charged aerosol particles is then elevated by convective updrafts. The lifetime of space charge accumulated by aerosols is typically in the range of 15–40 min. This allows the spacecharge plumes to be elevated to altitudes of up to several hundred meters and even more, depending on the atmospheric conditions. Summarizing the above, the principle of weather-modification methods by means of unipolar air ionization at lowelevation heights above the surface of the Earth is based on a local electrical disturbance of the GEC caused by the longlasting space charge of aerosol particles acquired by the attachment of produced ions. Under certain conditions, this may lead to the enhancement of the electric field strength at altitudes of non-thunderstorm clouds and thus external charging of the latter in the GEC due to a high conductivity gradient on the clear-to-cloudy air interface on cloud boundaries. This additional artificial charging, which appears as an increase in cloud particle electric charge, negative near the bottom and positive near the top of clouds, changes electrically sensitive cloud microphysical processes, such as droplet condensation, coalescence, and freezing of cloud droplets by contact with charged aerosol particles acting as contact-freezing nuclei. In this way, a local artificial modulation is added to the GEC. This artificial modulation can be used as a base for weather-modification applications at local or regional scales.
4.05.4.5.2 Development of the technology The first experiments using corona discharge devices for fog dissipation were reportedly carried out in the Soviet Union before World War II, although a detailed description of the trialed devices cannot be found in the literature available now.
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Using charged particles of fine sand for cloud and fog modification was also considered at that time. However, as was later stated by Smirnov (1992), delivering artificial charges to clouds was a major challenge. Vonnegut et al. (1957) were the first who hypothesized the existence of a link between aerosol charging near ground level and electrification of non-thunderstorm clouds. A hypothesis describing convective cloud charging based on the physical delivery of aerosol particles, charged near ground level, into clouds was proposed. To test this hypothesis, a large open-air DC corona discharge installation was proposed to charge natural aerosol particles in large volumes of atmospheric air. In that configuration, the emitter electrode in the form of a straight wire, several kilometers long with a diameter of 0.2 mm, was elevated on masts at 10 m above the ground and powered with an HVDC source at a voltage of 10 kV relative to the ground, the latter acting as the collector electrode of corona discharge. The external cloud-charging mechanism was not understood at that time. Although the results of experiments conducted by Vonnegut’s team to prove convective cloud charging were inconclusive, the new concept of charging atmospheric aerosols with open-air DC corona-discharge installations, aiming to remotely modify the electrical state of clouds, was introduced. Vonnegut et al. (1962) performed aircraft measurements of the electric potential gradient at different altitudes above the plume of aerosols charged with a 14-km long wire. They observed that the produced space charge mixed rapidly in the lower atmosphere and caused large perturbations in the fairweather electric potential gradient, which were extended downwind 10 km or more. When there was convection, the charge was rapidly carried aloft by thermal updrafts. Bradley and Semonin (1969) carried out similar measurements and attempted to detect precipitation–modification signals with the assessment techniques available at that time. In the early 1970s, experiments on warm-fog dissipation with aerosol chargers were carried out in the USA (Loveland et al., 1972). Electrical fog-dissipation experiments resumed in Russia from the 1990s onward. Large open-air corona-discharge systems were used comprising an emitter electrode in the form of a long thin wire supported and elevated to a few meters height above the ground with one or more poles. The wire was connected to the negative electrode of an HVDC source, the positive electrode of which was earthed. In a number of proposed embodiments, the wire was arranged in different ways and different support structures used in attempts to improve the basic design of Vonnegut et al. (1958). In other embodiments, such as that shown in Figure 20, metallic collector electrodes electrically coupled with the earthed positive electrode of HVDC source were provided. Fog-dissipation experiments were conducted using various embodiments on a trial-and-error basis. According to members of one of the Russian fog-dissipation research teams, best results were achieved with embodiments where the wire was arranged in parallel segments when wound in one strand around the sides of a wooden pyramidal frame (Rostopchin et al., 2001). A number of such emitter-electrode assemblies supported by individual poles, or grouped on a rectangular
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7m
100 m
(+)
(−)
100 m Figure 20 Corona discharge installations used for fog dissipation in Russian airports.
rack and supported by multiple poles, were operated at a voltage of 50–70 kV. In peer-reviewed literature, the studies of electrical fog dissipation in Russia were presented, for example, by Afanasiev et al. (1996) and later by Chernikov and Khaikine (1999). Although the primary concern of research, technical development, and field tests on weather modification, by means of ground-base DC corona-discharge installations, was the dissipation of fog, later such installations were also used in cloud-modification experiments conducted with varying degrees of success in Russia and other countries such as the USA, Mexico, the UAE, and Australia. In Russia, the press reported a number of trials, some of them commercial, performed with ground-based corona systems to reduce rainfall during harvesting and public events, or increase it during dry seasons, in particular, for extinguishing peat and forest fires. By the end of the last century, characterized by a sharp economic downturn in the former Soviet Union, a number of Russian experts in the field of electrical weather modification, previously employed by Soviet state meteorological institutions, had reportedly joined companies based in Mexico and the USA, such as ELAT SA (Mexico City), Earthwise Technology Inc. (Dallas, TX, USA), and Ionogenics (Marblehead, MA, USA). Marketed by the competing companies under various names, the technology was referred to either as ionization of the local atmosphere (IOLA) or as electrification of the atmosphere (ELAT). Support and funding were sought from both US and Mexican governments. According to media reports, limited funding was granted to ELAT by the Mexican government for field trials in a drought-affected area in Mexico. According to a report by Moore (2004), Earthwise Technology proposed corona-discharge systems that were 7 m high, shaped like short open-topped air-traffic control towers, housing proprietary ion generators and blowers to lift the produced space charge. The ELAT installations were rather simple in appearance, consisting of a 37-m-high central tower surrounded by 8-m posts arranged hexagonally at a distance of 150 m. The tower and posts were interconnected by the
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emitter wires, electrically coupled to an HVDC source powered by a 2-kW generator. Field trials were carried out by Meteo Systems AG, Switzerland, in the United Arab Emirates (in the year 2006), and later in Australia (in 2007 and 2008) under the scientific-assessment program of the University of Queensland, Australia. During 4 months of operations, a number of major rain-enhancement events were observed and quantitatively evaluated. Performance analyses based on measurements of ionic current and the assessment of aerosol-charging efficiency were carried out for different configurations of the corona installation deployed in field tests, which was based on the pyramidal design of Rostopchin et al. (2001). The major achievement of Meteo System’s research team was discovering the basic principles of weather modification by unipolar aerosol charging with ground-based systems. This opened the way for Meteo Systems to engineer a number of high-performance installations of the next generation, which are optimized for particular weathermodification applications. This would have been impossible without understanding the physical processes upon which the technology is based. Until recently, the mechanisms of how the space charge introduced into the lower atmosphere affected cloud development were not well understood. Field experiments relied on a trial-and-error approach. The prevailing belief, inherited from the former hypothesis of convective cloud charging, was that space charge should be delivered into a cloud in order to cause electrical modification of the latter. This and other hypotheses, such as those based on the vertical ion transport to clouds, were strongly criticized by the scientific community, and any observed evidence was not accepted but questioned, with reference to natural weather variability which is still not reliably predictable. However, as List (2005) stated, this reservation applies to any weather-modification technology regardless of the methods and techniques used.
4.05.4.6 Evaluation and Recommendations Although a large number of experimental trials have been carried out over the past years using ground-based coronadischarge installations, there is not a single report available in the scientific literature containing detailed results of the trials, and practical experiences. To the knowledge of the authors, only one paper has been published which provides, in the appendix section, some general information about trials that were planned for Mexico and Webb County, TX, USA, from 1996 to 2002 (Kauffman and Ruiz-Columbien, 2005). In contrast to strategies adopted by companies involved in marketing cloud-seeding technologies, the proponents of ionization-based technologies were very protective in their activities. As a result, some elements within the scientific community see the electrical weather-modification technology as voodoo science (Park, 2000). Subsequently, potential clients (e.g., representatives of water authorities) approach the technology with suspicion even when water scarcity is severely threatening local people and the agricultural community. The review of the scientific fundamentals of the ionizationbased technology presented in Section 4.05.4.5.1 reveals that this technology is far from being scientifically untenable. When applied in a scientifically sound manner, the technology
exhibits a significant potential, even higher than that of cloud seeding. During cloud-seeding operations, only a fraction of the available cloud cover can be influenced. In contrast, clouds covering a significantly larger area can be modified by remote cloud charging at low cost, especially if a grid of multiple ground-based installations is deployed. Moreover, the technology is highly scalable and suitable for applications which cannot be implemented or are difficult to implement by cloud seeding. In summary, the ionization-based technology, if accepted by the scientific community and supported in more detail with scientifically sound knowledge, has enormous potential to become a viable and widely used weather-modification technology. As in the case of cloud seeding, acceptance of the ionization technology will come with sound evaluations of achieved success. The chosen method of evaluation should not necessarily be based on randomized trials. As discussed in Section 4.05.4.4.2, application of the method is likely to be successful when done under the right meteorological conditions. In particular, the presence of updrafts is essential for vertical plume transport and thus for the success of operations. The art of applying the technology successfully depends on knowledge and correct interpretation of actual meteorological conditions in the three-dimensional space around the ion emitter, taking into account the variation of these conditions over time. Therefore, monitoring meteorological conditions, data collection and evaluation, with the aid of mathematical models, and operation of the technology according to the results of data evaluation are the three pillars on which any success rests. Applying the technology in such a manner is highly recommended. Unipolar ion emission should not be confused with the emission of electromagnetic waves, which are suspected of posing some health hazards. Nevertheless, the question whether or not the ionization-based technology is environmentally friendly needs to be scrutinized. It is highly unlikely that the emitted ions impose any threat to plants, animals, or commercial applications such as air traffic. Indirectly, health risks cannot be excluded, however. Corona discharge may cause generation of hazardous gases such as nitrogen oxides and ozone. The latter is known to be a strong oxidant, which may react with atmospheric pollutants. In a worst-case scenario, oxidation reactions may be incomplete, leaving molecules behind which may be toxic or even carcinogenic. It is highly recommended to invest in research to minimize the release of hazardous gases, for example, by optimizing the operating regime of corona discharge, and to clarify the extent of incomplete oxidation and the resulting toxic residues prior to commercial applications of the technology. As mentioned in Section 4.05.4.4.2, weather-modification technologies alone cannot solve the water-supply problems on the Earth unless the technology is treated a part of IWRM. In the case of the ionization-based technology, integration into the general policy of regional water management is particularly important since the methods affect rather large areas. At the ground, conflicts between different interest groups are very likely to develop, for instance, conflicts between agriculture which needs rainfall, and the tourist industry which wants clear skies. Political conflicts may arise between states when the technology is applied in areas close to borders.
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The question about which authority owns the atmospheric humidity requires international regulations, presumably at the United Nations (UN) level. Last but not the least, regulations need to be set with respect to liability. In the case of heavy rainfall caused by the application of any weather-modification technology, damages and even accidents may occur. In this case, is it the authority which ordered application of the technology, or the company which operated the technology that is legally responsible? Agreements at state level and with the major insurance and re-insurance companies need to be settled prior to commercial applications.
4.05.5 Rainwater Collection, Purification, and Storage 4.05.5.1 Incentives for Action As already mentioned in Section 4.05.2, there is enough water available in the atmosphere to support life on the Earth, and to satisfy the demands of people, agriculture, and industry, today and, most probably, in the future. The water contained in the atmosphere is transferred to the surface of the Earth by means of natural or deliberately forced precipitation processes. Once on the ground, the water may evaporate, be intercepted and used by plants, penetrate into the groundwater table, or form surface water bodies such as wetlands, lakes, creeks, or rivers (Figure 3). To minimize evaporation losses, the water can be collected on the spot, infiltrated, and stored in underground aquifers for subsequent use, or collected and stored in cisterns, ponds, or dams. One serious problem is that, on the local scale, precipitation in the form of rain or snow does not necessarily correspond to the actual demands of nature and humans – neither in time nor in intensity. There are areas which regularly receive high volumes of precipitation, far more than is actually needed. The west coast of the south island of New Zealand is a good example of the too-much-rain dilemma. In the area of Hokitika, for instance, annual rainfall exceeds 2800 mm. On the other hand, there are areas where rainfall is rare but water demand is extremely high. Southern California, where the annual rainfall is less than 40 mm, is an example of this extreme condition. Worldwide, people have tended to settle in areas where sunny days are predominant. Unfortunately, such areas are frequently affected by water shortages (e.g., Dubai), necessitating huge technical and financial efforts to satisfy urban and peri-urban water demand. Likewise, agricultural production is concentrated in sunny areas where solar radiation allows highquality grains, vegetables, and fruit to grow, provided enough water is available from local or distant sources. Water shortages may even develop in wet countries when the populations of cities and the resulting water demand exceed the capacity of rainfed water resources. The problems of the London metropolitan area are an example of this.
4.05.5.2 Rainwater Collection As people need water in sufficient quantities year round, collecting rainfall and temporary storage has been common practice in arid and semiarid countries ever since humankind
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shifted from migrating, gathering, and hunting to settlements and farming. In areas notoriously exposed to long-lasting periods of dry weather, people have learned to survive by collecting rainwater from roofs, and by storing the water in cisterns placed above or below ground. This traditional method of collecting and storing rainwater is called rainwater harvesting. With the advent of modern water technologies people have lost interest in rainwater harvesting, even in areas not served with tap water. This was certainly an important factor in the improvement of water supply in poorly served areas when organizations such as the Centre of Science and Environment (CES), based in New Delhi, India, started to revitalize the concept of rainwater harvesting. In 2004, the achievements of CES were rewarded with the Stockholm Water Prize. Since then, this old-fashioned but life-securing method has begun to receive worldwide attention, again. To help people understand that the perils resulting from lack of water can be mitigated by taking individual initiatives, CES produced a video which depicts a man in distress as rain starts pouring and a gust of wind blows away his umbrella. Eventually, the umbrella lands upside down on the road and fills quickly with rainwater. This is observed by passersby who excitedly start collecting rainwater in whatever container is available, including a police officer’s helmet. This video was an eye-opener for many people in the world who suddenly realized that, besides high-tech solutions, there are simple ways of coping during droughts. In Australia, for instance, under the pressure of a 6-year drought, a modern version of rainwater harvesting has become common practice (Lancaster, 2005). Large-scale rainwater-collection plants could be built, based on the upside-down umbrella metaphor, using devices which may only be opened in the case of a rain event. The advantage would be that pollution of the collected water by deposits, and loss of water through evaporation, could be minimized. The collected water could then be diverted to a storage tank to protect it against quality deterioration caused by algae growth, for example, as depicted in Figure 21.
4.05.5.3 Pollution and Purification of Stormwater Runoff Although the collection and storage of rainwater is certainly a clever method to deal with water shortages, it is not without drawbacks. Rainwater is by nature low in mineral-salt content. When consumed in large quantities, it may affect the osmotic pressure at a cellular level, resulting in health risks. More severe health risks result from pollution picked up by the raindrops when passing through the atmosphere, and by the rainwater once it comes into contact with the collection surfaces (roofs, terraces, courtyards, roads, etc). Analyses performed by Wallinder et al. (1998), Athanasiadis (2005), Schriever (2007), and Helmreich (2009), and by many others have revealed that rainwater collected from roofs and roads is severely polluted in rural areas, and to a much larger extent in cities (Fuchs et al., 2002). The collected water may contain soluble and particulate as well as dissolved materials. Airborne pollutants such as SO2, NOx, NH4, and volatile organic substances (VOSs) are of particular concern, as is
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Figure 21 Upside-down-umbrella farm to collect rainwater in arid areas: a visionary approach. Collecting devices open when it rains (right), and fold up during dry weather conditions (left) to allow free access to the native land.
small particulate matter (respirable dust). A multitude of organic and inorganic substances may be picked up from roofs and roads; some of these are dissolved in others in particulate form, for example, leaves and droppings of birds, cats, and other animals. Some pollutants may have deposited on the surface during dry periods. During the rain event, they get washed away and transferred into the collected rainwater. The resulting concentration of pollutants is especially high at the beginning of the rain event (first flush). Of particular concern are abraded materials from tires, brakes, catalytic converters in cars, and road covers because these materials contain heavy metals such as zinc, copper, cadmium, chromium, and platinum. Mangani et al. (2005) measured, during first-flush situations, concentration values of 346 mg l1 for copper, 412 mg l1 for zinc, and 37 mg l1 for lead. Organic pollutants of concern are benzene, poly-aromatic carbon (PAC), methyl- and ethyl-tert-butyl-ether (MTBE and ETBE, respectively). In Munich, Germany, the mean concentration of dissolved organic carbon (DOC), as a sum parameter for all those substances, was monitored over a period of 2 years. The concentration of DOC was in the range of 20 mg l1 in the runoff from a heavily frequented highway (Helmreich, 2009). The total organic carbon (TOC) was 70 mg l1 on average, and the total suspended solids was more than 350 mg l1. With respect to runoff from metal roofs, particularly high concentrations of copper and zinc, as a result of corrosion and subsequent wash-off effects, were detected by Athanasiadis (2005) and Schriever (2007). The wash-off rate for copper varied between 0.7 and almost 2 g m2 a1. For zinc, a mean value of 3.7 g m2 a1 was measured (Helmreich, 2009). In some cases, the concentration of zinc exceeded the 30 mg l1 margin. Copper concentrations in the runoff from a copper roof varied between 0.4 and 11 mg l1.
4.05.5.4 Purification of Stormwater Runoff in Decentralized Treatment Units Rainwater collected from roofs and roads can be considered as a supplementary source of water for households, industry, and agriculture, provided the collected rainwater is properly treated. Particulate material as well as dissolved pollutants need to
Figure 22 Buried road runoff treatment plant next to a gully of a highway in Ishijama, close to Lake Biwa, Japan.
be removed or, at the very least, lowered in concentration. If the treated water is designated for human consumption, disinfection is necessary to achieve hygienic safety. Since the extent of water demand and the availability of runoff hardly match actual water needs, temporary storage in containers (cisterns), ponds, or dams has to be provided. When the stormwater runoff is collected from a large area, the water is sent to a central treatment plant for purification. Such plants are large in size, and the treatment units often are placed in a standby phase as these can only be operated when rainfall actually occurs. Matsui et al. (2001) developed a decentralized concept for road runoff treatment and infiltration of the treated water into an aquifer. The pilot unit was positioned next to an individual street gully (Figure 22). Such treatment units are small since the volumetric loading remains comparably low even under peak-flow conditions. The treatment could be limited to
Abstraction of Atmospheric Humidity
filtration, absorption, and ion-exchange processes as barriers against the transfer of pollutants into the subsoil and the aquifer (Matsui and Lee, 2003). Filtration into the porous subsoil would take care of the removal of pathogenic organisms. The above idea was adopted by Koenig (1999) who proposed to pass the collected roof runoff through a layer of biologically activated top soil for filtration and biodegradation, and store the water in a buried cistern made out of prefabricated concrete for further use or infiltration into the aquifer (Figure 23). The system was commercialized and is used at various locations in Germany. The individual units are designed to receive runoff from roofs, 150 or 300 m2 in size. The prefabricated units are available in different volumetric sizes (3–10 m3). The total depth of the units varies between 2.4 and 3.1 m.
Soil layer consisting of special substrate
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Boller and his team developed a special filtration method which combines solid–liquid separation and the binding of heavy metals to granular iron-hydroxide (Boller and Steiner, 2002; Steiner and Boller, 2006). Advanced scientific investigation followed by field trials was conducted in Munich, Germany, by Athanasiadis (2005), Athanasiadis and Helmreich (2005), Athanasiadis et al. (2006), Hilliges (2007), and Helmreich (2009). In Munich, chemically conditioned clinoptilolite was selected as an ion exchanger to remove dissolved heavy metals from roof runoff (Athanasiadis and Helmreich, 2005). The treated clinoptilolite is packed in a filter column with the intention to remove, in addition, heavy metals in particulate form by filtration processes. Figure 24 shows the experimental setup which, after extensive testing, was applied at full scale at the bottom of a large copper-roofed building in Munich
Option: inlet
Inlet or overflow Manhole with cover Drainage pipe with coconut-fiber sleeve Empty conduit to domestic water station Monolithic container
Outlet, e.g., Mall percolation box
Figure 23 Rainwater storage reservoir with a soil filter top for purification of the collected roof runoff.
Emergency overflow
Effluent
Packing
Influent
Hydro-cyclone
Sedimentremoval pipe Figure 24 Schematic of the treatment system used to purify stormwater runoff from a copper-roofed building in Munich, Germany, according to Helmreich B (2009) Stoffliche Betrachtung der dezentralen Niederschlagswasserbehandlung (Pollution of stormwater runoff to be treated in decentralized system). Berichte aus Siedlungswasserwirtschaft. Munich, Germany: TU.
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Copper roof copper already patinated
Clinoptilolite filter cage
Manhole cover Figure 25 Copper-roofed building in Munich, and a view into the treatment chamber with the casket holding the granular clinoptilolite filter material.
(Figure 25). The total roof surface area is 4800 m2. Ten treatment units were installed but only four of them, each serving a roof area of 500 m2, were monitored. The influent was first introduced into a hydro-cyclone chamber to remove particulate material. The filtration chamber was packed with 750 kg of conditioned clinoptilolite, grain size 1–2 mm. For conditioning, the sieved clinoptilolite particles were submerged in a 1 M NaCl solution at room temperature for 24 h. Afterward, the treated material was washed 3 times with deionized water. During exposure to NaCl, the ion-binding sites were expected to be occupied by sodium ions, and it was assumed that sodium ions were readily exchanged by heavy metal ions during the purification process. During a 1-year observation period, 20 rainfall events were monitored. Samples of the rainwater before and after coming into contact with the copper roof, and of the effluent of the filter, were collected. The total volumetric loading of the four units was 744 m3. The copper concentration in the influent of the filter units varied between 38 and 980 mg l1. The effluent concentration varied between 19 and 84 mg l1. Considering mass loading, copper could be retained by 97%. Comparable removal values were observed for zinc. Similar experiments were performed with the aim of purifying runoff of a busy main road in downtown Munich, Germany. On average, 57 000 motor vehicles use this road every day. The catchment area of the filter system was about 300 m3. The composition of the road runoff differs significantly from the runoff of the copper roof discussed above. It contains a much higher load of particulate material, both inorganic and
organic. To remove large and heavy particles at the curb site, a simple and easily cleanable sedimentation/filtration trench was installed, followed by a hydro-cyclone placed in the lower part of the treatment unit (Figure 26). Particles of smaller size and lower density were removed by means of a fine sieve with a mesh size of 0.7 mm. Finally, the water was forced to pass through a medium consisting of a low-cost granular activated carbon material (1–2.5 mm in size) based on brown coal, before it was sent to a seepage trench. The experimental unit was monitored for 6 months. During this period, 63 major rainfall events were observed. The concentration of the various organic and inorganic pollutants varied across a wide range (DOC between 3.6 and 80 mg l1, PAC up to 1.3 mg l1). During winter, a significant increase in salt concentration was measured. The sodium concentration varied between 17 and 10 400 mg l1. The median concentration of zinc was 1 mg l1. The copper concentration varied between less than 0.1 and 0.6 mg l1. During the observation period, the removal rate of all organic pollutants and heavy metals was in the range of 90– 95%. The filter system installed along the curb eliminated firstflush effects almost entirely. Enhanced salt concentration during winter did not affect treatment efficiency in any significant way.
4.05.5.5 Large-Scale Storage of the Collected Rainwater Rainwater collected from large areas is traditionally stored in reservoirs and dams open to the atmosphere and to sunlight. Here, the problem is that some of the collected water is lost to
Abstraction of Atmospheric Humidity
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Manhole Handle to remove the fine sieve for cleaning Effluent
Packing Fine sieve Influent
Hydro-cyclone
Sediment trap Figure 26 Schematic representation of the treatment system used to purify road runoff, according to Helmreich B (2009) Stoffliche Betrachtung der dezentralen Niederschlagswasserbehandlung (Pollution of stormwater runoff to be treated in decentralized system). Berichte aus Siedlungswasserwirtschaft. Munich, Germany: TU.
evaporation. A wide variety of pollutants, pathogenic organisms among them, are transmitted from the atmosphere and from the surrounding land to the water stored in reservoirs. Moreover, algae growth leads to deterioration of the water quality as it transmits metabolites into the water, and also via biological degradation processes which follow the decay of algae. To keep the collected water from deteriorating in quality and quantity, infiltration into underground storage compartments is advisable. This process is commonly termed ‘groundwater recharge’. Strobl and his colleagues (Strobl and Zunic, 2006) developed a so-called infiltration dam concept (Figure 27). Infiltration dams serve two basic functions: as barriers preventing flooding of downstream areas and as reservoirs for the recharge of downstream aquifers. Stormwater runoff from upstream areas is collected and temporarily stored in the reservoir. To enable controlled groundwater recharge, the effluent of the reservoir is introduced to the downstream wadis (a dry riverbed that contains water only during times of heavy rain) for percolation (Figures 28 and 29). The flow is regulated so that evaporation losses are kept to a minimum. The water introduced in the wadis infiltrates toward the aquifer under the force of gravity. By raising the groundwater table, saltwater intrusion from the ocean can be counteracted. Thus, groundwater recharge can be considered as a measure to control salinization of aquifers. During the time the water resides in the aquifer, chemical interactions may take place with rock, gravel, and sand materials whereby the water picks up minerals. Filtration processes take place as water passes through sand and gravel layers. Both of these processes in combination are known to contribute to a significant improvement of water quality. Most likely, the water can be considered hygienically safe, and can be used as drinking water, and also for irrigation of gardens and agricultural fields. Full-scale trails carried out in Wadi Ahin, northern part of the Sultanate of Oman, combined with numerical modeling
(Haimerl et al., 2002; Haimerl, 2004; Strobl and Zunic, 2006) demonstrated the feasibility of this innovative approach. During a case study conducted in 1996, almost 4 million m3 of water was collected and temporarily stored in the reservoir. The stored water was discharged into the wadi system at a rate of 3 m3 s1. The efficiency of groundwater recharge increased with the moisture content of soil, with the water level of the surface flow, and the time of infiltration. These are obviously the aspects to be considered in efforts to further improve the efficacy of the technology. During the observation period, 85% of the water accumulated in the reservoir was able to be transferred to the aquifer. Evaporation losses were only around 0.1%, and 2.5% was captured in the topsoil.
4.05.6 Overarching Aspects Traditionally, water for human, industrial, and agricultural consumption is abstracted from rainfed sources such as rivers, lakes, and aquifers. After purification and transportation to customers, the water is used for specific purposes such as drinking, cleaning, and irrigation. Some of the water evaporates and thus gets incorporated in the original source, that is, atmospheric vapor. Most of the used water, however, is discharged into natural water bodies such as rivers, is transported downstream, and, in many cases, is abstracted again for subsequent use. In such cases, we are talking about unintended water reuse. As discharge of polluted water in rivers and lakes poses threats to both aquatic organisms and downstream users, and because such threats disturb the functioning of the aquatic environment as well as downstream economies, efforts have been made to regulate the abstraction of water from the Earthbased resources, and discharge of the used water back into the aquatic environment. Over recent decades, it has been realized that water management on the level of entire river basins is necessary to secure the economic, ecological, and social development of regions – in short: sustainable development.
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Rain in the mountains
Wells
Reacharge dam Flood detention basin
Surface runoff in wadi channels
Sea
Aquifer
Seawater intrusion into aquifier
Figure 27 Schematic representation of the concept of infiltration dam concept, provided by Zunic, Institute of Water and Environment, TUM, Germany.
Wadi inflow Reservoir Dam Outlets to wadi channels
Rock
Groundwater
Figure 28 Discharge of the water from the reservoir into wadi channels for subsequent percolation toward the aquifer. Schematic provided by Zunic, Institute of Water and Environment, TUM, Germany.
The United Nations Declaration on Environment and Development (Anonymous, 1972) may be considered as the starting point of what is now called the IWRM concept. IWRM is an alternative to the dominant sector-by-sector, top-down water management of the past. Water resources are
now managed at the basin or watershed level. The watersupply side is taken into account simultaneously with the water-demand side. IWRM integrates the use of land, groundwater, surface water, and coastal water. It integrates the interests of upstream and downstream regions. An
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Figure 29 Wadi channels in the Wadi Ahin area, Oman. The wedded riparian zone indicated progress of infiltration. Photo provided by Zunic, Institute of Water and Environment, TUM, Germany.
intersectoral approach is taken to decision making. IWRM calls for the integration of policy, regulatory, and institutional frameworks, such as the implementation of the polluter-pays principle, water-quality norms and standards, and marketbased regulatory mechanisms. According to the Global Water Partnership’s definition, 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 an equitable manner without compromising the sustainability of vital ecosystems (Anonymous, 2000; Anderson et al., 2008). The European Water Framework Directive (WFP) implemented in the European Union in October 2000 (Kaika, 2003) is an example of IWRM application. It calls for legal responsibility for the quantitative and qualitative status of all water bodies in Europe, including marine waters. It is a framework in the sense that all member states are committed to taking appropriate action to reach the common goal of achieving and maintaining high ecological quality of all surface and groundwater bodies. So far, however, atmospheric humidity and its quality are not included in the WFP, although they should be. Similar legislation has been implemented in many other parts of the world, including in developing countries. In Kenya, for instance, a Water Resource Management Authority (WRMA) has been established. It is a state corporation under the Ministry of Water and Irrigation, established under the Water Act 2002, and charged with being the lead agency in water-resources management (Anonymous, 2004). With the advent of methods designed to make active use of atmospheric water as an alternative water resource, a new dimension is added to the IWRM concept. Some of these methods are described in Section 4.05.4, and are comprehensively categorized in Figure 30. Other methods may be
developed in the very near future in response to the steady increase of water shortages around the world, caused by factors such as climate change, population growth, growth of urban areas, and lifestyle changes. A variety of options are available when it comes to the question of how to deal with the water abstracted from the atmosphere. Some of these options are shown in Figure 30. Precipitation generated by cloud seeding or ionizationbased technologies may be directed toward agricultural fields for sustained crop growth. In this case, farmers would be expected to pay for the service. Alternatively, it could be directed toward forest areas to prevent the outbreak of fires and the subsequent loss of property value and biodiversity. In this case, governmental organizations or insurance companies may be obliged to cover costs. To meet diurnal, weekly, and seasonal variations in water demand, temporary storage is certainly an option to be considered. The size of the storage facilities may vary from a small tank up to a large dam. Atmospheric water is low in salt content. When human consumption is concerned, mineralization of the water may need to be considered. When coming into contact with various surfaces, the water may pick up pollutants, dirt, droppings of animals, and even heavy metals (see Section 4.05.5.3). While being stored in a reservoir, the water may deteriorate in quality. In summary, the water needs to be treated prior to delivery to consumers in order to secure the health and welfare of the customers. The costs of abstracting, holding, treating, and delivering the water should be covered by the consumers (domestic and industry) based on volumetric consumption. This option is still a controversial subject in many regions of the world, however. Surprisingly, methods of abstraction of atmospheric humidity have not been taken into account by regulating
Abstraction of Atmospheric Humidity
Rainwater harvesting
Fog collection
Heat island
Clud seeding
Ionization
No regulations, yet
Condensation
Source: water vapor
Cloud formation and precipitation
Collection and storage
Piping or bottling
Purifying
Metering and prizing
Consumption Options:
S
S
S
S
S
Regulations exist and are mostly enforced
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People − Industry − Agriculture/forestry − Groundwater recharge Natural precipitation
Engineered capturing
Delivery
Figure 30 Overview of technologies to abstract atmospheric humidity, and the various options of treating the captured water on its way to the customers (people, industry, and agriculture).
authorities around the world, even though methods such as cloud seeding have been in use since 1946 (see Section 4.05.4.4). It is also surprising that atmospheric water has no obvious ownership – at least to the best knowledge of the authors (Wilderer, 2009). Influencing the atmosphere with the aim of changing weather conditions is not regulated anywhere in the world, with the exception of weather modification for hostile, military purposes (Anonymous, 1976). Over the centuries weather conditions have very often been decisive in military confrontations (Durschmied, 2000); therefore, it is only to be expected that engineered weather modification with the aim of abstracting atmospheric humidity to the advantage of one party or another may engender conflicts among stakeholders, regions, and even state authorities. Thus, it is high time to enter into national and supranational agreements concerning the exploitation of atmospheric humidity. Moreover, it is necessary to clarify liability issues, particularly regarding insurance. It has to be made clear as to which authority has the power to decide when and where certain weather-modification actions are allowed to be conducted. The authority in charge would then be responsible in case the activity has unintended extreme consequences, for instance, flooding, release of avalanches or mudslides, or car accidents.
Consequently, insurance companies need to be prepared to cover damages of any kind in case things do not go according to plan. A discussion of this type may appear superficial when considering technologies which obviously have only a very local effect, such as condensation or fog-collection technologies. In principle, however, the international community and governments everywhere would be well advised to take the negative effects of humidity abstraction seriously before a major accident occurs. Other than under natural conditions, it will not be force majeure which causes misery, but deliberate man-made actions.
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Abstraction of Atmospheric Humidity
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4.06 Safe Sanitation in Low Economic Development Areas BJ Cisneros, Universidad Nacional Auto´noma de Me´xico, Coyoaca´n, Mexico & 2011 Elsevier B.V. All rights reserved.
4.06.1 4.06.2 4.06.3 4.06.3.1 4.06.3.1.1 4.06.3.1.2 4.06.3.1.3 4.06.4 4.06.4.1 4.06.4.1.1 4.06.4.1.2 4.06.4.1.3 4.06.4.1.4 4.06.4.1.5 4.06.4.1.6 4.06.4.1.7 4.06.4.2 4.06.4.3 4.06.4.4 4.06.5 4.06.5.1 4.06.5.2 4.06.5.3 4.06.5.3.1 4.06.5.3.2 4.06.5.3.3 4.06.5.3.4 4.06.5.3.5 4.06.6 4.06.6.1 4.06.6.1.1 4.06.6.1.2 4.06.6.1.3 4.06.6.1.4 4.06.6.1.5 4.06.6.1.6 4.06.6.2 4.06.6.2.1 4.06.6.2.2 4.06.6.3 4.06.6.4 4.06.6.4.1 4.06.6.4.2 4.06.6.4.3 4.06.6.5 4.06.6.5.1 4.06.6.5.2 4.06.6.5.3 4.06.6.5.4 4.06.6.5.5 4.06.6.5.6 4.06.6.5.7 4.06.6.5.8 4.06.6.5.9
Introduction Historical Background Sanitation as Part of The Hydrological Cycle or Properly Closing the Water Loop Sources of Pollution Municipal discharges Industrial discharges Nonpoint and nonconventional pollutant sources to water Pollutants Biological Pollutants Viruses Bacteria Protozoa Helminth eggs Biological indicators Emerging pathogens Biological analytical techniques Conventional Parameters Emerging Pollutants Risks Sanitation in Low-Income Countries: A Complex Current Situation Sanitation Needs a Definition Millennium Development Goals Present Situation General overview Regional situation Situation at the national level Low-income countries sanitation specificities Sanitation Costs Wastewater Management Systems Basic Sanitation Facilities Traditional latrines Ventilated improved pit latrine Septic tank Composting toilets Pour-flush toilets Additional recommendations to set up basic sanitation facilities Toilets Water-saving toilets Toilets not using water Sludge Extraction from On-Site Sanitation System Sewerage Systems Small sewers Conventional sewers Pluvial sewers Wastewater Treatment Conventional pollutants treatment Pathogens treatment Emerging chemical pollutants Slow filtration Waste stabilization ponds Wetlands Land treatment Reservoirs and water storage tanks Upflow anaerobic sludge blanket
147 147 148 148 148 148 148 149 149 149 150 150 151 154 155 155 157 159 161 161 161 161 162 162 162 162 162 163 164 164 164 165 165 165 171 172 173 173 173 173 173 173 174 174 174 175 175 175 175 175 176 176 176 178
147
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Safe Sanitation in Low Economic Development Areas
4.06.6.5.10 4.06.6.5.11 4.06.6.5.12 4.06.6.5.13 4.06.6.6 4.06.6.7 4.06.7 4.06.7.1 4.06.7.1.1 4.06.7.1.2 4.06.7.1.3 4.06.7.2 4.06.7.3 4.06.7.3.1 4.06.7.3.2 4.06.7.3.3 4.06.7.3.4 4.06.8 4.06.9 4.06.9.1 4.06.9.2 4.06.9.2.1 4.06.9.2.2 4.06.9.2.3 4.06.10 4.06.10.1 4.06.10.2 4.06.10.3 4.06.10.4 4.06.11 4.06.12 References
Activated sludge Coagulation–flocculation Rapid filtration Disinfection Sanitation and Wastewater Treatment Costs Criteria for Selecting Wastewater Treatment Processes Wastewater Disposal versus Reintegration Soil Disposal or Reintegration of Used Water to Soil and to Groundwater Leach drains Evapotranspiration beds Soil aquifer treatment and aquifer storage recovery system Disposal into Surface Water Bodies or Reintegration of Used Water to Surface Water Bodies Reuse Types of water reuse Unintentional reuse Intentional or planned reuse Graywater reuse Sludge and Excreta Management Policy Integrated Water Resources Management Need for an Own Policy for Developing Countries Issues to address Challenges to face Strategies that can be used Funding Funding Options Why Sanitation Needs to be a Public Process Why Private Participation can be Involved Differences between Low- and Middle-Income Countries Science and Innovation: Need to Develop Individual Knowledge Conclusions
4.06.1 Introduction Before reading this chapter, it should be considered whether it is justifiable to have a specific section dealing with sanitation for low economic development areas (developing countries). Evidently, the editors of this book think so. The reasons include
• •
an increasing evidence that wastewater quality in high and low economic areas is different regarding some parameters that determine treatment options and differences in economic conditions necessitate alternative solutions not only at the technical level but also in terms of the ways to implement them.
To protect health, raise the quality of life, and increase the economic level, a good sanitation service is required in developing countries. While in developed countries, sanitation coverage is almost 99% as a result of a clear commitment of governments to provide it as part of the public services, in developing ones it is only around 50% (WHO–UNICEF, 2006). In addition, in the developed countries, the term sanitation applies not only to the installation of sewers but also to the full implementation of systems for the safe disposal and reuse of treated wastewater, sludge, and septage. In contrast, in developing countries, the term sanitation mostly
179 179 179 179 180 180 180 180 180 180 180 181 181 181 181 183 188 188 189 189 189 190 190 190 193 193 194 194 195 195 196 197
applies to the use of sewers not always ending in treatment plants. In fact, reported sanitation figures frequently do not reveal the disposal of wastewater or excreta uncontrolled into the environment, the existence of malfunctioning wastewater treatment plants, or the use of rudimentary and inefficient basic sanitation facilities sometimes contributing to increased environmental pollution rather than to control it. As a result, waterborne diseases affect millions of people in the developing world, and the water quality of surface and groundwater bodies is increasingly deteriorating. The aim of this chapter is to assist the process of increasing sanitation in low-income regions by contrasting the differences in needs and solutions’ options with high-income regions. Most technical publications have traditionally grouped developing countries together as low-income societies without considering that in them there are high- and low-income areas and that among the latter ones there are several factors that create differences that need to be taken into consideration to provide suitable solutions, that rarely fall under the logic used in developed countries to provide sanitation. Most people lacking sanitation include the millions of poor people (Figure 1) living under precarious institutional conditions and under an economical and social situation that avoids the use of conventional solutions. This renders the provision of sanitation in low-income areas a major challenge.
Safe Sanitation in Low Economic Development Areas Living below the poverty line
Living above the poverty line
149
Total
2000
Million people
1500
1000
500
0 Low-income countries
Middle-income countries
Total
Figure 1 Poverty distribution of the global population without access to basic sanitation in low- and middle-income countries (with information from Lenghton et al. (2005).
4.06.2 Historical Background The history of sanitation is mainly about three aspects: toilets, sewers, and final disposal. As sanitation is a broken subject in developing countries, the story of these three is also the same. When mankind was nomadic and lived in very small communities, sanitation was not an issue. Nature could absorb human wastes. Later, when villages grew, there was the need to set up special practices and facilities. In ancient Egypt (B3000 BC), each household had the responsibility to dispose of their garbage and excreta at the communal dump, in irrigation canals, or in open fields. Irrigation canals were the first drainage and waste disposal systems. At that time, toilets were a luxury that only the wealthier people could afford in cities. Toilets were carved of limestone, and the used water was disposed of into pits in the streets (MSU, 2009). Flushing toilets – some of them communal – existed in India since the twenty-sixth century BC. Reports on the use of toilets and other safe sanitation practices in ancient civilizations from Asia, Latin America, and Africa were common in places where nowadays lack of sanitation is a problem. The earliest covered sewers reported are from the Indus Civilization (2600–1900 BC) where Pakistan is located today. Cities used sewers to control inundations caused by pluvial water. The Cloaca Maxima or Roman sewer dates from around 600 BC. Initially, it was an open drain that was covered and left below the urban level, as the city building space became costly (Wikipedia, 2009). Later, when water began to be supplied in large quantities to households, getting rid of the used water became a problem and water was considered as a waste. It was then when sewers were found to be a useful infrastructure to convey wastewater out of the city in addition to stormwater. Concerning disposal, land application of wastewater and excreta has a long tradition in many countries. For centuries, farmers in China used human and animal excreta as fertilizers. The oldest references to the use of excreta in aquaculture come from some Asian countries, where it was employed to increase fish production (WHO, 2006). Further, even now in China, Mexico, Peru, Egypt, Lebanon, Morocco, India, and Vietnam
wastewater is used as a source of crop nutrients (Jime´nez and Asano, 2008). According to Rusong (2001), in contrast to the ‘mechanical’ ideas predominant in industrial societies, human ecological thoughts in ancient China emphasized the use of systems advocating ‘man and nature as one’. This principle is considered as equivalent to the sustainability principle and is based on terms describing concepts that are dissociated in modern civilizations, such as
• • • • •
Tian – heaven or nature; Di – Earth or resources; Ren – people or society; Wuxing – the five fundamental elements and movements within any ecosystem, that need to be in equilibrium by promoting and restraining each other; and Zhong Yong – describing that things should never go to their extremes but should be kept at equilibrium.
For several centuries, based on these ecological principles, China has developed and supported 21% of the world’s population with only 7% of the world’s arable land and less than 7% of the world freshwater resources (Rusong, 2001). Once again, similar conceptions can be found in ancient civilizations from Asia, Africa, and Latin America, in the same places where there are environmental crises now.
4.06.3 Sanitation as Part of The Hydrological Cycle or Properly Closing the Water Loop The urban water cycle is a relatively new concept used to analyze water quality problems in cities (Jime´nez, 2009b), which is depicted in Figure 2. It is useful in identifying conventional and nonconventional sources of pollution, in particular those that are specific to developing countries. It is important to understand the difference in order to be able to apply proper solutions to sanitation that go beyond the simplistic approach of merely installing wastewater treatment plants. A similar analysis could be made for rural areas.
150
Safe Sanitation in Low Economic Development Areas
Water from: Rivers Lakes Reservoirs Water distribution through the network.
Water treatment plant Possibly
Wastewater treatment plant Disposal Occassionally
Rain Air pollution
Water tanks, water vendors, water bottlers
Urban agriculture with wastewater
Infiltration from Water network Ground water
Polluted rivers or open wastewater drains
Rivers lakes reservoirs
To the next city
Agriculture
Considerably higher
sewerage septic tanks industries storage tanks
Irrigation
Infiltration
Aquifer
Figure 2 Hydrological urban cycle. Differences compared to developed countries are shown in red. Adapted from Jime´nez B (2003) Health risks in aquifer recharge with recycle water. In: Aertgeerts R and Angelakis A (eds.) State of the Art Report Health Risk in Aquifer Recharge Using Reclaimed Water, pp. 54–172. Rome: WHO Regional Office for Europe.
The urban water cycle is important because of the large increase in urban population that is being experienced worldwide. By 2030, the urban proportion of the global population is expected to be around 60%. Over the next 50 years, in developing countries, most of the population growth will occur in urban and periurban areas. Furthermore, most of the 19 cities with the most rapid growth are located in chronically water-short regions in the developing world (UNHabitat, 2006). Providing water sources to urban areas from the developing world is a challenge because nearly one-third of the population (31.2% compared to a 6% in developed countries in 2001) are poor people living in slum areas. The slum growth rate is of 2.37%, a value significantly higher than the average world urban growth rate of 1.78%.
4.06.3.1.1 Municipal discharges
pollutants, such as biological, biodegradable, and nonbiodegradable organic matter, and heavy metals, in that order of importance. The content of almost all these of pollutants is similar around the world, tending to be more concentrated in arid and semiarid areas because of lack of water. In some cases, higher concentrations of pollutants result from increased industrialization of cities. Unfortunately, even when treated, municipal discharges introduce used water containing used compounds, some of which are pollutants, to water bodies. Municipal wastewater is never treated to recover its original quality (the one it had at the water source) as the selfcleansing and dilution capability of nature is used to complete the task. This is confirmed by the increasing amount of trace pollutants, such as endocrine disrupters, found in water sources. The presence of these compounds might be considered as an indicator that we have surpassed the natural depollution capability of the environment. Despite this, the idea of using water bodies or soil to depollute wastewater is still very common, and it could be reduced in water bodies as the depollution capability is lost as result of the water temperature increase due to climate change. In developing countries, the environment is frequently used to depollute wastewater, included when not treated at all, explaining the low quality of water bodies and the widespread presence of diarrheic diseases.
Municipal discharges are those produced by cities and small towns. They are considered to be point sources of pollution where they are produced and collected in sewers and thus disposed of as a well-identified source. When not treated, the main environmental concerns relate to conventional
Industrial wastewater has very variable quality and volume depending on the type of industry producing it. It may be highly biodegradable or not at all, and may or may not
4.06.3.1 Sources of Pollution Traditionally, pollution sources are classified as point and nonpoint sources. Municipal and industrial wastewater discharges are considered to be point sources, while agriculture (considered as the surface return flow from irrigation), storm runoff, and a wide variety of others are considered as nonpoint sources (Jime´nez, 2009a).
4.06.3.1.2 Industrial discharges
Safe Sanitation in Low Economic Development Areas
contain compounds recalcitrant to treatment. These include organic synthetic substances or heavy metals whose content in developing countries’ wastewater may be considerably different (in quantity and quality) from that of developed ones. The main concern with industrial wastewater is the increasing amount (in quantity and variety) of synthetic compounds contained in and discharged to the environment. A list of the most common pollutants in industrial discharges can be found in Jime´nez (2009a). Due to the difficultly in tracking toxic compounds and their fate, combined with the need to use complex and costly treatment methods to remove them from wastewater, it is advisable and cost effective to consider the implementation of cleaner production methods in industries (such as the replacement of toxic recalcitrant compounds with others that are less harmful or not harmful at all) and, also to raise awareness of society to reduce the use of such types of compounds (Jime´nez, 2009b).
4.06.3.1.3 Nonpoint and nonconventional pollutant sources to water Water pollutants come not only from urban and municipal wastewater discharges, but also from nonpoint sources, some of which are not perceived as such. Most of the nonpoint sources have been initially recognized as such by groundwater experts (Foster et al., 2003) who realized that soil (urban or rural) was an important means of transporting pollution to ground and surface water through complex interactions. A list of such pollutants is presented in Table 1 and a detailed description of some of the pollution sources can be found in Jime´nez (2009a).
4.06.4 Pollutants In this section, the types of different pollutants are reviewed, emphasizing those of special interest in developing countries.
4.06.4.1 Biological Pollutants Biological pollutants are the major threat to low-income countries as diseases caused by them are rapidly manifested and have important effects on children and the elderly, sometimes even resulting in fatalities. According to WHO (2004), diarrheal diseases accounts for an estimated 4.1% of the total daily global disease burden and is responsible for 1.8 million deaths every year. It is estimated that 88% of that burden is attributable to unsafe water supply, sanitation, and hygiene. Biological pollutants cause hydraulic diseases that are frequently divided into three categories: 1. Waterborne diseases that are caused by pathogenic organisms ingested when consuming water polluted with fecal contamination or food irrigated with polluted water. Examples of these types of diseases are giardiasis and amebiasis. 2. Water-washed diseases that are caused by the lack of safe water or simply any water for hygiene purposes. Disease transmission is linked to skin or eye contact. An example is trachoma, a disease that causes blindness. Some 6 million people have been blinded by trachoma. Another 150 million need treatment, and an estimated 500 million are at
151
risk. The disease is endemic in 55 countries, with only China and India accounting for 2 million cases. Productivity losses caused by trachoma are estimated to be US$2.9 billion (WHO, 2004). 3. Water-based diseases that are caused when water accumulates and stagnates, promoting the breeding of vectors such as mosquitoes that cause dengue or malaria. There are four groups of organisms that can be found in waste and polluted water: viruses, bacteria, protozoa, and helminths (in the form of eggs, Jime´nez (2003)). The general characteristics of these organisms can be found in specialized literature. In the following sections, properties relevant to developing countries will be highlighted for each type of group. A list of pathogens that have been detected in wastewater is presented in Annex 1. The main aspect to highlight is the notable difference in the quantity and variety of pathogens found in wastewater between developed and developing countries (Table 2).
4.06.4.1.1 Viruses Viruses are the smallest (0.01–0.3 mm) infectious agents. There are more than 150 types of enteric viruses capable of producing infections or illnesses that multiply in the intestine and are expelled in feces. Unlike bacteria, pathogenic viruses are found in wastewater and feces when people are infected, independently of whether they display symptoms. In regions where viral diseases are endemic, they are constantly isolated from wastewater. The presence of viruses and their concentration in wastewater is linked to the season of the year and the age distribution of the population. Concentrations are usually higher during summer and lower in the autumn months. The composition, type, and especially the content of viruses contained in wastewater are poorly known, particularly in developing countries, as a result of the complex and costly analytical techniques required to identify them (Jime´nez, 2003). The enteric viruses most relevant to man are enteroviruses (polio, echo, and coxsackie viruses), Norwalk, rotaviruses, reoviruses, caliciviruses, adenoviruses, and hepatitis A viruses. Rotaviruses are responsible for between 0.5 and 1 billion cases of diarrhea per year in children under 5 years of age in Africa, Asia, and Latin America and up to 3.5 million deaths. Usually, between 50% and 60% of the cases of children with gastroenteritis that are hospitalized are caused by rotaviruses. Reoviruses and adenoviruses are the main causes of respiratory illness, gastroenteritis, and eye infections and have been isolated from wastewater. To date, there is no evidence that the human immunodeficiency virus (HIV) causing the acquired immunodeficiency syndrome (AIDS) can be transmitted via a waterborne route. It is recognized that low virus levels may cause infection or illness; wastewater contains thousands of them, some of which are much more resistant to chlorine disinfection than bacteria (Jime´nez, 2003). Viruses discharged in polluted water can migrate long distances in soil and groundwater. The reported horizontal migration varies between 3 and 400 m, while vertical migration ranges from 0.5 to 70 m depending on soil conditions.
152 Table 1
Safe Sanitation in Low Economic Development Areas Sources of pollution for surface and groundwater
Origin
Relative importance
Concern Developing countries
Developed countries
Cl, NMA ED, F, N, OM, T, PCP, sediments ED, N, OM, PCP Variable DBP, HC, OM, T ED, H, OM, PCP, S, T A, ED, EP, H, HC,NMA OM, PCP, S, T EP, S, T HC, OM, T ED, H, OM, PCP, S, T F, M, N, NMA, OM
þ þ þ þ þ þ
þ þ þ þ þ þ
þ þþ þ þ þ þ
þþþ þþ
þþþ þþþ
þ þ
Variable, more relevant synthetic compounds Variable, more relevant synthetic compounds N, P, T NMA, T HC, T Depending on the type of substance stored
þþþ
þþþ
þþ
þþþ
þþþ
þþ
þ þ þ þþ
þ þ þ þþþ
þþ þþ þ þ
EP, F, N, OM, T EP, F, H, HC, N, OM, T ED, EP, F, OM, N, PCP, S, T
þþþ þþ þþþ
þþþ þþþ þþþ
ED, EP, F, N, OM, PCP, S, T DBP, ED, EP, N, NMA, PCP ED, EP, F, N, OM, PCP S, T ED, EP, H, OM, PCP, S, T
þ þ þ þ
þþþ þ þ þþþ
þ þþ
Other urban sources Atmospheric pollutants deposition Urban run-off Saline intrusion Industrial accidental spillage
A, EP, H, HC, N, M A, B, EP, HC, M NMA EP, T, HC
þþ þþþ þþ þ
þ þ þ þ
þ þþ þþ þ
þþ þþ þ þ
Industrial sources Industries located in urban or rural areas, in general
Variable, mostly synthetic compounds
þþþ
þþþ
þþ
Agricultural sources First use water Treated wastewater Nontreated wastewater
N, P EP, N, P, S EP, F, N, OM, P, S,
þþþ þ þþþ
þþþ þ þþþ
þþþ þþ þ
þþ
þþþ
þ
þþþ þþ
þþþ þþþ
þ
Urban infrastructure Water network Sewerage system Septic tanks and latrines Storage or treatment ponds Storage tanks Municipal landfills Hazardous wastes confinement sites Highways drainage soakways Pipelines Injection wells Cemeteries Urban activities Industries Factories and small commerce Irrigation of amenity areas Application of ice melting substances Transport and transference of material Storage of substances in tanks and reservoirs Urban disposal options Unsewered sanitation Transportation of polluted water in channels or rivers Nontreated sewage disposal in soil with impact on water bodies Nontreated sewage discharge in rivers and lakes Treated wastewater disposal Sludge disposal Uncontrolled dumping sites
Rural areas On-site sanitation systems and unsewered areas Storage of substances in tanks and reservoirs Disposal of solid wastes Transportation of polluted water in channels or rivers
Main polluting agents
EP, F, N, OM, Depending on the type of substance stored EP, ED, F, H, NMA, OM, PCP, S, T EP, F, H, HC, N, OM, T
þ þ þ þ þ
þ þ þ þ
þ þ þ þ þ þ
þ þ þ þ þ
þ
þ
þþ þ þ þþ
Adapted from Jime´nez B (2009a) Coming to terms with nature: Water reuse new paradigm towards integrated water resources management Encyclopedia of Biological, Physiological and Health Sciences, Water and Health, Vol. II: Life Support System, pp. 398–428. Oxford: EOLSS Publishers/UNESCO; Jime´nez (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324 Paris: UNESCO Leiden: Taylor and Francis Group. (a): May include industrial compounds. (b): Only present in industrial areas. A: Acids; Cl: Residual chlorine; DBP: Disinfection by-products; ED: Endocrine disrupters; EP: Emerging pollutants; F: Fecal pathogens; H: heavy metals; HC. Hydrocarbons; N: Nutrients; NMA: Nonmetal and anions; OM: Organic matter; P: Pesticides; PCP: Personal care products; S: Salinity; T: Toxics; þ : Magnitude increase.
Safe Sanitation in Low Economic Development Areas Annex 1
153
Biological disease-causing agents that have been reported in wastewater
Agent
Classification
Illness
Adenoviruses (31 to 51 types) Arbovirus Astroviruses (five types) Calcivirus or Norwalk agent Coronavirus Coxsackie A (enterovirus) Coxsackie B (enterovirus)
Viruses Viruses Viruses Viruses Viruses Viruses Viruses
Echovirus (enterovirus) Enterovirus 68–71
Viruses Viruses
Flavirus Hepatitis A virus Hepatitis E virus Norwalk virus Parvoviruses (three types) Poliovirus (enterovirus) Reoviruses (three types) Rotaviruses (four types) Snow Mountain Agent Small and round viruses Yellow fever viruses Brucella tularensis Campylobacter jejuni Escherichia coli enteropathogenic Legionella pneumophila Leptospira spp., 150 types
Viruses Viruses Viruses Viruses Viruses Viruses Viruses Viruses Viruses Viruses Viruses Bacteria Bacteria Bacteria Bacteria Bacteria
Clostridium perfringens Mycobacterium leprae Mycobacterium tuberculosis Salmonella spp., 1700 a 2400 strains (parathyphi, schottmuelleri, etc.) Salmonella thyphimurium Shigella spp., 4 types Treponema pallidum-pertenue Yersinia enterocolitica Vibrio cholerae Aspergillus fumigatus Candida albicans Balantidium coli Cyclospora cayetanensis Cryptosporidium parvum Entamoeba histolytica Giardia lamblia Naegleria fowleri Plasmodium malariae Trypanosoma spp. Toxoplasma gondii Ancylostoma duodenale Ascaris lumbricoides Echinococcus granulosis Enterobius vermicularis Necator americanus Schistosoma spp. Strongyloides stercoralis Taenia solium Trichuris trichiura Toxocara spp.
Bacteria Bacteria Bacteria Bacteria
Respiratory illness, conjunctivitis, vomiting, diarrhea Arboviral disease Vomiting, diarrhea Vomiting, diarrhea Gastroenteritis, vomiting, diarrhea Meningitis, fever, herpangina, respiratory illness Myocarditis, congenital heart anomalies, rash, fever, meningitis, respiratory illness, pleurodynia Meningitis, encephalitis, respiratory illness, rash, diarrhea, fever Meningitis, encephalitis, respiratory illness, acute hemorrhagic conjunctivitis, fever Dengue fever Infectious hepatitis Hepatitis Epidemic vomiting and diarrhea, gastroenteritis Gastroenteritis Poliomyelitis, paralysis, meningitis, fever Not clearly established Diarrhea, vomiting, gastroenteritis Gastroenteritis Diarrhea, vomiting Yellow fever Tularemia Gastroenteritis, diarrhea Gastroenteritis Acute respiratory illness, Legionnaire’s disease Leptospirosis (septic meningitis, jaundice, neck stiffness, haemorrhages in the eyes and skin) Gaseous gangrene, food poisoning Leprosy Pulmonary and disseminated tuberculosis Salmonellosis
Bacteria Bacteria Bacteria Bacteria Bacteria Fungi Fungi Protozoa Protozoa Protozoa Protozoa Protozoa Protozoa Protozoa Protozoa Protozoa Helminths Helminths Helminths Helminths Helminths Helminths Helminths Helminths Helminths Helminths
Typhoid fever, paratyphoid or salmonellosis Bacillary dysentery, Shigellosis Yaws (frambuesia) Gastroenteritis, Yersiniosis Cholera Aspergillosis Candidiasis Mild diarrhea colonic ulceration, dysentery, balantidiasis Severe infectious, dehydration: diarrhea, nausea, vomiting Diarrhea and cryptosporidiosis Amoebic dysentery Giardiasis Amoebic meningo-encephalitis Malaria Trypanosomiasis Congenital or postnatal, toxoplasmosis Anaemia, ancylostomiasis Ascariasis Hyadatidosis Enterobiasis Anaemia Schistosomiasis Diarrhea, abdominal pain, nausea, Strongylodiasis Taenisis, cysticercosis Diarrhea Fever, abdominal pain, nausea
The presence of biological disease-causing agents is not necessarily an indication of a confirmed risk. From Jime´nez B (2003) Health risks in aquifer recharge with recycle water. In: Aertgeerts R and Angelakis A (eds.) State of the Art Report Health Risk in Aquifer Recharge Using Reclaimed Water, pp. 54–172. Rome: WHO Regional Office for Europe.
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4.06.4.1.2 Bacteria Bacteria are single-celled microorganisms ranging from 0.2 to 10 m in size with different shapes. They reproduce and grow in an appropriate environment at defined ranges of temperature, Table 2 Comparison of the biological pollutant content in wastewater from developing and developed countries Organism Enteric viruses, PFU 100 ml1 Salmonella, MPN 100 ml1 (M,
(U, I) U, F, SA, IN,
Developed world
Developing world
102–104 100–104
104–106 106–109
104–106 101 1–103 1–103
106–4107 103 102–103 ND
H)
Fecal streptococci, No. 100 ml 1 (U, B, K) Protozoan cysts, organisms l1 (U, M) Giardia lamblia, cysts l1 (U, E, K) Cryptosporidium parvum, oocysts l1 (U, E)
Helminth ova, egg l1
1–9
6–800
Data from: E, England; H, Holland; In, India; I, Israel; K, Kenya; M, Mexico; SA, South Africa; U, USA; ND, No data. Adapted from Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO Leiden: Taylor and Francis Group.
Table 3
salinity, pH, etc. They may or may not be encapsulated. The environmental distribution of bacteria is ubiquitous and has different nutritional requirements. Many species of bacteria are not harmful to man. In fact, some even live inside humans forming intestinal colonies. Bacteria are expelled in feces at high concentrations (Jime´nez, 2003). Table 3 shows some characteristics of pathogenic bacteria that can be found in the feces of infected people. In wastewater, pathogenic bacteria are always present but at a variable concentration, depending on the local health conditions. As shown in Table 3, due to the high rate of diseases caused in developing countries, Salmonella, Shigella, and Helicobacter pylori are bacteria of importance as agents causing endemic diseases. In contrast, Vibrio cholerae is present only when an epidemic exists.
4.06.4.1.3 Protozoa Protozoa are the group of parasites most closely associated with diarrheas. They are single-celled organisms (2–60 mm in size) that develop in two ways: as trophozoites and as cysts. Infections are produced when mature cysts are consumed. Cysts are resistant to gastric juices and transform themselves into trophozoites in the small intestine, lodging in the wall where they feed on bacteria and dead cells. In time,
Characteristics of some bacteria frequently found in wastewater (with information from Jime´nez (2003) and Lenghton et al. (2005))
Characteristics and effects in humans Escherichia coli is commonly found in wastewater at high concentrations. Different E. coli strains can cause gastroenteritis in both animals and humans and pose a high risk to newborns and children under 5 years of age. E. coli strains implicated with human diseases are: (1) enteropathogenic E. coli ; (2) E. coli that is the common cause of traveler’s diarrhea, which provokes a liquid and profuse diarrhea with some mucosity, nausea, and dehydration; (3) enteroinvasive E. coli that invades the intestinal mucus lining like Shigella spp., and (4) E. coli (EHEC) that produces a similar toxin to Shigella causing hemorrhagic colitis. Infective doses are relatively low (102 organisms). Salmonella spp. is frequently present in wastewater at content always lower than that of fecal coliforms by 1–2 log. There is a wide variety of strains capable of infecting humans and animals. The incidence in humans is lower than in animals and has a seasonal variation. The most severe form of salmonellosis is typhoid fever caused by Salmonella typhi. Typical symptoms are chronic gastroenteritis with diarrhea, stomach cramps, fever, nausea, vomiting, and headache. In severe cases, collapse and death might occur. Transmission is through ingestion of polluted water or food, and is very common in developing countries. Infective dose is of the order 105–108 microorganisms, but for Salmonella typhi doses as low as 102–103 have been reported. Shigella is similar to Salmonella spp. but less frequent in wastewater. There are more than 40 strains, but S. sonnei and S. flexeneri represent almost 90% of total wastewater isolations. It rarely infects animals and lives for a shorter period in the environment. One route of transmission is through swimming in polluted water. Shigella spp. produces bacillary dysentery or shigellosis. This is light watery diarrhea that can develop into full-blown dysentery. The symptoms are fever, nausea, vomiting, abdominal pain, migraine, and myalgia. The classic form of dysentery is characterized by the expulsion of feces containing blood with or without mucus. The infective dose is less than 103 microorganisms. Helicobacter pylori is found in wastewater. Its major habitat is the human gastric mucosa. Three species are human pathogens: H. pylori, H. fennelliae, and H. cinaedi. The pathway of transmission is not entirely clear but water could be involved. In developing countries, H. pylori is acquired early in childhood, and up to 90% of children are infected by the age of 5. This contrasts with the low infection rate during childhood observed in developed countries (0.5–1%). Campylobacter jejuni usually is a pathogen to animals but it can cause severe gastroenteritis in humans. The main source of infection is nonchlorinated water supplies. Mycobacterium tuberculosis along with M. balnei (marinum) and M. boris causes pulmonary diseases and tuberculosis. For M. tuberculosis, contaminated water is the main source of infection. Vibrio cholerae is the cause not only of epidemic but also eight pandemics, the last one between 1990 and 1995. Cholera epidemics are caused by V. cholerae group O1 and some non-O1. Symptoms are abundant liquid diarrhea with significant loss of hydro-electrolytes and severe dehydration associated with vomiting. V. cholerae is rare in developed countries but frequent in poor ones. Humans are the only known hosts. The most frequent pathway of transmission is water, either through direct consumption or when used to irrigate produce that is consumed uncooked. Fish grown in polluted water are another source of transmission. Since 2007, there have been outbreaks of cholera in India, Iraq, Congo, Vietnam, and Zimbabwe. In 2005, West Africa suffered more than 63 000 cases of cholera, leading to 1000 deaths.
Safe Sanitation in Low Economic Development Areas Table 4
155
Protozoa related to sanitation problems and that are of interest for developing countries (with information from Jime´nez (2003))
Characteristics and effects in humans Entamoeba histolytica is one of the most important parasites detected in municipal wastewater and is commonly known as Amoeba. Trophozoites measure 20–40 mm and t cysts 10–16 mm. Amoebae usually lodge in the large intestine; occasionally they penetrate the intestinal wall, traveling and lodging in other organs. They are the cause of amoebic and hepatic dysentery. Entamoeba histolytica infects 10% of the world’s population – mostly in the developing world – resulting in approximately 500 million infected persons; there are between 40 and 50 million cases of invasive amebiasis per year resulting in up to 100 000 annual deaths (placing it second after malaria in mortality caused by protozoan parasites). Ninety-six percent of these cases occur in poor countries, especially on the Indian subcontinent, West Africa, the Far East, and Central America. Giardia spp. are common in wastewater as it frequently causes endemic diseases. It especially affects children under 5 suffering from malnutrition. The total number of sick people is of the order 1.1 billion, 87% of whom live in poor countries. Giardia spp. is the most common parasite of humans but water is not necessarily the main pathway of transmission. Cysts (that are 8–14 mm long and 7–10 mm wide) can survive in water bodies for long periods, especially in winter. Giardia lives in the intestines of a large number of animals as trophozoites. The disease is characterized by very liquid and smelly explosive diarrhea, stomach and intestinal gases, nausea, and loss of appetite. Cryptosporidium spp. is a parasite widespread in nature. Oocysts are resistant to chlorine and due to their small size (4–7 mm) are difficult to remove from water, as many other protozoan. Cryptosporidium spp. infects a large spectrum of farm animals and pets and was recently recognized as a human pathogen that is why it is considered as an emerging pathogen. Cryptosporidium spp. is capable of completing a life cycle within the same host and causing reinfection. Once an individual has been infected, the person carries the parasite for life and can be reinfected. The disease rate in developing countries has been poorly studied, in particular due to the higher occurrence of other types of diseases. Cryptosporidiasis in developing countries has shown a greater incidence among immune depressed people and in rural areas (Snelling et al., 2007). The main symptoms of cryptosporidiasis are stomach cramps, nausea, dehydration, and headaches. Although it is known that the infectious dose varies between 1 and 10, outbreaks have always been associated with large concentrations in water.
trophozoites become once again cysts that are expelled in feces. Infected persons may or not display symptoms. Protozoa do not reproduce in the environment, only in their host. However, they are able to survive in the environment and remain active for periods ranging from some months to up to several years, depending on the environmental conditions. Most intestinal protozoa are transmitted through polluted water and food contaminated with polluted water or unsanitary handled (Jime´nez, 2003). Table 4 shows the characteristics of some protozoa. In the developing world, the more relevant protozoa because of their effects on humans are Giardia and Amoeba. Cryptosporidium is a threat to developed countries, as was unfortunately demonstrated in Milwaukee, US, when 403 000 people became ill and more than 50 died after an infection was transmitted through the drinking water supply (Hrudey and Hrudey, 2004).
4.06.4.1.4 Helminth eggs Helminths are worms some of which are parasites in humans. Where helminths are the origin of waterborne diseases, they are mainly transmitted through the consumption of contaminated food (crops, meat, or fish). Helminths can also be transmitted through the oral–fecal route and, therefore, hygiene is important as a factor in their control. As helminths are associated with turbid water, they normally are not a concern in drinking water. Helminths are pluri-cellular worms and because of this they are poorly addressed in environmental microbiology books. The eggs – their infective form – are microscopic and travel along with wastewater. Helminths occur in different types and sizes (from 1 mm to several m in length), and have diverse and complex life cycles compared to most of the microorganisms known in the sanitary field (Jime´nez, 2008a).
Before infecting humans, in some cases, they may have an intermediary host as is the case for Schistosoma spp. that temporarily lives in snails. There are three different types of helminths: (1) plathelminths or flat worms, (2) nemathelminths, nematodes or round worms, and (3) annelids. If plathelminths have their body formed by segments, they are called cestodes; if not, they are then called trematodes. Only the first two types are of sanitary importance. Although common in sanitary engineering literature, it is improper to use the terms nematodes, Ascaris, and helminths as synonyms. This misunderstanding comes from the fact that Ascaris (a nematode) is the most common helminth egg in wastewater and sludge. A list of helminth eggs found in wastewater and sludge and its classification can be found in Jime´nez (2008a). Helminthiases are diseases of high incidence in developing countries compared with developed ones. Globally, there are around 1–2 thousand million people suffering of helminthiases but most of them are from developing countries where it affects up to 10% of the population. The incidence rate may reach 90% in regions where poverty and poor sanitary conditions prevail. In contrast, in developed countries, helminthiases’ incidence is at the most 1.5% and affects mainly poor immigrants (Jime´nez, 2008a). Helminthiases have different manifestations but, in general, they cause intestinal wall damage, hemorrhages, deficient blood coagulation, and undernourishment. They can degenerate into cancer tumors. Helminthiases affect mainly children, the elderly, and poor people (Jime´nez, 2008). Around 94% of the more than 4 billion cases of diarrhea in the world are caused by helminths (Murray and Lo´pez, 1996). There are several kinds of helminths with different local names (Annex 2). This along with the fact that it is hard to properly identify them clinically unless a costly laboratory analysis is performed, makes it difficult to track the actual incidence of all the
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Safe Sanitation in Low Economic Development Areas
Annex 2
Examples of local names given to helminth and helminthiases diseases
Common name
Technical name
Examples of local names given to diseases
Number of infected people (million)
Region affected
Foodborne trematodes and schistosomiasis
Trematode
4240
Found in 74 countries.
Blood fluke
Schistosoma
Trematodiases, clonorchiasis, schistosomiasis, fasciolasis Schistosomiasis, bilharziosis or snail fever
200 half of which live in Africa (20 with severe consequences)
Liver fluke
Clonorchis sinensis
Clonorchiasis
40 (10% of the world’s population thought to be at risk)
Liver fluke
Fasciola hepatica and F. gigantica
Fascioliasis
Intestinal Fluke
Fasciolopsis buski
Fasciololopsis
Hookworms
Ancylostoma duodenale
Ancylostomiasis, anchylostomiasis, helminthiasis, miners’ anemia, tunnel disease, brickmaker’s anemia and Egyptian chlorosis Necatoriasis
Asia, Africa, and South America. (80% of whom live in sub-Saharan Africa China, Russian Federation, Republic of Korea, Vietnam Temperate areas of Africa, Europe and Central/ South America Kazakhstan, Lao Peoples0 s Democratic Republic, Poland, Russian federation, Thailand, Turkey, Ukraine, Viet Nam Middle East, North Africa, India and (formerly) in southern Europe
Necator americanus
Tapeworm
All cestode
Tape worm Tapeworm
Taeniasis, Cysticercosis
Roundworm nematode
Taenia Hymenolepis nana and diminuta All nematode (Ascaris, Toxocara, Trichuris Enterobius) Ascariasis lumbriocides
Pinworm Whipworm
Enterobius vermicularis Trichuris trichiura
Roundworm
1300
The Americas, Subsaharan Africa, Southeast Asia, China, and Indonesia Asia, Africa, South America, parts of Southern Europe and pockets of North America
Nematode infection
4000
Latin America, Asia, Africa, far East
Ascariasis
1500
Africa, Asia and Latin America, Far East
Oxiuriasis Enterobiasis Trichuriasis
600 1050
helminthiases. That is why frequently figures are underestimated. Technically, helminthiases take their name from their causative agent. For instance, trichuriasis is named after Thrichuris. Ascariasis, affecting nearly 1500 million people, is the most common of the helminthiases and is endemic in Africa, Latin America, and the Far East. Even though the mortality rate is low, most of the people infected are children under 15 years of age with problems of faltering growth and/ or decreased physical fitness. Around 1.5 million of these children will probably never bridge the growth deficit, even if treated (Silva et al., 1997; Jime´nez, 2008a).
The helminthiases’ infective agents are the eggs, not the worms. Actually, worms cannot live either in wastewater or in sludge because they need a host. Helminth eggs are transmitted through (1) the ingestion of crops polluted with wastewater or sludge, (2) direct contact with polluted sludge or fecal material, and (3) the ingestion of polluted meat or fish (Jime´nez, 2008a). Each type of helminth has its own pathways of infection. Eggs of different helminths generally occur in different shapes, sizes, and resistances (Figure 3). As a result of the higher incidence of ascariasis, in wastewater and sludge, these
Safe Sanitation in Low Economic Development Areas
Egg fertile roundworm Ascaris 40−80 μm × 25−50 μm
Ascaris egg, four-cell stage
Ascaris egg. With eight or more cells
Ascaris egg with a young worm (200−300 × 14 μm)
Ascaris egg, the shell loses resistance to allow hatching
Ascaris egg hatching
Nonfertile Ascaris egg 80−90 μm × 30−40 μm
Egg of the tapeworm Hymenolepis nana 30−47 μm
Egg of the tapeworm Taenia 30−40 μm
Egg of the tapeworm Hymenolepis diminuta egg 80 μm Hymenolepis diminuta 70−80 μm
157
Hymenolepis diminuta hatching
Figure 3 Examples of helminth eggs most frequently observed in wastewater and sludge, Photographs courtesy of Catalina Maya, Treatment and Reuse GROUP, UNAM.
are the eggs found in the highest concentrations (Figure 4). The percentage of types of helminths might vary from one region to another following the disease’s pattern. Due to differences in health conditions in developed and developing countries, their helminth eggs content is very different in wastewater and sludge (Table 5). Eggs contained in sludge are not always viable and infectious. To be infectious, the larvae need to develop, and, for that, a certain temperature and moisture are needed. The necessary conditions are frequently met in soil or crops, where eggs are deposited when polluted wastewater, sludge, or excreta is used as fertilizer. Under such conditions, the larvae develop in 10 days. According to previous information (that has not been updated using better analytical techniques), Ascaris eggs remain viable 1–2 months in crops and many
months in soil, freshwater, sewage, feces, night soil, and sludge – periods which are much longer than those for microorganisms (Jime´nez, 2008a, Figure 5). This high resistance is due to a cover composed of 3–4 layers that gives mechanical resistance to eggs and protects them from desiccation, strong acids and bases, oxidants, reducing agents, detergents, and proteolytic compounds (Jime´nez, 2008a). The resistance of different helminth eggs genera under environmental conditions has not been reported in literature. To inactivate helminth eggs, it is recommended to raise the temperature above 40 1C for 10–20 days for Ascaris or to reduce moisture levels below 5%. These conditions are not ease of use during wastewater treatment; thus, helminths are usually removed from wastewater to be subsequently inactivated in sludge. Helminth ova of interest in the sanitary field
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Safe Sanitation in Low Economic Development Areas
Egg of the roundworm fertile Toxocara 85−95 μm
Toxocara larva inside the egg, infective stage (300−400 × 40 μm)
Toxocara egg, two-cell stage
Toxocara egg, four-cell stage
Toxocara hatching
Toxocara larva
Egg of the whipworm fertile Trichuris egg, infectious stage Trichuris 50−54 mm × 22−23 μm)
Egg 50−60 μm × 20−30 μm of pinworm Enterobius vermicularis with larva
Figure 3 Continued.
Trichosomoides 80 μm × 50 μm egg of a nematode with larva
Trichuris egg hatching
Trichosomoides sp. with damaged larva
Safe Sanitation in Low Economic Development Areas
Mexican wastewater
Mexican sludge
159
South African ecosan sludge
90 80 70 60 50 40 30 20 10
Uncinaria
p. Taenia sp
Enterobiu
s spp.
. oides spp Trichosom
spp. Toxocara
pis spp. Hymenole
spp. Trichuris
Ascaris s
pp.
0
Figure 4 Content of different helminth egg genera in Mexican wastewater and sludge and from an on-site sanitation system in South Africa. Data from Maya C, Jime´nez B, and Schwartzbrod J (2006) Comparison of techniques for the detection of helminth ova in drinking water and wastewater. Water Environment Research 78(2): 118–124 and Jime´nez B and Wang L (2006) Sludge treatment and management. In: Ujang Z and Henze M (eds.) Municipal Wastewater Management in Developing Countries: Principles and Engineering, pp. 237–292. London: IWA Publishing.
Table 5 Helminth ova content in wastewater and sludge from different countries Country/region
Municipal wastewater (HO l1)
Sludge (HO g1 TS)
Developing countries Brazil Egypt
70–3000 166–202 6–42
Ghana Jordan Mexico
No data 300 6–98 in cities Up to 330 in rural and peri-urban areas 840 800 60 9 No data No data 1–8
70–735 75 Mean: 67; maximum: 735 76 No data 73–177
Morocco Syria Ukraine France Germany Great Britain United States
No data No data No data 5–7 o1 o6 2–13
From Jime´nez B (2008a) Helminth ova control in wastewater and sludge for agricultural reuse. Water reuse new paradigm towards integrated water resources management. In: Grabow WOK (ed.) Encyclopedia of Biological, Physiological and Health Sciences, Water and Health, Vol. II. Life Support System, pp. 429–449. Oxford: EOLSS Publishers/UNESCO.
measure 20–80 mm, have a specific density of 1.06–1.2, and are very sticky. These properties are used to remove eggs from wastewater (Jime´nez, 2008a). Helminth ova criteria. As shown in Table 5, not all wastewater and sludge contain significant amounts of helminth ova. For this reason, they are not included in all countries’ wastewater, sludge, or fecal sludge norms, as is the case with biochemical oxygen demand (BOD) or fecal coliforms, which are universal parameters used to design wastewater treatment (Jime´nez, 2008a). Based on toxicological and epidemiological studies, the World Health Organization WHO (2006) suggested a value of r1 egg l1 in wastewater intended for the irrigation of crops that are eaten uncooked. Wastewater used for the culture of fish should contain 0 egg l1, since trematode eggs (Schistosoma spp., basically) may multiply in an intermediary host (a snail) before infecting fish and humans. For excreta, the recommended criterion is of 1 egg g1total solids (TS).
4.06.4.1.5 Biological indicators Thermotolerant coliform bacteria (commonly referred as fecal coliforms) are the group most frequently used as indicators of fecal pollution because they behave in a similar way to most pathogenic bacteria in the environment, and, during treatment, they are abundant and easy to determine.
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Fresh and wastewater Ascaris lumbricoides ova E. histolytica cysts Shigella spp / Vibrio cholerae Fecal colifrom / Salmonella
Enteroviruses
0
20
40
80
60
100
120
140
100
120
140
Days
Crops Ascaris lumbricoides ova E. histolytica cysts Vibrio cholerae Shigella spp Fecal coliform / Salmonella Enteroviruses
0
20
40
80
60
Days
Soil Ascaris lumbricoides ova → Shigella spp / Vibrio cholerae / E. histolytica cysts Fecal coliform / Salmonella Enteroviruses
0
20
40
60
80
100
120
140
Days Figure 5 Survival time of different pathogens in fresh and wastewater, soil and crops at 20–30 1C. Data from Feachem R, Bradley D, Garelick H, and Mara D (1983) Sanitation and Disease: Health. pp. 349–356. New York, NY: Wiley.
Safe Sanitation in Low Economic Development Areas
Thermotolerant coliforms are less specific indicators of fecal contamination than Escherichia coli, since they may sometimes arise from nonfecal sources, especially in tropical climates (WHO, 2004). However, it is becoming increasingly evident that they are not useful to simulate the behavior of all enteric viruses, protozoa – in particular with regard to Giardia and Amoeba – and helminth eggs that are of concern in low-income regions. Despite this, it is frequently, but wrongly, assumed that fecal coliforms are indicators of all kinds of biological pollution. Even though they can be useful indicators of fecal pollution in developed countries’ drinking water, this is not always the case for water and wastewater from developing ones, owing to the presence of a wider variety and larger quantities of microorganisms (Jime´nez, 2009). This does not mean that fecal coliforms are not useful for developing countries; it simply means that care must be taken to select additional indicators for specific purposes, such as for wastewater and sludge reuse in agriculture and aquaculture. In these cases, the helminth egg content (WHO, 2006) needs also to be specified. It is worth mentioning that the treatment procedures to inactivate helminth eggs are frequently developed using Ascaris eggs as models as they have been informally considered as indicators for all helminth eggs, although this has not been fully proven experimentally. In other cases, Taenia saginata or Ascaris galli, types of eggs that are rarely present in wastewater, are used to test treatment procedures.
4.06.4.1.6 Emerging pathogens Some pathogens that are not usually followed during conventional monitoring have been linked to outbreaks in developed countries. These pathogens have been called ‘emerging’ pathogens. They have led to new regulations as well as to improvements in water and wastewater treatment procedures. Some of the microorganisms considered as emerging pathogens are Giardia lamblia, Cryptosporidium parvum, Cyclospora cayetanensis, Blastocystis hominis, Legionella pnuemophila, E. coli 0157H7, Campylobacter, Mycobacterium, and Norovirus (Jime´nez, 2009b). In developing countries, some of these pathogens are endemic, while others have either not been studied or not reported as disease-causing agents.
4.06.4.1.7 Biological analytical techniques Assessing the biological quality of water is always a challenge due to the diversity of organisms and the need for different and proper methods to identify and enumerate them, some of which are complex, time consuming, and costly. In the following sections, a short description on the techniques used for different type of organisms is described. Viruses. Identification and quantification of viruses in wastewater, sludge, or excreta is complicated due to the low level of recovery from wastewater and the need to use complex and costly techniques to analyze them. A laboratory requires 14 days, on average, to determine the presence or absence of a virus in water and another 14 days to identify them, using conventional procedures. Polymerase chain reaction (PCR) techniques have considerably speeded up the process, as they can be used to determine viruses online. These techniques are based on the amplification of a single or few copies of a piece
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of DNA allowing the identification of different types of viruses. However, quantification with the precision required in the sanitary field remains a challenge. In addition, the method is sophisticated, and requires highly specialized equipment and highly trained personnel. Due to these difficulties, it is sometimes preferred to detect bacteriophages, that is, bacteria infected by viruses. Bacteriophages are used as informal indicators of viruses and not been linked to human diseases; therefore, their presence has no health significance (Jime´nez, 2003). Bacteria. As mentioned previously, thermotolerant bacteria are the common accepted indicator of bacterial fecal pollution. They are detected by using a selective medium and incubating it after inoculation at 35 or 3770.5 1C and/or 44 or 44.570.25 1C, depending on the medium used. The materials and equipment used for this analysis are very common in most wastewater laboratories. PCR techniques to detect E. coli are useful as well. Protozoa. There are enough accessible techniques to determine the presence of the main protozoan pathogens in wastewater and sludge; however, fewer techniques are available to quantify them with the required precision for the sanitation field. The presence of protozoa on samples does not necessarily always imply a risk, since this requires them to be also viable. To determine the viability, several days are required. PCR techniques for protozoa are not as well developed as they are for bacteria and viruses. Helminth eggs. Helminths eggs require laborious techniques to detect them and even more so to enumerate them. Fortunately, the technique is readily available and does not use complex equipment, although it does require well-trained laboratory personnel. Currently, there is no standardized method and most of the few laboratories trained to detect them are using either different analytical procedures or similar ones with modifications. Moreover, most of the laboratories, instead of reporting the total content of helminth eggs, only report the Ascaris content, as is done in developed countries where it is frequently the single type of helminth eggs present (Jime´nez, 2008a). Analytical techniques for quantifying helminth eggs can be divided into two: direct and indirect techniques (Jime´nez, 2008a). The first consists of separating helminth ova from the other particles contained in wastewater or sludge (where there are many) and then identifying and counting different genera using a microscope. Some examples of these techniques used the US-EPA (United States-Environment Protection Agency), the membrane filter, the Leeds I and Leeds II, and the Faust techniques. The most widely used technique seems to be the US-EPA (1992). A comparison of the performances of the above-mentioned methods has been made by Maya et al. (2006). The recovery rate among them varies from 20% to 80%. Sensitivity for each notably varies as well and not all are capable of measuring the criteria values set by WHO (2006) of 1 egg l1 for wastewater and 1 egg g1 TS for sludge. The second types of techniques are indirect ones, and these have been applied only for wastewater. They are based on measuring either the total suspended solids (TSS) content or the particle size distribution (PSD), and then correlating the concentration to the helminth egg content. Calibration curves need to be established for each type of wastewater and
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Box 1 Endocrine compounds. From Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO Leiden: Taylor and Francis Group. Endocrine disruptors are chemicals that mimic hormones or have antihormone activity interfering with the functioning of endocrine systems in various living species. They derive from many sources including pesticides, persistent organic pollutants, nonionic detergents, and human pharmaceutical residues. Some of them have been identified in municipal wastewater and many of them may persist in the environment for some time. Endocrine disruptors have been also found in drinking water. Their presence in recycled waters also raises broader questions about the risks and benefits of water recycling and our approaches to anticipating the emergence of new contaminants. Human health effects potentially linked to exposure to these chemicals include breast, prostate, and testicular cancer; diminished semen quantity and quality, and impaired behavioral, immune or thyroid functions in children. Although direct evidence of adverse health effects in humans is lacking, reproductive abnormalities, altered immune function, and population disruption potentially linked to exposure to these substances has been observed in amphibians, birds, fish, invertebrates, mammals, and reptiles. Notably, feminization or masculinization on male or female animals, respectively, has been reported.
treatment process. Nevertheless, it is a worthwhile method because the helminth egg determination costs US$7–12 if TSS are used, and US$3 with the PSD, instead of US$70, which is the cost of direct methods. It is important to distinguish between fertile viable and nonfertile eggs as only the viable eggs are infectious. This can be done visually using stains or by incubation at 26 1C for 3–4 weeks (Jime´nez, 2008a).
4.06.4.2 Conventional Parameters Conventional parameters as understood in this text are those commonly used to design or select wastewater and sludge treatment processes worldwide, and they refer mainly to the organic matter content (measured as BOD or COD – biological or chemical oxygen demand), or suspended solids. In general, they are similar worldwide except for the heavy metals content that in general –and especially for sludge – is notably lower in developing countries than in developed ones (LeBlanc et al., 2008) as result of the difference at the industrialization level. However, at a local level, metal content in some industrialized areas of developing countries, notably where metal or tanning industries are placed, may be high. A detailed description of conventional parameters and their significance can be found in Jime´nez (2009a).
4.06.4.3 Emerging Pollutants The term (chemical) ‘emerging pollutant’ is used to describe a wide variety of complex organic chemical compounds that are candidates for future regulation and that have not usually been monitored. To detect them, complex and costly analytical equipment is needed, such as GC-MS or GC-MS-MS (gas chromatography coupled with one or two mass spectrometers) as these are the only ones capable to measure the very low concentrations at which the pollutants are present (in the order of micro- or ng l1) and to identify them. Emerging pollutants have been detected in untreated wastewater, treated wastewater, surface water, groundwater, and even in drinking water of both developed and developing countries (some). Among the countries that have measured and detect emerging pollutants, the following can be cited: Austria, Brazil, Canada, Finland, Germany, Italy, Japan, Mexico, the Netherlands, Spain, Switzerland, UK, and USA (Jime´nez, 2009b). The sources of emerging pollutants are diverse. They come from nonpoint sources, municipal wastewater (treated or nontreated), and industrial discharges. They are also the result
of the improper disposal of solid wastes. Two groups of compounds that are considered as emerging pollutants are: endocrine disrupter compounds (Box 1) and personal care and pharmaceutical products (PCPPs). Wastewater treatment processes have not been designed to remove them; thus, they are randomly removed during conventional treatment. From the limited literature currently available, emerging pollutants – as other organic compounds – are concentrated in sludge during wastewater treatment. Initial risk studies suggest minimal ecological and health effects through biosolids recycling to soils (LeBlanc et al., 2008). As most of these pollutants have only been recently studied, the knowledge of their fate, transport, behavior during treatment, and risks is still poor in the sanitary engineering field. Chemical emerging pollutants, in general, are not considered at the moment as a priority for the developing world as there are more pressing health and environmental pollutants of concern.
4.06.4.4 Risks It is important to bear in mind that the simple presence of a pathogen or a toxic chemical in wastewater, sludge, or excreta does not necessarily mean that a negative effect will occur. For that, several other things need to happen. These include (1) the need for a compound/pathogen to reach a certain concentration; (2) the existence of a pathway for transmission to human or the environment; (3) the ingestion or presence of a certain dose to cause long- or short-term effects; (4) sufficient exposure times to the pollutant; and (5) sufficient sensitivity of a person or of the environment to pollutants. In addition, it should be remembered that, for humans, water is not the only source of risk, as food and air are also sources of pollutant ingestion and, in some cases, they may be the main ones. In terms of the differences of biological risks to humans in developing and developed countries, there are additional aspects to consider as humans develop immunity to pathogens depending on the type of environment they are exposed to, and thus infectious doses may be higher. Genetic history, nutrition, and the combination of social patterns also intervene. For these reasons, data developed for developed countries are not always applicable to developing ones to perform risk analysis. In order to quantitatively assess risks, it is necessary (1) to establish the type and quantity of given microorganisms in a region, (2) to know the actual infectious dose, and (3) to define and evaluate the possible infection route. To
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quantitatively evaluate the risk from a chemical or microbial pollutant, several methodologies are available in literature, but the data needed to apply them may be lacking for special cases in developing countries.
In 2002, the World Summit on Sustainable Development (WSSD) provided a definition for basic sanitation that, besides considering the service itself, considered its impact on human health. This definition comprises the following:
• 4.06.5 Sanitation in Low-Income Countries: A Complex Current Situation 4.06.5.1 Sanitation Needs a Definition Sanitation is a term that has a clear meaning in the developed world. However, for the developing one, there is need to have a better definition. Traditionally, sanitation has been reported as the percentage of the population having access to the service. In practice, this service in low-income regions ranges from simple access to sewers that are discharging the wastewater just behind households or into the streets to sewers connected to sophisticated wastewater treatment plants coupled with water reuse projects and comprising safe sludge management practices. For basic sanitation – sanitation provided in rural or poor periurban areas, the term sanitation includes a wide variety of on-site sanitation options going from simple pit to highly comfortable package treatment plants, which may or may not be functioning. To overcome this, the Joint Monitoring Programme (JMP) from WHOUNICEF proposed in 2000 to introduce the term ‘improved sanitation’. Improved sanitation is a system in which excreta are disposed of in such a way that the risk of fecal–oral transmission to users and to the environment is reduced (WHO–UNICEF, 2008). Table 6 shows which options qualify as improved sanitation and which do not. Table 6 Improved and unimproved sanitation facilities according to WHO–UNICEF (2008) Improved
Unimproved
Connection to public sewer or septic tank Pour-flush latrine Pit latrine with slab VIP latrine Ecological sanitation
Service or bucket latrine Traditional latrine Public latrine or shared toilet Open pit or pit latrine without a slab Open defecation in bush or field
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• • • • • •
the development and implementation of efficient household sanitation systems; the improvement of sanitation in public institutions, especially in schools; the promotion of safe hygiene practices; the promotion of education and outreach focusing on children, as agents of behavioral change; the promotion of affordable and socially and culturally acceptable technologies and practices; the development of innovative financing and partnership mechanisms; and the integration of sanitation into water resources management strategies in a manner that does not negatively affect the environment (it includes protection of water resources from biological or fecal contamination).
As a result, the WSSD’s focus is not only on the construction of a particular number of toilets but also on the effective improvement of health and hygiene through basic sanitation. However, still new elements are needed to be added as problems caused by lack of sanitation are combined with those arising from the lack of economic resources and frequently also with lack of water in societies lacking even from social, economical, and political rights (Box 2).
4.06.5.2 Millennium Development Goals The Millennium Development Goals (MDGs) are drawn from the actions and targets contained in the Millennium Declaration that was adopted by 189 nations and signed by 147 heads of state and governments during the UN Millennium Summit held in New York City on September 2000 (WHO– UNICEF, 2009). They comprise eight goals and 21 quantifiable targets. Water is part of the 7th Goal under Target 7c: ‘‘Reduce by half the proportion of people without sustainable access to safe drinking water and basic sanitation.’’ Fulfilling this target represents the challenge of providing safe water supply to 1.1 million people and safe sanitation to 2.6 million people within 15 years.
Box 2 What sanitation should include, with some information from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan. * * * * * * * * * * *
Safe collection, storage, treatment and disposal, reuse, or recycling of human excreta (feces and urine). Drainage and safe disposal, reuse, or recycling of household wastewater (often referred to as sullage or grey water). Management, minimization, reuse, and recycling of solid wastes (trash or rubbish). Use of goods producing less solid wastes. Drainage, safe management, and even reuse or recovery of storm water. Treatment and disposal, reuse, or recycling of sewage effluents and wastewater by products. Collection and management of industrial waste products, and, the promotion of cleaner industries, vis-a`-vis water. Management of hazardous wastes (including hospital wastes and chemical, radio-active, mining, petrochemical, and other dangerous substances). The use of sanitation as a way to properly reintegrating water, organic matter, and nutrients into the environment in order for them to be safely used again. Provision of water in a sufficient amount to maintain clean households and to allow proper hygienic habits. The recognition of a right for sanitation at the same level of the right to water. The sanitation as an instrument to differentiate social classes, gender, children, and ethnic groups.
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Box 3 Some figures for global sanitation * * * *
For each four persons that do not have access even to a simple pit latrine, six have it. For each one person that does not have access to sanitation, another one has it. In rural areas, for each two persons, only one person has access to a sanitation service. For each 7 l of wastewater that is nontreated, 1 l is treated.
4.06.5.3 Present Situation Reporting figures concerning the state of sanitation in the developing world is a difficult task. First, there is a lack of information; second, the information available is generally presented in a heterogenic way; and third, different sources tend to contradict each other despite national and international efforts to produce consensus.
4.06.5.3.1 General overview The worsening situation with regard to sanitation in developing countries can be described using different indicators (Box 3). Contaminated water and poor sanitation account for the vast majority of the 1.8 million child deaths each year from diarrhea – almost 5000 every day – making it the second largest cause of child mortality (UNDP, 2006). The expansion of water services is essential to reduce the burden of waterrelated diseases and to improve the well-being of a large part of the world’s population. It is also vital for economic development and poverty alleviation (WHO, 2004). According to the figures presented by WHO–UNICEF (2006), despite the efforts made and due to population growth, between 1990 and 2004, the population with access to sanitation services has increased from 2569 million to 3777 million (47%), while the net number of people without improved sanitation decreased by only 98 million.
4.06.5.3.2 Regional situation The difference between the level of sanitation in developed and developing countries is high: 99% versus 50% (Table 7). However, between 1990 and 2004, the percentage of people with access to improved sanitation increased from 35% to 50% with countries’ variations ranging from 37% to 88% (WHO–UNICEF, 2006). The difference observed between rich and poor countries is also observed between urban (77%) and rural (33%) areas from developing countries and as well between rich and poor people living there following the inequities of wealthy distribution.
4.06.5.3.3 Situation at the national level The sanitation coverage as percent of the population with service per country is presented in the map of Figure 6 for the year 2004. Annex 3 contains a table with countries with less of 60% of the total, urban, or rural population.
4.06.5.3.4 Low-income countries sanitation specificities Sanitation in developing countries is quite a complex issue, because the lack of it is combined with other several problems, some of which are geographically described on the Maps 1–8
Table 7
Sanitation coverage per region for 2004
Region World Developed regions Commonwealth of independent states Developing regions Northern Africa Sub-Saharan Africa Latin America and the Caribbean Eastern Asia Southern Asia South-eastern Asia Western Asia Oceania
Coverage as % of the population Total Urban Rural 59 80 39 99 100 98 83 92 67 50 73 33 77 91 62 37 53 28 77 86 49 45 69 28 38 63 27 67 81 56 84 96 59 53 81 43
Coverages below 60% are highlighted in red and those above 80% in blue. Data from WHO–UNICEF (2006) Meeting the MDG Drinking Water and Sanitation Target: The Urban and Rural Challenge of the Decade. Geneva: WHO and UNICEF.
from Annex 4. By analyzing these maps, the following conclusions may be drawn: 1. Several low-income countries are located in arid or semiarid regions; thus, besides sanitation problems, they face the problem of water scarcity. 2. Many of the areas under greatest stress (where people are already overexploiting rivers by tapping water that should be reserved for environmental flows) coincide with areas that are heavily developed for irrigation to provide water for food, that is, mostly in developing countries. 3. Water withdrawal for agriculture is mainly performed in developing countries as a result of low water availability and the high dependence of agriculture. 4. Areas where poverty and hunger are prevalent coincide with areas lacking sanitation. 5. In the future, it seems that the situation may worsen as water availability will decrease in the countries already experiencing water-related problems, including lack of sanitation. As result of the past and present situations, sanitation has different aspects on developing countries that cannot be described simply using the percent of population-covered index. In the following, some of these aspects will be described. Basic sanitation versus sanitation. Providing services for excreta management in poor rural or urban areas is frequently known as basic sanitation. Thus, it has to do with excreta management rather than with sewerage and wastewater treatment plants (Box 4 and Figure 7). The quality of the service is frequently associated with peoples’ economic level, and thus, is
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Pacific Ocean
Equator Pacific Ocean Population using improved sanitation, in percentage Less than 50% 50 to 75% 76 to 90% 91 to 100% No data
Equator Atlantic Ocean
Indian Ocean
Source: World Health Organization
Figure 6 Sanitation coverage per country in 2004 (with information from WHO–UNICEF, 2006).
also a sign of status. Another aspect to consider is that the lack of basic sanitation frequently is associated with lack of water. LeBlanc et al. (2008) highlights that research and experience suggest the following hierarchy of risk to human health:
•
‘‘living in a dense community without basic sanitation4(is more risky thany) irrigation of crops with untreated, pathogencontaminated wastewater4use of untreated, pathogen-contaminated excreta or wastewater sludge on soils4use of untreated, pathogencontaminated animal manures on soils4use of treated manures, wastewater, or biosolids on crops4use of these treated materials in accordance with strict modern regulations that address heavy metal and chemical contaminants.’’
•
•
• • •
Differences on sanitation services. Possibly, one of the aspects that contributes the most to render sanitation in developing countries a challenge is the variety of needs and circumstances arising from social differences. As shown in Figure 8, for instance, poor people not only are less served but also the quality of the services is lower. One of the deepest disparities is between urban and rural areas as for the former the coverage is twice as much than for the latter in developing countries. Traceable differences in sanitation services have been reported as well among indigenous and nonindigenous people and minorities such as castes and women (Box 5). Among these differences, the following common challenges can be identified:
•
The need to provide the service in poor areas with large population increases.
•
For urban areas, a very fast service demand growth in slums that are spread out in cities, have high population density, and there is no land to place the infrastructure. For rural areas, the need to assist a population frequently dispersed and hence at higher cost. The need to fund projects combining liquid and solid waste collection and treatment infrastructure. The need to develop new or different management structures to provide services in social and political complex areas. The need to include health education and awareness programs on sanitation projects. The need to use public funding to provide services that are to be subsided. The existence of regions having high income where services can be provided in a similar way to developed countries.
Sanitation versus wastewater treatment. As described previously, sanitation coverage does not necessarily result in wastewater being treated or safely disposed of. To illustrate this, figures for the situation in some developing countries are provided. Two comments on this figure are that (1) it is really difficult to find data on wastewater treatment, notably for the Asian and African regions and (2) although there should not be a full correspondence between the sanitation coverage and the wastewater treatment – as some people are served using basic sanitation facilities – the figures should not be as different as they are for some countries. In Latin-America, for instance, although the sanitation coverage was 78% in 2006, only 18%
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Annex 3 Classification of countries per range of sanitation coverage, with information Total
Urban
Sanitation Coverage o 20% Ghana Guinea Cambodia Burkina Faso
Comoros
Sao Tome and Principe
Congo Congo, Democratic Republic of Coˆte d’Ivoire Gabon Guinea-Bissau Haiti India Lao People’s Democratic Republic Lesotho Liberia Madagascar Mauritania Micronesia, Federated States of Mozambique Namibia Nepal Sao Tome and Principe Solomon Islands Somalia Sierra Leone Sudan Timor-Leste Togo Sanitation coverage 4 40 but o 60% Azerbaijan Afghanistan Belize Angola Bolivia Bangladesh Botswana Benin Cameroon Bolivia Cape Verde Botswana China Burkina Faso Equatorial Guinea Burundi Gambia Cambodia
Continued
Total
Urban
Rural
Indonesia Kenya Kiribati
Cameroon Comoros Central African Republic Coˆte d’Ivoire
Iraq Kazakhstan Kyrgyzstan
Congo, Democratic Republic of the Equatorial Guinea Ethiopia Guinea-Bissau Haiti India Kenya Kiribati Korea, Democratic People’s R. Liberia Madagascar Mali Mauritania
Marshall Islands
Rural
Cape Verde Solomon Islands Togo Micronesia, Federated States of
Ethiopia Niger Chad Eritrea Sanitation coverage 4 20% but o 40% Afghanistan Chad Angola Congo Bangladesh Eritrea Benin Gabon Burundi Ghana Central African Guinea Republic
Annex 3
Azerbaijan Belize Brazil China El Salvador Lao People’s Democratic Republic Lesotho Mongolia Nepal Peru Senegal Timor-Leste Yemen
Korea, Democratic People’s R. Kyrgyzstan Maldives Mali Mongolia Nicaragua Nigeria Pakistan Papua New Guinea Rwanda Senegal Swaziland Tajikistan Tanzania, United Republic Uganda Vanuatu Yemen Zambia Zimbabwe
Mozambique Namibia Nicaragua Niger Nigeria Somalia Rwanda Sudan Sierra Leone Swaziland Tanzania, United Republic of Uganda Zambia
Maldives
Mexico Moldova, Republic of Morocco Palau Panama Pakistan Papua New Guinea Philippines South Africa Tajikistan Turkmenistan Vanuatu
Viet Nam Venezuela Zimbabwe
From WHO–UNICEF (2006) Meeting the MDG Drinking Water and Sanitation Target: The Urban and Rural Challenge of the Decade. Geneva: WHO and UNICEF.
of the wastewater was treated (CONAGUA and WWF, 2006). To give an idea of the situation in other regions, for the year 2004, when the Latin America and the Caribbean region reported a treatment capacity of 14%, this was of the order of 35% for Asia and nearly 0% for sub-Saharan Africa (WHO/ UNICEF, 2000; Figure 9).
4.06.5.3.5 Sanitation Costs Colombia Djibouti Egypt Fiji French Guiana Gambia Guyana Honduras Indonesia
According to Lenghton et al. (2005), the amount of money needed to fulfill the sanitation MDGs ranges from US$24 billion to US$42 billion representing, in mean conditions, an annual average investment of US$2.2 billion. To put these figures in perspective, the above-mentioned authors mention that each year Europe and the United States spend US$17 billion on pet food and Europe spends US$11 billion on ice cream. The overall cost estimation of the current water and sanitation deficit is of the order of US$170 billion, equivalent to 2.6% of developing countries’ gross domestic product (GDP). For each US$1 invested for sanitation, the economic
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Annex 4
No data
Upper middle income countries ($3,056 - 9,386)
High - income countries ($9,386 or more)
Low middle - income countries ($766 - 3,056)
Low- income countries ($766 or less)
Map 1 Economic income per country, with information from World Bank 2009.
No data
<10
10−25
25−50
50−75
>75 %
Map 2 People living at under 2 USD/day, UNDP, 2006 with data from http://earthtrends.wri.org/povlinks/index.php (Continued )
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No data
<500
500−1000
1000−1700
1700−5000
>5000
Map 3 Renewable water resources (surface and ground water) per inhabitant for 2005, with data from: FAO-Aquatat, 2009 http://www.fao.org/nr/ water/aquastat/globalmaps/
No data
<10
10−25
25−50
50−75
>75 %
Map 4 Water stress or water use intensity index (surface and groundwater withdrawal as percentage of the total renewable water resources) for 2001, with information from http://www.fao.org/nr/water/aquastat
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No data
<5
5−10
10−20
20−45
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>40 %
Whit data from 2001 Map 5 Surface water and groundwater withdrawal for agricultural purposes as percentage of the total actual renewable water resources for 2001, with information from http://www.fao.org/nr/water/aquastat/globalmaps
No data
<5
5−15
15−25
25−35
>35−50 %
>50 %
Map 6 Prevalence of undernourished people as percentage of total population for 2002–2004, with information from http://www.fao.org/nr/water/ aquastat/globalmaps (Continued )
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Improved sanitation prevalence 75−100
0−50
50−75
20−40
Note: Data on prevalence of improved sanitation are for 2000. Diarrhea Data on prevalence of diarrhea are for various years, prevalence 1991−2000, and indicate prevalence in two weeks (%) before may vary by season. Because country surveys were administered at different times, data are not comparable across countries.
10−20 0−10
with data from FAO-AQUASTAT, 2007 Map 7 Prevalence of diarrhea and improved sanitation 2000 With information from: United Nation Children’s Fund Programme and The Joint Monitoring Programme Lenghton et al. (2005) UNPD Earthscan.
Vistuta Oder Elba Rhine Meuser
Dnleper Don
Kura
Darya
Tigris & Euphrates
Colorado Rio Grande Grande de Santiago Balsas
Syr
Panuco
Amu Darya
Indus
Yangtze Ganges
Narmada Tapti Krisna Volta
Jubba
Mahanadi
Huang He Hong Chao Phrya
Godavari
Limpopo Orange
<500
500−1000
1000−1700
1700−4000
4000−10000
<10000
No data
Map 8 Projected annual renewable water supply per person by river basin for the year 2025. With information from From: Water Resources eAtlas, 2007 http://earthtrends.wri.org/pdf_library/maps/2-4_m_WaterSupply2025.pdf
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Box 4 Some challenges to provide basic sanitation in low-income countries * *
* *
Open defecation is practiced by 48% of the population in Southern Asia and 28% in sub-Saharan Africa. In Ouagadougou, the capital of Burkina Faso, the access to sanitation facilities is 53% while the figure for the country is 15.6%, a figure that reduces to only 10% for rural areas (Paskalev, 2008). In Yaounde´, Cameroon’s capital with 2 000 000 inhabitants, the available facilities for most people (88%) are external and in shared proprieties (Figure 7). Basic sanitation and sanitation figures reported are not the same. For instance, for Cote d’Ivoire, a coverage of 45% is reported for rural areas, but, in fact, 36% refers to basic facilities and only 9% to adequate systems (Angoua, 2008).
Percentage of people covered (%)
50 40 30 20 10 0 Flush toilets indoor, 75 l
External latrine, 50 l
Common latrine, 25 l
Private latrine, 20 l
Common latrine, 20 l
Other
Figure 7 Type of toilets used in Yaounde indicating the amount of used water per day and inhabitant: estimation for 2007. Data from Mfoulu N (2008) Cameroon. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 169–179.
Flush toilet
Pit latrine
No facility
100 80
%
60 40 20
Colombia, 2005
Kyrgyzstan, 1997
Namibia, 2000
Peru, 2000
Poorest 20%
Richest 20%
Poorest 20%
Richest 20%
Poorest 20%
Richest 20%
Poorest 20%
Richest 20%
Poorest 20%
Richest 20%
0
Zambia, 2001−02
Figure 8 Type of facilities provided for the richest and poorest quintiles in some countries. Data from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan.
return would be between 3 and US$34, depending on the region and the type of technologies used (WHO–UNICEF, 2004). Studies performed in Egypt and Peru showed that just providing access to flush toilets reduced the risk of infant death by 57–59% (Lenghton et al., 2005).
4.06.6 Wastewater Management Systems Even if sanitation represents an economic benefit, its cost is still important to societies in which this is not the only
requirement. Therefore, it is useful to combine options that involve building infrastructure with others that do not (such as washing or cooking produce that has been irrigated with polluted water) in order to improve health conditions while the sanitation services can be gradually provided. Such an approach is described in WHO (2006). In the next sections, options to build up wastewater management systems are reviewed. A wastewater management system (WWMS) is understood in this chapter as the combination of one or several of the following components: (1) basic sanitation facilities or toilets; (2) wastewater collection systems (sewers) or
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Box 5 Women and sanitation (with information from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan.) One explanation for the low effective demand for sanitation is gender inequality. Women tend to place a higher value on household toilets than do men for a number of reasons, among them privacy, cultural norms, care-giving responsibilities, and the risk of sexual harassment and assault. In addition, the unique sanitation needs of women and girls (e.g., during menstruation and during and after pregnancy) receive little recognition when discussions about sanitation and hygiene occur. Yet, the limited political and personal power of women in many developing countries means that some of sanitation’s strongest advocates are virtually absent from decision making and priority-setting processes.
% Sanitation
% Wastewater treatment
100 80 60 40 20
Yemen
Tunisia
Syria
Panama
Pakistan
Morocco
Mexico
Libya
Lebanon
Jordan
Iran
India
Guatemala
Ghana
China
Brazil
Algeria
0
Figure 9 Percentage of sanitation coverage and treated wastewater. Data from CONAGUA and WWC (2006) Regional Document for the Americas Prepared for the 4th World Water Forum. Ciudad de Me´xico, Mexico, 16–22 March. Vienna: UN, Hashimoto (2009, personal communication), Wikipedia (2009), and Bahri A (2008) Water reuse situation on the Middle Eastern and North African countries. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs, pp. 27–48. London: IWA Publishing.
excreta extraction mechanisms; (3) wastewater treatment plants; (4) sludge management and disposal units; and (5) wastewater disposal or reuse facilities. Before presenting these components in detail, the two options in which they can be managed (centralized or decentralized) are discussed. Conventionally, to handle wastewater, sewers connected to wastewater treatment plants have been used. This is known as a centralized system and is a well-mastered and well-managed technology approach applicable to cities, provided funds for its construction and operation are available. In terms of operation, centralized systems are often cheaper and easier to handle than decentralized ones. For isolated slums and dispersed rural areas and even for cities where new sewerage systems is too costly, it is advisable to use decentralized wastewater management systems. In these, sewers of reduced size result in a lower capital cost (around 30%) due to the smaller diameter and length of the used pipelines. In addition, they offer the following benefits (Lenghton et al., 2005; Correlje and Schuetze, 2008): (1) they allow investments to be made stepwise, in line with available funds, local development, and population growth; (2) they are used in smaller areas of service that are easier to manage; (3) they allow the use of different technologies to provide services to different socioeconomic groups; and (4) they facilitate the reuse of water on-site. Nevertheless, all these advantages need to be assessed in practice, as they cannot be taken for granted
universally. As for many water utilities, decentralized systems represent a higher number of systems to manage, which is difficult and complex; to overcome this limitation, centralized management of decentralized systems is recommended. This way it is possible to ensure high performance and reliable operation, reduce costs, and also ensure the need for specialized operators (Hughes et al., 2006).
4.06.6.1 Basic Sanitation Facilities From a technical point of view, there are four important components to consider when providing a basic sanitation service: (1) the type of toilet, (2) the storage facility for feces which frequently are associated to the toilet, (3) the way in which feces are extracted from the pit, and (4) their further management. This section deals with the first two components. Their main characteristics are discussed here; for design, it is recommended to consult specialized books. A good option to begin with is the United Nations Environment Programme (UNEP) website (see section titled ‘Relevant websites’).
4.06.6.1.1 Traditional latrines Latrines are the most widespread type of on-site sanitation facility. They are used in rural settings and deprived areas in cities. They consist of a makeshift pit dug in the ground and
Safe Sanitation in Low Economic Development Areas
generally covered with any material (a wooden, plant, or metallic cover, whichever is available). When latrines are full they can be emptied (this is unpleasant and has an associated cost) or closed to build another one (this requires the availability of land).
4.06.6.1.2 Ventilated improved pit latrine These latrines, instead of having a single vault, are made up of a shallow pit divided into two 1–2 m3 vaults. Their major advantage is that they are a permanent facility due to the alternate use of each pit. The name comes from the inclusion of a properly designed pipe allowing ventilation, which also requires a screen to avoid the accumulation of flies. The pit cover is made of precast concrete, wood, palm leaves, or metallic material, and is removable. Emptying is performed manually in low-income areas, but can be done mechanically every 3–4 years. The ventilated improved pit (VIP) latrine with multiple pits can be built for collective use, such as in schools, markets, fueling stations, and administrative buildings (Mamadou, 2008).
4.06.6.1.3 Septic tank The septic tank is commonly used as primary treatment in rural areas, low-income urban settings, isolated households, or on sites where soil is not suitable for the installation of sewers (Jime´nez and Wang, 2006). They are built where a constant water supply is available and are used to partially treat domestic wastewater and to digest the settled sludge. They remove around 50% of the organic matter and suspended solid content in 2–4 days. For sludge digestion, 0.5–1 year is required; during this time, sludge is mineralized and its volume is reduced. Septic tanks are made up of a series of communicating chambers. They must be water sealed to avoid underground infiltration and are built using bricks, mortar, or concrete. A variation of the septic tank is the Imhoff tank, having the advantage of a shape that allows the removal of suspended solids and the control of foul odors in a better manner. Septic tanks need to be periodically cleaned (1–2 times per year, leaving 20% of the mature sludge as inoculum for digestion). This represents an additional cost that cannot always be afforded by poor people. Septage (the slurry taken out of septic tanks) is sent to wastewater treatment plants or treated separately. To treat septage, lime is frequently added until a pH of 12 is reached, over a period of 30 min (Jime´nez and Wang, 2006). Effluents from septic tanks are discharged into trenches for subsoil infiltration or diverted to the sewerage system (when available). Septic tanks are widespread sanitation systems but are often responsible for environmental pollution due to poor purification effects and leakages notably affecting groundwater.
4.06.6.1.4 Composting toilets Composting toilets are characterized by the separation of urine and feces. For this reason, they are also referred to as urine diversion (UD) toilets. They are constructed with two vaults or chambers. When the first vault is full, the pedestal is moved over to the second vault, and the first hole is closed. When the second vault is full, the first vault is emptied and so on. The urine is diverted to a soakaway. In comparison to VIP
173
latrines, they have a lower cost associated with emptying the pits (Snyman, 2008). Urine is collected in small cans (10–20 l) and can be used to enrich the soil after a stabilization period of 30 days. Feces are treated using an aerobic composting process. To control odors and to assist in the mineralization of feces, materials, such as ashes or pieces of wood, are used daily to raise the pH. The pathogens in fecal matter are inactivated over time through the drying process so they can be safely removed by the owner at no cost to the municipality. Once the sludge is digested, disinfected, and removed, it is used as fertilizer. UD toilets are seen as a viable option for rural applications. The main reasons are that they are cost-effective and, since the rural community is accustomed to the use of manure, the UD toilet is socially acceptable. However, its use in periurban areas is more problematic. The emptying of the vaults requires large-scale programs for which small businesses can contribute to the emptying of tanks (from UD or VIPs) either manually, using appropriate safety equipment, or by the use of a tanker. The disposal of the fecal matter in periurban areas is challenging due to the lack of land. If space allows, fecal sludge is buried on-site. Where this is not feasible, the sludge is blended into the waterborne system. This frequently leads to the complete overloading of the wastewater treatment plant (Snyman, 2008). There are several options of composting toilets (see section titled ‘Relevant websites’).
4.06.6.1.5 Pour-flush toilets Pour-flush toilets have been developed based on traditional flush toilets, which rely upon a water seal to perform cleansing and to control odors and insect infestations. The system works via a manual flush, where 2–3 l of water are poured into the toilet. The water, urine, and excreta are collected in an anaerobic chamber, which works similarly to a septic tank. The chamber needs to be periodically emptied and the partially treated wastewater needs to be disposed of, normally to land (Hughes et al., 2006). In the context of water-scarce areas, a very interesting option is combining graywater reuse with basic sanitation using pour-flush toilets. This concept was developed by United Nations International Children’s Emergency fund (UNICEF) on a system called the Wise Water Management scheme (Godfrey et al., 2007). This system was conceived to provide both water supply and sanitation services for water-scarce areas and can be used for both rural and lowincome urban areas. It was conceived in Madhya Pradesh, India, a densely populated and poor area. The WWMS uses groundwater as the primary source of water and also includes rainwater harvesting, used to dilute groundwater when polluted with fluoride to reduce its content for human consumption (Figure 10). First-use water is employed for cooking, handwashing, and bathing. Water from these two activities is recovered and properly treated in a sand filter to be used for toilet flushing and kitchen garden irrigation. The graywater reuse system can be installed independently of the rainwater harvesting system. By matching water demands, in quantity and quality, to different conventional and nonconventional water sources, the WWMS increases water availability by nearly 60%. Sanitation using low-consumption reused water flush
174
Safe Sanitation in Low Economic Development Areas
Cooking and human consumption
Fresh water source Ground or surface water
House cleaning activities
Pluvial water
Bathing
Gray water treatment
Hand washing
Reclaimed water Kitchen garden
Soil disposal
Toilet flushing
Leach pit
Soil disposal
Soil disposal
Figure 10 Flow diagram of the Wise Water Management Scheme.
Box 6 Poor people have a globalized attitude towards excreta management As described for Senegal by Ba (2008), in most poor areas of the developing world, water from baths and in some cases from showers are routed to septic tanks from which the effluent is sent to infiltration wells or trenches. Kitchen and laundry water is generally poured directly into the street, discharge areas in the wild, a well, a nearby river, or riverbed. Wastewater and noncollected solids are also frequently mixed creating breeding sites, odor problems, and development of flies.
toilets has proven sustainable under the prevailing local conditions and has eradicated open defecation.
4.06.6.1.6 Additional recommendations to set up basic sanitation facilities One important aspect to keep in mind when selecting the technology is that facilities need to be operational and, to achieve this, there is a need to sustain them under operation from the economical, technical, and cultural perspectives. Investment costs are linked to the type of sanitation system selected, the construction materials, and labor. Frequently, to reduce costs, cheap materials and the users are employed to build the facilities. However, this may result in failures, as cheap material frequently means low quality and the users are not people experienced enough, even if trained. It is thus preferable to invest in good and durable material and to use experienced workers. In India, for instance, sanitation programs using professional well-trained masons are being implemented in which the same masons for whom sanitation is a source of income become at the same time sanitation promoters. Norms and institutional capacity to provide basic sanitation constitute another weak link in the complex chain needed to implement and provide services. How to build institutions, policies, and human resources to provide successful sanitation services is better known in high-income countries
than in developing ones. Each country/region needs to look for the proper way to solve their problems. Finally, concerning basic sanitation, it needs to be considered that in several places, providing basic sanitation means to change open defecation habits and to handle domestic solid wastes (Box 6). It means as well to properly dispose of the toilet paper.
4.06.6.2 Toilets Under this section, only the toilets using less water or none at all are described as compared to the others (pour flushing toilets using 415 l of water is a well-known technology widely spread commercially). Concerning these toilets, one aspect to highlight is that even if convenient from the point of view of the used water, care must be taken when designing treatment plants as wastewater will be not only lower in volume but also highly concentrated, notably in terms of its organic matter content.
4.06.6.2.1 Water-saving toilets These toilets are based on the same working principles as common flush toilets but they are specially designed to fully operate with less water (6–8 l). In such toilets, it is possible to select either a full flush (with 4, 6, or 9 l depending on the model) for solids or a half flush (2–4.5 l) for liquids.
Safe Sanitation in Low Economic Development Areas
These toilets are also available with separate drainage for urine to reduce the impact of nutrients and pharmaceuticals on the sewage and to facilitate the reuse of urine as a fertilizer. However, most water-saving toilets available on the market are designed to be connected to typical drainage systems. There are several technological options on the market, some of which use a vacuum to transport feces at a much higher cost. The investment cost for low-volume toilets is comparable to high-volume toilets. However, dual flush toilets may cost more than common ones (nearly double). The installation of water-saving toilets must be stimulated by education (e.g., in the form of campaigns to raise awareness concerning watersaving issues), water metering, and pricing. Water-saving urinals, using 1–3 l, are also available (Correlje and Schuetze, 2008).
4.06.6.2.2 Toilets not using water The idea of dry toilets is not new. They have been used for thousands of years in East Asia (China, Japan, and Korea). Dry toilets are available as industrial prefabricated products and can also be constructed in local workshops; however, knowhow for its good operation and to avoid foul odors is required. Investment, construction, or installation costs vary significantly and depend on the specific system and design. The cost ranges from low investment for simple dry toilets to comparatively high cost for industrialized composting toilets. Due to the large size of the storage and composting chambers, these toilets require a large space underneath; if this is not possible, then they need to be regularly emptied and feces need to be transported to treatment facilities. User acceptance depends on cultural background and awareness. Generally, people who are already using flush toilets do not readily switch to dry toilets because the image of dry toilets is less attractive than that of flush toilets.
4.06.6.3 Sludge Extraction from On-Site Sanitation System Equally important as the type of on-site sanitation system selected is the provision of all the services associated. Past experiences (Water Decade, 1980–90) have shown that massive sanitation infrastructure provision without a proper planning of the whole scheme can be a complete failure (Kone´, 2010). Besides the technical aspects that are discussed later, the most worrying aspect is the lack of financial, institutional, and regulatory framework in most of the developing countries to establish the network required. Management of on-site sanitation infrastructure comprises on-site sanitation systems emptying, fecal sludge haulage, treatment, and safe reuse or disposal (Kone´, 2010). Fecal sludges refer to sludge collected from on-site sanitation systems such as latrines, nonsewered public toilets, or septic tanks. The criteria to select an extraction method – a task that is never pleasant – depend on (1) the TS content and (2) the funds available. Sludges with less than B2% TS, such as those produced in septic tanks, can be pumped; but, for the rest of facilities producing all sludge with 10% TS, pits need to be emptied using cesspit trucks or manually by laborers (Kone´, 2010). Even though when mechanically emptied and water is used for toilet cleansing, 20–50% of the contents in the lower pit part need to be manually emptied to extract the thicker
175
sludge. The use of mechanical equipment allows carrying away the sludge several kilometers for disposal on controlled sites or on treatment facilities, but this is often expensive and needs proper equipment and skilled laborers. In contrast, when sludge is manually emptied, this is deposited in nearby lanes or on open spaces representing a source of risk. According to Kone´ (2010) 30–50% of the on-site sanitation facilities from West African countries are emptied manually. In addition, in almost every developing country, fecal sludge collection and haulage are conducted by private entrepreneurs. However, their important role and responsibilities as key stakeholders are not yet fully recognized and legalized (Kone´, 2010).
4.06.6.4 Sewerage Systems 4.06.6.4.1 Small sewers In many low-income areas, the sanitation problem begins with the lack of sewerage. One option is to build sewers of small extent coupled with on-site sanitation systems. Sewers carry the treated effluent to disposal (usually to soil for infiltration, to irrigation canals, or into water receptors), to wastewater treatment plants, and/or to reuse sites located within a short distance. As these sewers frequently convey partially treated wastewater (such as septic tank effluents), they are designed for self-cleaning using a high wastewater velocity and/or a steep slope. This option is applicable for rural areas or urban ones where adequate land is available. Another option is to use simplified sewers. These are recommended where an uncertain population increase is occurring, as normally happens in periurban areas or slums. Small sewers are built to reduce the infrastructure and maintenance costs, as well to allow high operational flexibility. Inspection chambers such as manholes are replaced by inspection cleanout. The life expectancy of such sewers is in the order of 20 years rather than the 30 years quoted for conventional sewers. Such sewers are short and shallow (Hughes et al., 2006). One example of simplified sewers are condominial ones in which pipelines are laid through housing lots instead of on the side street, in a way that allows isolated and stepwise construction (UNEP, 2002). Condominial sewers were developed in the 1980s in Brazil with the aim of extending sanitation services to low-income communities. This technology has now become a standard sanitation solution for some urban areas in Brazil, irrespective of income levels. Condominial sewers reduce the per capita costs of service by replacing the traditional model of individual household connections to a public sewer with a model in which household waste is discharged into branch sewers, and eventually into a public sewer through a group (or block) connection (Watson, 1999 cited in Lenghton et al., 2005).
4.06.6.4.2 Conventional sewers These are structures that are bigger and deeper than those previously discussed. Details for design can be found in conventional literature on sewers.
4.06.6.4.3 Pluvial sewers Many developing countries are located within regions subject to tropical storms, or in areas where there are only two seasons per year: wet and dry. Therefore, urban hydraulic infrastructure
176
Safe Sanitation in Low Economic Development Areas
needs to be designed accordingly to have sewers that can handle large peaks of stormwater and the normal wastewater flows (wastewater treatment plants should also be capable of dealing with the varying wastewater characteristics in quantity and quality, at least in large cities). Sewers in tropical areas produce a high amount of sediments to be disposed off, which turns out to be a peculiar and difficult-to-solve problem not frequently commented upon in specialized literature but that needs proper methods to extract sludge and handle it. In addition, when conveyed in sewerage systems, stormwater must be treated in treatment plants at the same time as wastewater; but, if transported separately, it can be discharged to surface water or into wells for groundwater infiltration receiving treatment in soil. In this case, it must be kept in mind that stormwater quantity and quality are determined by rainfall, catchment processes, and human activities, which cause its flow and composition to vary in space and time. Normally, for the first rains of the year, stormwater has higher suspended solids, heavy metal content, and bacterial numbers than nontreated wastewater, and lower dissolved solids, nutrients, and oxygen demand than secondary-treated sewage effluent.
4.06.6.5 Wastewater Treatment Wastewater treatment is the typical method applied for sanitation, and is the predominant option used in developed countries for that purpose. Although it cannot be considered a caveat for all the negative impacts produced by wastewater, it is still a very important option, and, in many cases, the only one. There are several steps to treat wastewater. The primary step basically serves to remove easily decantable and floating solids. The secondary one, generally a biological process, is used to remove biodegradable (mostly) dissolved suspended material. The tertiary step is used to refine the quality of the effluent produced by a secondary treatment. It may have different purposes, most commonly being the removal of nutrients (N and P). As the treatment steps were conceived following treatment needs, in practice, they are usually implemented in separate tanks or in well-defined sections of wastewater treatment facilities; however, it is possible to use compact processes eliminating physical separation among steps and thus reducing costs (Jime´nez, 2003). Wastewater treatment plants are not common facilities in low-income countries. In contrast to developed countries, in developing ones, the sanitation figure (50% according to WHO–UNICEF (2006)) does not include the treatment of wastewater, which barely reaches 15% (US-EPA, 1992). Moreover, when available, the treatment merely consists of a primary step or including eventually a secondary step that is not always properly functioning. In many developing countries, the main issue concerning treatment is still the proper disposal of feces, particularly in low-income urban or rural areas. This, combined with a high content of pathogens in wastewater, sludge, or fecal sludge, implies the need to properly select the treatment process in order to effectively control disease dissemination. In general, coupling any kind of secondary wastewater treatment process (biological or physico-chemical) with a filtration step before disinfection will considerably reduce the pathogen content. However, this is rarely feasible for economic reasons and therefore it is sensible to consider the use
of other technologies alone or combined with other type of intervention methods to build up a multiple barrier system to control wastewater risks (Jime´nez, 2009b). In the following sections, guidance will be provided to support the selection for treatment options, based on the type of pollutants.
4.06.6.5.1 Conventional pollutants treatment To address problems caused by suspended solids, organic matter, nutrients, and fecal coliforms, there is a wide variety of available technologies supported by literature and practical results. Their affordability in economic terms and the suitability of the processes for local conditions are among the important aspects to consider for developing countries. It is beyond the scope of this chapter to provide a full description of treatment technologies for conventional pollutants, which can be found elsewhere in the literature. Table 8 shows the removal of pollutants by different processes so that it is possible to identify those acting upon the same type of pollutants.
4.06.6.5.2 Pathogens treatment Table 9 presents organisms’ removal or inactivation achieved by different wastewater treatment processes. This table is a guide for selecting a process. However, to design complete treatment schemes, the operating conditions need to be properly selected as well as the pre-and post-treatment. Table 9 differs from the one presented by WHO (2006) in showing the removal efficiency data for helminth eggs in terms of a percentage instead of log removal. This is because helminths eggs’ content is by far much lower and log units are meaningless. For developing countries, the removal of protozoa and helminths eggs is the main concern, considering their content and the occurrence of diseases caused by these types of agents. To remove protozoa, filtration is a good treatment option. Conditions used to remove Cryptosporidium oocysts – the targeted protozoan for developed countries – can be used as well to remove protozoa relevant to developing countries. Helminth eggs are not affected by conventional disinfection methods (chlorination, ultraviolet (UV) light, or ozonation); thus, they are first removed from wastewater using sedimentation, coagulation–flocculation, or filtration processes to be subsequently inactivated in sludge (Jime´nez, 2008a). Removal occurs because eggs are particles 20–80 mm in size. It is estimated that for contents of 20–40 mg l1 of TSS in treated wastewater, the concentration of eggs is around 3– 10 eggs l1, while for values below 20 mg l1 it is around 1 egg l1or less (Jime´nez, 2008a). However, for a process to be reliable, besides the removal efficiency attained, it is important for it to produce an effluent with constant concentration.
4.06.6.5.3 Emerging chemical pollutants The removal efficiency of emerging chemical compounds during conventional treatment can be found in Jime´nez (2009b). It is recommended that experimental tests be performed under laboratory conditions, prior to treatment selection. In the following, a description of main wastewater treatment processes is made, highlighting aspects that are relevant to developing countries, notably concerning their efficacy to control pathogens.
Table 8 Removal of pollutants by different wastewater treatment process that can be used to buildup a multiple barriers treatment scheme (with information from Jime´nez (2003); Jime´nez (2009), and Correlje and Schuetze (2008) Cost 1 Low Medium High Sophistication/complexity Low Medium High
Pollutant
Process
ONSS
PS
BT
BT + NR
CF
FI
NO
NO
NO
NO
NO
NO
Suspended solids Dissolved solids
Cl-D
UASB
LmP
NO
3
NO 3
UV-D
O-D
NPh
NO
NO
NO
NO
NO
NO
BOD
NO
TOC
NO
Volatile organics
NO
Heavy metals
NO
Nutrients
NO
Viruses∗
NO
Bacteria∗
NO
2
NO 2
NO
Protozoan∗
NO
21
NO
7
NO
11
7
NO
11
2, 10
Helminth eggs Pesticides
NO
NO
NO
NO
NO
NO
3,4,5
5
6
NO
9
NO
NO
UV-O
NO
NO
NO
NO
NO
NO
NO
NO
NO
Pp
Ads
3
NO
NO
NO
UF
NO
NO
NF
RO
NO
9
NO
11
11
MF
15
?
No on AC
9
2
9
2
9
2
11
12
No on AC
12
8
8, 12
Oz-O
NO NO
NO NO
Cl-O
NO
NO
8 NO
3
9 NO
Disinfection by products Chemical emerging pollutants
10
WT
NO
NO
21
SAT
8 20
19
?
12,13
12
14
5
16
NO
17
17
18
Processes: AC, activated carbon; Ads, adsorption; BT, biological treatment (any technology); BT+ NR, biological treatment with nutrient removal; CF, coagulation−flocculation (any technology) Cl-O, chlorine oxidation Cl-D, chlorine disinfection; FI, filtration; Clo, chlorine oxidation; Lmp, lime precipitation; MF, microfiltration; UF, ultrafiltration; NF, nanofiltration; NPh, natural photolysis; O-D, disinfection with ozone; ONSS, on-site sanitation systems; Oz-O, ozone oxidation; PS, primary sedimentation; Pp, precipitation; RO, reverse osmosis; UV-O, UV light oxidation; UASB, upflow anaerobic sludge blanket; SAT, soil aquifer treatment and river bank filtration; UV-D, UV-light disinfection; WT, wetlands. Low removal 1 2 3 4
Medium removal
High removal
Depending on the treatment level (primary, secondary, or tertiary). Depending on the type of technology used. Might increase the content. Mostly in biological secondary treatment plants; widely depending on the chemical composition of the pollutant; removal might represent only the transformation of the compound or its adsorption into. 5 Depending on the specific compound. 6 If coupled with chemicals. 7 Produce the pollutant as by-product or increase its value. 8 With low reliability. 9 For phosphorus. 10 Depending on the operating conditions.
11 12 13 14 15 16 17 18 19 20 ?
Noxious by-products can be formed. If there is no competition with organic matter (BOD or COD). Doses are several orders of magnitude higher than those used for disinfection. If granular carbon is used. High for nonpolar organic compounds with log KOW > 2 and when there is no competition with organic matter. Medium to high depending on the presence of cations and organic matter. High but not for low molecular weight uncharged compounds. Effective for several EC but not for carbamazepine, primidone, and iodinated X-ray contrast media. High for some EC, as it depends on the strength of solar irradiation removal will be different for different latitudes, or conditions. Can be enhanced with photosensitizers. Unknown or insufficient information ∗, Can be removed or inactivated. NO, not applicable for the pollutant.
178 Table 9
Safe Sanitation in Low Economic Development Areas Reduction or inactivation of different biological pollutants in wastewater
Treatment process
Log unit microorganisms removal
Removal (%)
Viruses
Bacteria
Protozoan (oo)cysts
Helminth eggs
Natural systems Waste stabilization ponds, WSP Wastewater storage and treatment reservoirs Constructed wetlands
1–4 1 to 2/4 1–2
1–6 1 to 3/6 0.5–3
1–4 1–2 0.5–2
90–100a, e, HR 70–95a, d, LR, g 90?a, e, L, R
Primary treatment Primary sedimentation Chemically enhanced primary treatment or advanced primary treatment Anaerobic upflow sludge blanket reactors, UASB Filtration
0–1 1–2 0–1 0–1
0–1 1–2 1–2 0–0.5
0–1 0.5–2 0–1 0–1
90a, LR 90–99a, 60–96a, 90–95
Secondary treatment Activated sludge þ secondary sedimentation
0–2
1–2
0–1
90–95a,
0–1 1–2
1–1 1–2 1–2
0–0.5 0–1
85–90c 95–100c 90c
Tertiary treatment Coagulation/flocculation High-rate granular sand filtration Dual-media filtration Membrane bioreactors
1–3 1–3 1–3 2.5 to 46
0–1 0–3 0–1 3.5 to 46
1–3 0–3 1–3 46
95–99a, 90–99a, 100c 100c
Disinfection Chlorination (free chlorine) Ozonation UV irradiation
1–3 3–6 1 to 43
2–7 2–6 2 to 44
0–1.5 1–2 43
0a, f, b 30–70b 0c
Trickling filters þ secondary sedimentation Aerated lagoon or oxidation ditch þ settling pond Slow filtration
e, HR e, LR
L, R
e, HR f, HR
a
Have been tested at full scale. From laboratory data. c Theoretical efficiency based on removal mechanisms. d Total helminth egg removal is only achieved when wetlands are coupled with a filtration step. e Tested with high helminth egg content. f Tested only with low helminth egg content. g Efficiency highly depends on size and operating conditions, notably the hydraulic retention time. LR, low reliability; HR, high reliability. Based on Shuval HI, Adin A, Fattal B, Rawitz E, and Yekutiel P (1986) Wastewater irrigation in developing countries: Health effects and technical solutions. World Bank Technical Paper No. 51. The World Bank, Washington; WHO (1989) Guidelines of the Safe Use of Wastewater and Excreta in Agriculture and Aquaculture. Prepared by D. Mara and S. Cairncross: Geneva: WHO. Von Sperling (2003, 2004); Rose (1999); Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO Leiden: Taylor and Francis Group; WHO (2006) Guidelines for the Safe Use of Wastewater, Excreta and Greywater, Vol. 2: Wastewater Use in Agriculture. Geneva: WHO. b
4.06.6.5.4 Slow filtration Slow filtration is recognized in water potabilization as an efficient method to control microbial pollution in rural and low-income communities. The few studies carried out on slow filtration of wastewater have demonstrated a removal range of 60–80% of suspended solids and 1–2 E. coli log, with coarse sand (Jime´nez, 2003). In rural areas, it may be coupled with absorption wells, irrigation reuse, or a soil aquifer treatment (SAT) system.
4.06.6.5.5 Waste stabilization ponds Waste stabilization ponds (WSPs) are shallow basins that use natural factors such as biodegradation, sunlight, temperature, sedimentation, predation, and adsorption to treat wastewater (Mara, 2004). WSPs are capable of removing organic matter with efficiencies similar to the activated sludge process and all kind of pathogens. They are easy to design and operate but require long retention times (several weeks). WSP systems
comprised several ponds connected in series. Lagoons are made through the shallow excavation of around 1–2 m, and they are frequently unlined to reduce investment costs. After a period of time, soil percolation and sedimentation form an impermeable barrier. If the water table is very high at the site, ponds need to be impermeable from the beginning. WSPs remove up to 6 bacteria log, up to 5 viruses log, and almost all the protozoa and helminth ova. To control Cryptosporidium spp., almost 38 days’ retention time is needed (Jime´nez, 2008). In developing countries with wet warm climates, the use of stabilization ponds is recommended if land is available at a reasonable price. For arid and semiarid regions, high evaporation rates limit their application as there is a net loss of water of 20–25% due to evaporation. This, in addition, increases the salinity of the effluent limiting its use for agricultural irrigation (Jime´nez, 2008). Sludge production in ponds is low but if extracted it needs disinfection as helminth ova remain viable in ponds for more than 9 years (Nelson et al., 2004).
Safe Sanitation in Low Economic Development Areas
WSPs can be coupled with aquaculture systems that are shallow ponds or wetlands where fish, duckweed, or aquatic vegetables are produced as is frequently done in Indonesia, China, and Thailand. Ponds can be used to produce only one crop such as duckweed that is used as food for the next pond where grass carp are grown. Different species can also be cultured in the same pond, as happens in nature. To operate the system, wastewater is applied to ponds at the required rate (estimated in terms of the organic load applied per hectare of ponds per unit time), and the organic matter and the nutrients contained serve as food for plant and animal production (Hughes et al., 2006). In order to avoid health problems, wastewater needs to be previously disinfected according to WHO guidelines (2006).
4.06.6.5.6 Wetlands Constructed wetlands are used to naturally remove organic matter, pathogens, and nutrients from wastewater through biodegradation, adsorption, or filtration in a similar way to WSPs. Nutrients are also removed by plant uptake and pathogens by competition and sun UV-light inactivation (Jime´nez, 2003). Wetlands are shallow ponds where aquatic macrophytes are planted in soil, sand, or gravel. There are three main types: surface-flow, horizontal-flow subsurface, and vertical-flow systems. Juncus spp. or Phragmites are commonly used plants but any local plant can be employed. Construction requires expertise and skilled labor. Once installed, operation is relatively easy. Wetlands remove nitrogen, phosphorus, and heavy metals. Up to 90–98% of thermo-tolerant coliforms, 67–84% of MS2 coliphages, and 60–100% of protozoa are inactivated or removed using hydraulic retention times of 4–5 days. In practice, pathogen removal is highly variable and depends on climate, type of wetland, and the kind of plant used. To completely remove helminth ova, it is necessary to couple wetlands with filtration, otherwise effluent with variable content may be produced. Breeding of mosquitoes and unpleasant odors can be a problem if wetlands are not operated correctly. Subsurface wetlands are used to avoid mosquito breeding (Correlje and Schuetze, 2008). Wetlands are a good solution for wastewater treatment in urban or rural areas where space is available; as a rule of thumb, 0.5–2.5 m2 per person is required for the treatment of graywater and 1–3 m2 per person for domestic wastewater. They are considered environmentally sound technology by UNEP for the treatment of graywater and stormwater urban runoff. They are used as secondary or tertiary treatment units, in which case, they treat effluents from septic tanks, anaerobic ponds, upflow anaerobic sludge blanket (UASB) reactors, or conventional wastewater treatment plants. Treated wastewater can be reused for agricultural irrigation, although its nutrient content is low. Wetlands have been used in Bangladesh and China to treat wastewater and to cultivate fish and ducks. In addition, they have the advantage of producing a low quantity of sludge.
4.06.6.5.7 Land treatment Soil can be used to treat wastewater by infiltration. It has a greater depollution capacity than water receptors, as there is no limit for the oxygen transfer needed for biodegradation.
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Land-based treatment is recognized as an environmentally sound technology by UNEP (2002) that has a low cost when used for primary effluents. Among its disadvantages is the high demand for land (Jime´nez, 2003). In the case of land treatment, depollution takes place in the unsaturated zone through biodegradation, adsorption, ion-exchange filtration, and precipitation. For the removal of organisms, in addition to predation and humidity, the temperature also plays a role. Heavy metals and trace organic compounds (such as emerging pollutants) are removed mainly by adsorption. To operate, wastewater is to be applied at specific rates; if pretreatment is needed primary sedimentation or sand filtration might be used (Brissaud and Salgot, 1994; Jime´nez, 2003; Bouwer, 2002). In developed countries, pre-treatment usually consists of a secondary treatment. Wastewater application occurs in cycles at a rate that depends on the soil infiltration characteristics. In a typical situation, the cycle involves 1 week of wastewater flooding where infiltration is reduced by organic buildup, and 1 week of drying where bacteria consume the organic matter and soil drying takes place. There are several types of land treatment options in specialized literature that can be consulted. For efficient functioning, hydraulic loads (29–111 m3 m2 yr1) and mass loads should be limited. To avoid aquifer pollution, application of wastewater (preferably partially treated) is restricted to sites where groundwater is a minimum of 3 m in depth. Applied as primary or secondary treatment, land treatment produces a consistently high-quality effluent (TSS o1 mg l1, organic carbon 3 mg l1, and total nitrogen 6 mg l1, with a phosphorus removal of almost 50% with minimal pre-treatment). As tertiary treatment, it removes 492% of BOD, 85% of COD, 100% of TS, 455% of detergents, 499% of ammoniacal nitrogen, 55% of total nitrogen, and 98% of phosphorus. Land treatment is effective for the removal and/or inactivation of helminth eggs, protozoa, bacteria, and even viruses (Jime´nez, 2003).Treated wastewater can be used for irrigation or any other use and can be collected on the surface or underground.
4.06.6.5.8 Reservoirs and water storage tanks Reservoirs or wastewater storage tanks can be used as well to treat wastewater. While wastewater is stored during the wet season to provide water for irrigation during the dry season, pathogens are removed or inactivated via sedimentation, UVsunlight inactivation, predation, and other similar processes, which also occur in WSPs. Nevertheless, the efficiency is lower. Procedures for designing wastewater storage and treatment reservoirs are detailed in Juanico´ and Milstein (2004) and Mara (2004). Reservoirs and storage tanks are easy to operate and maintain, and if considered as part of the irrigation system, they result in a low investment cost. However, they facilitate vector breeding if they are not well maintained and operated, and algal development in effluents may interfere with irrigation applications. Effluent storage reservoirs remove 2 4-log of viruses, 3 6-log of bacterial pathogens, and 1 2-log units of protozoan (oo)cysts. If treatment reservoirs are operated as batch systems with retention times over 20 days, the complete removal of helminth eggs can be achieved (Juanico´ and
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Safe Sanitation in Low Economic Development Areas
Milstein, 2004). In addition to large storage reservoirs, small storage ponds can be utilized for pathogen removal when used for urban agriculture irrigation as intermediate water storage reservoirs. Such reservoirs reduce the helminth ova content by around 70% (Keraita et al., 2008).
initial helminths egg content of 20–120 eggs l1, effluents with 3–10 eggs l1 are produced (Jime´nez, 2008). Other biological secondary treatment options include aerated ponds, oxidation ditches, and trickling filters. Much specialized literature exists describing the processes that are used to treat effluents before discharge into water bodies.
4.06.6.5.9 Upflow anaerobic sludge blanket The UASB is used to remove organic biodegradable matter. A UASB is a kind of attached system where microorganisms adhere to themselves, forming flocs. UASBs are considered as the most successful anaerobic process applied to treat wastewater due to low hydraulic retention time compared to other anaerobic processes thanks to the high density of biomass attained in the blanket (Campos, 1999). The reactor is designed to not only produce the biological reaction but also to sediment and filter suspended solids from wastewater. In addition, sludge retained in the bottom part of the reactor is anaerobically digested (Campos, 1999). The UASB produces better results when the wastewater has a high organic matter content. As by-products, it produces methane and partially treated sludge. The gas can be used as a source of energy, while the sludge remaining, after proper treatment to control the pathogen content, can be used to fertilize soil. UASBs remove 65–75% of BOD and COD and helminth eggs through filtration in the sludge blanket and through sedimentation. However, their efficiency with regard to the removal of helminth eggs is very variable. From wastewater containing 64–320 eggs l1, they produce effluents with 1–45 eggs l1 (60–96% removal). Therefore, UASBs are frequently coupled with other treatment process such as stabilization ponds or filtration to completely and reliable remove helminth ova and to inactivate other pathogens. Several stand-alone UASB plants or those coupled with WSP are currently under operation in Curitiba, Brazil. UASB reactors require careful design and operation to avoid bypasses (Campos, 1999). The construction, operation, and maintenance of improved anaerobic technology such as biogas installations require considerable expertise and skilled labor as well as space (Correlje and Schuetze, 2008). UASB reactors have a low capacity for tolerating toxic loads, need several weeks to start up the process, and require a post-treatment step.
4.06.6.5.10 Activated sludge It is the most common way to treat wastewater in developed countries. Compared to other secondary biological processes, activated sludge is effective for pathogen control as it removes 10% more than trickling filters. Both sedimentation and aeration play an important role in this. Sedimentation eliminates heavy and large pathogens, while aeration promotes antagonistic reactions between different microorganisms, causing their elimination. As a result of becoming entrapped within the flocs (which are subsequently sedimented), there is fairly good removal of small nonsedimentable microorganisms, such as Giardia spp. and Cryptosporidium spp., which remain concentrated within the sludge (Jime´nez, 2003). Helminths eggs are also removed, but due to continuous difficulties in achieving efficient and reliable sedimentation of suspended solids in secondary decanters, protozoan and helminths eggs may be found in effluents along with flocs. For an
4.06.6.5.11 Coagulation–flocculation This is a process that was almost abandoned for the treatment of municipal wastewater in the 1960–70s due to the high sludge production, which considerably increased the overall wastewater treatment cost. The introduction of new chemical products, in particular flocculants, combined with the possible reuse of treated effluent for agricultural irrigation and ocean disposal, has been instrumental in its reintroduction. Coagulation–flocculation removes helminths eggs while preserving nutrients and organic matter in contents suitable to grow plants. When this process is applied using low coagulant doses combined with a high molecular weight and high charge density flocculants, it is called chemical enhanced primary treatment (CEPT). If, a high-rate settler is used instead of a conventional settler, it is referred to as advanced primary treatment (APT). As a result, CEPT has a total hydraulic retention time of 4–6 h while, for APT, this is only 0.5–1 h. Among the coagulants that have been used, iron and alum compounds are the most common. APT removes 50–80% of protozoan cysts (Giardia, Entamoeba coli, and E. histolytica) and 90–99% of helminths eggs. From a content of up to 120 eggs l1, an APT can consistently produce an effluent containing 0.5–2 eggs l1. This process produces an effluent with a low content of suspended solids or turbidity, which leads to greater disinfection efficiency, either with chlorine or with UV light. Likewise, the process allows the use of sprinkler irrigation in high-tech countries or countries where water is scarce. The effluent quality is improved by the soil effect, and aquifers can be used as water supply storage (Jime´nez, 2003, 2008). APT and CEPT are useful in middle- and high-low-income countries on large urban areas as an economical alternative to an activated sludge process as the treatment cost for APT is one-third of this process when considering sludge treatment and disposal within 20 km. Coagulation–flocculation can also be applied as a tertiary treatment after a biological process. This is a very good method to remove enteric viruses (Jime´nez, 2003).
4.06.6.5.12 Rapid filtration Rapid filtration (at rates over 2 m3 m2 h1) is very efficient in removing protozoa and helminth eggs from wastewater, primary effluents, and biological or physicochemical effluents. It removes 90% of fecal coliforms, Salmonella, Pseudomonas aeruginosa and enteroviruses, 50–80% of protozoan cysts (Giardia, Entamoeba coli, and E. histolytica), and 90–99% of helminths eggs. Efficiency can be increased to easily reach 499% if coagulants are added (Jime´nez, 2008). For helminth ova removal, rapid filtration is performed in silica sand filters with 0.8–1.2 mm media size, a bed depth of at least 1 m and filtration rates of 7–10 m3 m2 h1. The helminth ova content
Safe Sanitation in Low Economic Development Areas in the effluent is constantlyo0.1 HO l1 in filtration cycles of 20–35 h for primary effluent (Jime´nez, 2003, 2008).
4.06.6.5.13 Disinfection The challenge for any disinfection method is that microorganisms respond differently. Efficiency depends on the disinfecting agent, the type and content of microorganism, the dosage, and the exposure time. The water matrix has as well a relevant influence, which becomes more important as its concentration and complexity increase. The most common disinfection processes for wastewater are chlorination, ozonation, and UV-light disinfection. 1. Chlorination. It is the most widely used process to control microorganisms. It is effective for the inactivation of bacteria, less so for viruses and protozoa, and not at all for helminth eggs. With regard to virus and bacteria, chlorine has inactivation efficiencies of up to 5–7 log. However, chlorine is a very reactive agent and, therefore, before attacking microorganisms, it reacts with many substances contained in wastewater, in particular with organic matter, hydrogen sulfide, manganese, iron, nitrites, and ammonia. As a result, chlorination is a process that, in order to be efficient, needs to be applied at the end of treatment schemes to avoid interferences. If, in treated wastewater, ammoniacal nitrogen and organic matter are still presented, chloramines and organo-chlorinated compounds are formed. These are compounds that increase cancer risks. Notwithstanding such risks, it is always preferable to chlorinate wastewater as microbial diseases have faster and often more dramatic health effects (Jime´nez, 2003). 2. Ozonation. Ozone is very effective at inactivating viruses and bacteria. It inactivates 3–4 log concentration units in a very short time, provided there is a low demand for oxidizing agents by wastewater. There is abundant information in the literature concerning the design and operation of the processes. Required ozone doses for several microorganisms are also available in the literature but, frequently, they are not affordable. As happens with chlorine, by-products generated during ozonation are a source of concern as many of them have been reported in the literature as toxic (Jime´nez, 2003). 3. UV light. Nowadays, UV-light disinfection closely competes with chlorination because it does not generate by-products that are too costly to remove from wastewater. Besides, compared to chlorination, UV light does not need storage facilities, does not imply the handling of hazardous chemicals, and uses very small-size treatment tanks as disinfection contact times are very small (in the range of seconds or minutes). Furthermore, due its simplicity of operation and high adaptive potential, it is suitable for rural and isolated communities.
4.06.6.6 Sanitation and Wastewater Treatment Costs Figure 11 presents estimated cost for different sanitation options, including from basic sanitation system to wastewater treatment plants. Simple services certainly are much cheaper to provide, but they do not necessarily represent what the society wishes to have due to the comfort level. As cost is an
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important barrier to spread sanitation services, one would expect that these data is a well-known parameter. Despite this, in many developing countries there are no reference costs, as exist in developed ones. As result of this situation, in many bids, costs are established using international data that do not necessarily reflect the local conditions (Table 10). Differences are due not only to build the sanitation facilities but also for the use of fuel and electricity, two important inputs to operate wastewater treatment plants. Sludge management and disposal (Figure 12) is another source of different affecting costs (Figure 12). Table 10 also shows that the cost of emptying onsite sanitation systems is not negligible.
4.06.6.7 Criteria for Selecting Wastewater Treatment Processes The selection criteria for wastewater treatment processes are presented in Table 11, emphasizing the needs of developing countries.
4.06.7 Wastewater Disposal versus Reintegration After treating wastewater, the next step is its disposal. Recently, some researchers have suggested (Asano, 2009) to use the term ‘dispersion’ instead of ‘disposal’ in order to change the perception of getting rid of used water, but this term has to an extent the connotation of wanting to dilute a problem. In this chapter, the term ‘reintegration’ is introduced in order to emphasize that water needs to be returned to the environment or used once again (reuse). By reintegrating the water to the environment, the responsibility of using it and then restoring it back to the environment in a proper way may be realized. As, well water can be reintegrated into the hydraulic cycles in which is been used by the society, thus reducing the negative impact of extracting water from the environment beyond the amount needed for ecological use (environmental flow). Water can be reintegrated to the environment by discharging it to the soil or into water bodies. In the following, different ways to reintegrate used water are discussed. This is followed by discussing the reintegration of water through reuse.
4.06.7.1 Soil Disposal or Reintegration of Used Water to Soil and to Groundwater Soil reintegration (disposal) consists of discharging treated or nontreated water into land. As discussed in the Section 4.06.6.5 the soil may act as a treatment step if a proper management is provided. The options to reintegrate treated wastewater into the environment are presented below. After discharging used water to soil, it will be evaporated, infiltrated, or will percolate to reach surface or groundwater bodies. The extent of each of these will depend on the soil and local conditions.
4.06.7.1.1 Leach drains They are used mostly for on-site sanitation effluents. They consist of a trench in which partially treated wastewater is discharged to allow its infiltration to the subsoil. The seepage in the trench allows uniform disposal of the wastewater over a given area. The leach drain is often filled with gravel or highly
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Safe Sanitation in Low Economic Development Areas Estimated cost per person in USD
10
Improved traditional practice
45
Simple pit latrine Ventilated improved latrine
65
Pour flush latrine
70
Septic tank latrine
160
Sewer connection with local labor
175
Connection to conventional sewer
300
Sewer connection and secondary wastewater treatment
450
Tertiary wastewater treatment
0
800
200
400
600
USD
800
Figure 11 Estimated cost for different options (with information from van de Guchte, and Vandeweerd, 2004).
Table 10 Comparisons of costs for wastewater treatment, diesel, and electricity in selected countries for the year 2008 (with information from LeBlanc et al. (2008)) Country
USD per m3 of wastewater
USD per 1000 l diesel fuel
Countries with high sanitation coverage England 2.98 2152 Norway 2.92 2292 Austria 1.24 1897 Australia 1.14 1234 USA 0.92 753 New Zealand 0.73 990 Russian 0.42 800 Federation Canada 0.39 1073 Italy 0.39 1899 Countries with low sanitation coverage Czech Republic 2.93 1752 Jordan 2.30 700 Slovakia 1.47 1764 Hungary 1.39 1697 Turkey 0.59 3588 Senegal 0.35 1044 Bulgaria 0.31 1298 China 0.08 834 Iran 0.05
Cameroon Nigeria Mali Ethiopia
Per truckload to empty latrines
USD per 1000 l diesel fuel
120 45 38.2 16.50
1120 935 1061 742
USD per kW h1 of electricity
0.29 0.07 0.18 0.11 0.04 0.12 0.08 0.26 0.26 0.06 0.14 0.14 0.17 0.17 0.59 0.09 0.03 USD per 1 kW h 1of electricity 0.12 0.21 0.06
permeable material and a perforated pipe – from which used water is distributed – is placed in the centre at about 0.2 m beneath the soil surface. The perforated pipe is typically around 0.1 m in diameter (Hughes et al., 2006). The size of the trench depends on the wastewater load and the soil type, groundwater depth, and precipitation. Leach drains are not recommended disposal options if the groundwater table is close to the surface (e.g.,o 0.5 m depth) or the soil has low permeability (e.g.,o3 mm d1).
4.06.7.1.2 Evapotranspiration beds They are convenient where soil is highly impermeable (e.g., clay) but can also be used in permeable soil from where water is both evaporated and infiltrated. In each case, plants are positioned to increase evapotranspiration and to remove nutrients from wastewater. If a limited area is available, evapotranspiration beds can be used in conjunction with a seepage trench. To increase dispersal of the wastewater throughout the whole bed, perforated pipes surrounded by gravel are used. The design of the bed should ensure it is large enough to hold wastewater loading and pluvial precipitation while, at the same time, providing sufficient water and nutrients to plants (Hughes et al., 2006).
4.06.7.1.3 Soil aquifer treatment and aquifer storage recovery system Soil disposal can be coupled with soil treatment in the soil aquifer treatment–aquifer storage recovery system (SAT-ASR). An aquifer storage recovery system (ASR) consists of holding water in an appropriate underground formation, where it remains available in such a way that it can be recycled by extraction when needed. An ASR can have several objectives, some of which are (Dillon and Jime´nez, 2008; Jime´nez, 2003) temporary or long-term storage; decrease of disinfection by-products; reestablishment of underground water levels;
Safe Sanitation in Low Economic Development Areas
183
60
%
40
20
China
Slovakia
Bulgaria
Turkey
Czech Republic
England
Russian Federation
Canada
Japan
Austria
Norway
USA
0
Figure 12 Estimated percentage of total wastewater treatment costs required for wastewater sludge treatment and management (with information from LeBlanc et al., 2008).
maintenance or improvement of underground water quality; prevention of saline intrusion; deferment of expansion of water supply systems; aggressive water stabilization; hydraulic control of contaminant plumes; and compensation of soil salinity lixiviation. The major advantages of underground storage is that evaporation losses are considerably lower than dams (B1%) and do not have the eco-environmental problems associated with them (Dillon and Jime´nez, 2008). Aquifers can be an economical option to reintegrate water to the environment in arid and semi-arid countries where it remains available for future use. They are also convenient in densely populated urban areas where, besides storing treated water, aquifiers can store stormwater runoff.
4.06.7.2 Disposal into Surface Water Bodies or Reintegration of Used Water to Surface Water Bodies Effluents from treatment plants can be used for the augmentation of surface water bodies, in which the effluent is diluted with freshwater and reused as a source for water. The water quality of receiving water should be preserved to facilitate a safe water supply. For this, it is important to control pollutant content in the effluent, notably pathogens, organic matter, and nutrients (especially for surface water bodies with slow flow). Two aspects need to be monitored: oxygen depletion in rivers and eutrophication in dams and lakes. To avoid oxygen depletion, biodegradable organic matter needs to be removed before introducing the wastewater. There is considerable literature available concerning this aspect as it has been the main target for most wastewater treatment processes. Control of eutrophication is achieved by removing N and/or P from effluents; this is an operation costly to perform in wastewater treatment plants for most developing countries. As an
alternative, land treatment can be used or treated wastewater used first for agricultural irrigation recovering it from the agricultural drainage before sending it to on lakes. Eutrophication of dams and lakes is a frequent problem in developing countries; alternatives for its control are discussed in Box 7.
4.06.7.3 Reuse Reuse is another option to reintegrate water to the environment but through its use. Due to the increase in the human population and the increased use of water for almost all human activities, water is becoming scarce and new tools are needed to use it better. Such tools are (1) the efficient use of water (using less water for the same activity – this is beyond the scope of this chapter) and (2) water reuse. Water reuse is a key component to alleviate the mismatch between water supply and water demand. At the global level, water availability is of around 8500 m3 inhab1 yr1 but with important variations at a regional, national, and local level. For instance, it is estimated that around 700 million people (11% of the total population) in 43 countries live in areas with less than 1000 m3 inhab1 yr1. By the year 2025, 38% of the total world population will live under such water stress, increasing to 50% (in 149 countries) by the year 2050 (UNDP, 2006). As shown in Maps 3, 4, and 9 (Annex 4), most of the affected people live in developing countries. For these countries, three aspects can be highlighted concerning water stress and water demand. First, water is needed for economic development and a better quality of life (even if industrialized countries are not completely making an efficient use of water; they use 30–50 times more water than developing ones (UN/WWAP, 2003)). Second, agriculture is the dominant user of water worldwide, but, in addition, for developing countries, agriculture is usually the
184 Table 11
Safe Sanitation in Low Economic Development Areas Criteria for selecting wastewater treatment operation and processes
Process applicability Must be evaluated based on past experience, data from full-scale plants, published data, and from pilot and full-scale plant studies. If few data or unusual conditions are encountered (atypical wastewater characteristics) pilot plant studies are essential. For developing countries: – Since much less experience is available, a good wastewater characterization is needed as well as a request during bids that the applicability of the processes should be demonstrated before construction. – Bids should encourage operating at lower costs at the same pace the process is optimized. – Technology complexity need to be in agreement with the type of community being served: rural areas, rural isolated areas, small urban towns, large towns, and megacities (low-, middle-, and high-income urban and periurban areas densely or dispersed populated). – Possibility to combine treatment technologies with soft intervention methods (management). Performance Performance needs to be expressed not only in terms of the effluent quality but also on its allowed variability, and both must be consistent with the effluent discharge requirements and the possible use of treated wastewater. Performance needs also to be considered in terms of its reliability, as it may vary according to the process type. Reliability is very important when the effluent is to be reused or treated water is to be discharged into sensitive aquatic environments. For developing countries:
Performance should be verified in terms of the disinfection needs locally required. Influent wastewater variability Consider wastewater characteristic variations in probabilistic terms. Consider wastewater variability in terms of climate change impacts and climate variability. For developing countries: – It is important to have a statistically representative wastewater characterization considering parameters not only defined in norms but also those that might interfere with the treatment processes or the future use of treated water. – Design data should not be based on bibliography data, especially that coming from other countries. – Since segregation and pretreatment of industrial discharge is not common, there are high chances that the wastewater to be treated will contain inhibiting constituents. An evaluation of these is important but not as intensive as the one required for the characterization of the targeted treatment parameters. – Consider wastewater quantity and quality possible variation if programmes to reduce water consumption (such as the use of water less toilets) are to be implemented. Reliability
Achievable performance needs to be expressed in statistical terms and in short and long terms, taking into account water flow and wastewater quality variations. For developing countries: – Unusual situations and emergencies are common. Selecting robust albeit more expensive processes might be cheaper long term, both economically as well as in terms of the negative effects that malfunctioning can produce. Process sizing
Reactor sizing is based on the governing reaction and kinetic coefficients. If kinetic data are not available, process loading criteria are used, but not always with good results, even in developed countries. For developing countries: – Most of the available information used in the design of biological process comes from the developed world, where wastewater and climatic conditions, among others, are different, and so bibliographic kinetic data and load criteria use should be avoided as much as possible. – For coagulation–flocculation process doses and mixing conditions determine at laboratory conditions are essential to minimize cost and sludge production. – For disinfection processes conditions need to be determined or checked up using laboratory data – If experimental data are not available, the adjustment of published data to local conditions, such as pressure and temperature, should always be checked in bids. Applicable flow range and flow variations
The process should be matched to the expected ranges of flow rates. Moreover, whenever possible, considering the presence of stormwater, notably considering impacts of climate change. For developing countries: – For those located in regions with high pluvial precipitation concentrated in short periods of time, treatment processes must be able to deal with flow and major variations in quality. – Alternatively, the use of flow equalization tanks and their cost should be considered. – Processes that can be operated as modules than can be easy to start should be preferred to match variable influents in terms of quantity and quality. Residual treatment and disposal The types and amounts of solid, liquid, and gaseous residuals produced must be estimated. Use pilot plant studies to identify and quantify residuals.
Safe Sanitation in Low Economic Development Areas Table 11
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Continued
For developing countries: – By-products and wastewater treatment residues are often disregarded in proposals in order to offer a lower operating cost. To avoid this, it is important to clearly state in bids that any residues must be quantified and the management options considered within costs. Sludge processing Design, operation, and maintenance must have the same degree of investment and complexity of its management as that of the wastewater treatment. For developing countries: Revalorization of sludge as biosolids (treated sludge) for soil fertilization, erosion control, or land remediation are to be considered as a priority. For urban areas, use of biosolids to cover landfill cells can be an interesting disposal option. Climatic constraints Temperature affects the reaction rate of most chemicals and biological processes; therefore, local water temperature should be taken into account when selecting a processes. For developing countries: – In most developing countries temperature is relatively high, so problems arise due to high temperatures not low ones. High temperature may accelerate odor generation and also limit solubilization of gases such as oxygen. In densely populated urban areas, temperatures may rise even more than expected due to the ‘heat islands’ phenomena. Environmental constraints Environmental factors, such as prevailing winds, may restrict or affect the use of certain processes, especially where odors are produced near residential areas. A wastewater treatment plant may have negative impact on the environment if not properly designed. The disposal site restrictions of the treated wastewater need to be considered regardless of the norms to be met. Water and sludge reuse Water reuse can be a way of making wastewater treatment more attractive in economic terms.
For countries located in water-stressed areas, besides being ecologically sound to reintegrate water to the environment as disposal option, reuse serves to alleviate water scarcity. For developing countries – Land degradation is costing 5–10% of their agricultural production (Young, 1998) and fertilizers have often a prohibitive cost for farmers; in both cases, biosolids can be used to remedy these problems. Ancillary processes Wastewater treatment plants are often accompanied by ancillary (complementary) processes that do not necessarily directly relate to the wastewater treatment process, such as power plants, special storing facilities for reagents, etc. It is important therefore to know, before selecting a process, what are those needs, their cost and viability to obtain them from the local market. Chemical requirements The type and amount of chemicals to be used need to be considered as well as their cost and market availability, both now and in the future. If chemicals are added during the treatment of wastewater or sludge and these are to be reused, their selection needs to be compatible. For developing countries: – Although the use of chemicals is often prohibited, an economic comparison is worth making, especially if chemicals are locally available. Energy requirements The present and future cost of the energy used is something to consider. In selecting and designing wastewater treatment plants, the location, efficient use of energy, and the possibility of recovering/producing energy for in-plant use must form part of the selection criteria that in the long term will contribute to properly closing the urban water cycle. The energy foot print of the wastewater and sludge treatment plant should be minimized to contribute to the reduction of GHG (greenhouse gases). Personnel requirements The amount of people as well as their skill levels need to be well defined. For developing countries – The most common situation is a high availability of low-skilled personnel working for low salaries. Thus, selected processes may have a high labor demand but cannot be very sophisticated. Alternatively, intense training programs should be considered; nevertheless, high indexes of personal rotation are frequently experienced in developing countries when personnel are trained. Complexity and compatibility Define operational needs under routine and emergency conditions. Define the type and need for repairs. It is important that the items selected be compatible for efficient operation. For developing countries: – It should be considered that cheap or obsolete equipment may become costly if frequent repair is needed. – Equipment and spare parts must be available within an appropriate period of time. Obsolete equipment is very difficult to repair. (Continued )
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Table 11
Continued
– Normally, few items are produced or available locally, therefore overall equipment selection needs to consider compatibility between different equipment traders. Adaptability Many treatment plants will need to adapt to future conditions and not all systems have the same capability to be adapted. Economic life-cycle analysis Cost evaluation must consider initial capital cost and long-term operating and maintenance costs. The plant with lowest initial capital investment may not be the most effective with respect to operating and maintenance costs. The nature of the available funding will affect the choice of the process. Land availability It is important to consider the size of the selected treatment process with respect to available land, including buffering zones for future expansions. For developing countries: – There is not always land or cheap land available, as frequently believed. – Considering the fast growth of cities in the developing world and the possibility of building plants in modules, it is very useful to consider buffering zones to increase treatment capacity, complete the treatment process or even to avoid building human settlements near to the facilities. Public acceptance
Communities reject systems producing foul odors or vector breeding. Communities also tend to more readily accept natural process that are integrated with the landscape.
Low-income communities accept better technologies that are a source of jobs for local people than rich ones. Adapted from Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO Leiden: Taylor and Francis Group.
Box 7 Eutrophication Control (with information from Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO; Leiden: Taylor and Francis Group.) Eutrophication is a process in which plants (such as water lilies or hyacinths (Eichornia crassipes), hydrilla (Hydrilla verticillata), cattail –(Thypa sp.), and duckweed (Lemna sp.)) proliferate in surface water bodies due to the presence of high concentrations of phosphorus and/or nitrogen that may come from wastewater, treated effluents, or agricultural runoff. It is commonly observed in polluted lakes or dams, but problems in low flow rivers and agricultural canals have also been observed. Aquatic plants cover the water surface preventing sunlight and oxygen from entering the water. Other negative effects that are provoked are (1) oxygen depletion in the hypolimnion; (2) release of Fe, Mn, NH4, and heavy metals from the sediments; (3) vector breeding, such as Schistosomas and mosquitoes; (4) loss of biodiversity, especially in higher trophic levels; (5) displacement of native species, (6) obstruction of hydroelectric plants and irrigation canals and drains; and (7) restrictions on tourist, recreational, and fishing activities. To reduce aquatic weed density (plants m2), five methods are available: *
*
*
*
*
Biological control. It consists of using living organisms to control weeds. In theory, it is a cheap option as no equipment or chemicals are required but it has an associated labor cost in order to perform maintenance. To be completely effective, the rate of grazing needs to be higher than the plant growth rate, which is very difficult to match in practice. A wide variety of fish, arthropod, fungi, and bacteria have been used for this purpose. Mechanical Control. These methods remove or cut weeds into pieces using mechanically or manually operated equipment. It is an expensive option that can play a role in quickly reducing the extent of infested areas prior to the application of another control method. Chemical control. Pesticides are also used to control weeds. Some substances that have been used are terbutryn, diquat, 2,4-D, glyphosphate, paraquat, and simazine. However, due to their toxicity, they can only be applied under controlled conditions and for a limited period of time. Water level control. In this method, the water level is decreased so the weeds located close to the edges of the water body dry out. The applicability of this method is limited to dams where water levels can be controlled, and to the dry season in which rain would not convey plants once again to the water. Nutrient control. Weed growth is caused by high N or P content in water, and so, lowering their concentration through wastewater treatment is another alternative. Unfortunately, the cost remains high.
Due to their low efficiency or cost implications, in practice, two or more methods are often used to control weeds.
main source of income and the main mean to feed a growing population. Third, the increasing demand for water by municipalities and industries is increasing the competition for its use with farmers. It is estimated that, in developing countries, water withdrawals will increase more (27%) than in developed ones (UNDP, 2006). Among the uses demanding water, sanitation needs to be considered and, in that respect, water reuse may be a component in some areas to promote it through the alleviation of water demand, saving water for
sanitation facilities or through coupling projects to treat wastewater with reclamation ones.
4.06.7.3.1 Types of water reuse Two types of water reuse can be distinguished: nonintentional and intentional or planned. As, in several developing countries, lack of sanitation is generating nonintentional reuse, national policy will need to encourage controlled options
Safe Sanitation in Low Economic Development Areas
instead of promoting practices to start up water reuse. This is the biggest difference with developed countries, where reuse is being promoted once wastewater is treated.
4.06.7.3.2 Unintentional reuse In literature, water reuse is considered merely as an activity where wastewater is intentionally treated to be used once again. Therefore, water reuse is understood as an artificial man-made practice. However, unintentional reuse also exists as part of the natural hydrological cycle, but this is frequently not acknowledged. (Jime´nez, 2009a). ‘Nonintentional’, ‘nonplanned’ or ‘incidental’ water reuse describe situations where used water is mixed with (or becomes part of) the water supply. In most cases, this unplanned reuse is difficult to identify, although it would be important to acknowledge it in order to properly control it. The nonplanned use of water is at the origin of the presence of emerging chemical pollutants in water sources and the reason why drinking water standards are becoming increasingly comprehensive and stringent and more sophisticated technologies to treat water are needed (Jime´nez, 2009b). Nonplanned reuse of wastewater is happening for agricultural irrigation, aquifer recharge, and human consumption. 1. Nonplanned reuse for agriculture. Three-quarters of the total irrigated area worldwide is located in developing countries, and, as a consequence, there is a high dependence on water for food production. Frequently, due to lack of sanitation in these countries, wastewater is used to irrigate land. This is a practice that happens almost naturally because of the combination of the high demand for water for irrigation (81% of total use compared to only 45% in developed countries, Figure 13), the availability of wastewater, the productivity boost that the added nutrients and organic matter provide, and the possibility to sow crops all year round (Jime´nez, 2006). It is estimated that at least 20 million hectares in 50 countries (around 10% of irrigated land) are irrigated with raw or partially treated wastewater (WHO, 2006).
Agriculture
Approximately one-tenth of the world’s population consumes crops irrigated with wastewater, diluted or not. As an example, in Hanoi, Vietnam, wastewater is used in the production of 80% of the vegetables consumed locally (Ensink et al., 2004). The use of nontreated wastewater is also common for urban agriculture, which is practiced in urban and periurban areas of arid or wet countries where there is local demand for fresh food products, and people live on the verge of poverty with no job opportunities (Jime´nez, 2009b). For urban agriculture, wastewater flowing in open channels is used to irrigate very small urban plots of land where trees, fodder, or any other product that can be introduced to the market in small quantities (flowers and vegetables) or be used as part of the family diet are grown (Ensink et al., 2004). In terms of volume, reuse of nontreated wastewater is at least 6 times higher than of treated wastewater (Jime´nez, 2006; Jime´nez and Asano, 2008). As a consequence, any sanitation project in localities using wastewater should consider its actual use. 2. Unintentional reuse for water recharge. Since groundwater is not water that can be observed as in lakes or dams, very often its pollution and nonintentional recharge is not perceived. Infiltration may result from agricultural irrigation, leakages from wastewater and water urban networks, unlined dams, tanks or reservoirs, and on-site sanitation systems. Little information on the extent of this problem is reported in literature, but some cases (a summary is presented in Table 12) have been described highlighting the importance of this phenomenon as a source of water supply. For the one referring to the Tula Valley, it has been the best documented (Jime´nez, 2008b) that recharge with wastewater amounts to at least 25 m3 s1, and the aquifer is used to supply 500 000 people. Infiltration and pollution of groundwater supplies varies from negligible to severe, and the recognition of unplanned reuse is needed in order to advance understanding of how to manage the risks. This may involve continuing groundwater recharge with water of improved quality and/or separating the
Municipal
Industrial
Low income countries Middle income countries High income countries
Developing countries Developed countries
World 0%
20%
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40%
60%
80%
Figure 13 Water use in developing and developed countries (with information from Earth Trends, 2009).
100%
188 Table 12 aquifers
Safe Sanitation in Low Economic Development Areas Examples of unintentional indirect potable reuse via
City
Recharged water
Groundwater uses
Hanoi, Vietnam
Sewer, storm water
Hat Yai, Thailand
Drainage canals, on-site sanitation facilities Primary effluent Mix industrial effluent
Irrigation and drinking Drinking
Ica Valley, Peru Leon, Mexico Merida, Mexico Mexico City (southern part), Mexico Santa Cruz, Bolivia Sana’a, Yemen Tula Valley, Mexico
Sewer, storm water On-site sanitation facilities On-site sanitation facilities Cess pits Untreated effluent
Drinking Irrigation and drinking Drinking Drinking
Drinking Irrigation and drinking
Adapted from Dillon P and Jime´nez B (2008) Water reuse via aquifer recharge: Intentional and unintentional practices. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and needs. London: IWA Publishing.
recharge areas further from points of water abstraction. Appropriate monitoring information will allow the most cost-effective investments to be identified (Dillon and Jime´nez, 2008). 3. Nonintentional reuse for human consumption. Nonintentional reuse for human consumption occurs as described previously, not only through aquifer recharge but also through surface water sources when effluents, treated or nontreated, are discharged into them. This has been documented in developed countries. For instance, in the River Thames in England, during dry periods, 70% of the water used as supply downstream comes from treated effluent. In California’s Santa Ana River, a large part of the supply consists of treated wastewater (Gray and Sedlak, 2003) and in Berlin, 17–35% of the city’s water supply comes from an advanced treated effluent that is discharged to a nearby water supply (Jekel and Gruenheid, 2008). The increasing evidence of the presence of emerging contaminants in water sources is an indication of the nonintentional reuse of water. Information on this subject for developing countries is very poor, and possibly only reported as pollution cases. Recognizing the nonintentional reuse of water for human consumption will help society to acknowledge that water reuse is unavoidable in the future and also to understand that, to properly reintegrate used water to the environment is needed. For this, tools other than wastewater treatment plants will be needed.
4.06.7.3.3 Intentional or planned reuse According to Asano (1998), wastewater reclamation involves the treatment or processing of wastewater to make it reusable; and wastewater reuse or water reuse is the beneficial use of treated water. Planned reuse may be performed for agricultural
irrigation, industrial purposes, environment restoration, and municipal uses. 1. Reintegrating water for irrigation. Most of the world’s poorest people, 800 million to 1 billion rural people, live in arid areas and depend directly on natural resources, including water, for their livelihoods (Dobie, 2001). In such a context, safe wastewater reuse can be a sanitation option that could also be coupled with food security and economic development goals. Under prevailing land and water management practices, a balanced diet represents a depleting water use of 1300 m3 inhab1 yr1, which is 70 times more than the 50 l inhab1 d1 required for basic household water needs (SIWI-IMWI, 2006). For several middle- and low-income countries, agriculture is currently, and will continue to be, a key sector representing 80% of export earnings. Limited and unreliable access to water is a determining factor in agricultural productivity in many regions, a problem rooted in rainfall variability that is likely to increase with climate change (Lenghton et al., 2005). To feed this sector, water reuse can be one option. Planned reuse of water for agricultural irrigation in developing countries is a convenient strategy for many reasons (Jime´nez and Gardun˜o, 2001; Jime´nez, 2006, 2009a; WHO, 2006; Keraita et al., 2008), such as • It is an easy option to increase controlled reuse when nontreated wastewater is already in use as it allows more profitable and safe products. • It can be a low-cost option to manage wastewater and to reintegrate water into the environment. • It allows the reclamation of nutrients (N and P, to increase soil fertility) and organic matter (to improve soil characteristics) at no cost. • Particularly in (but not limited to) arid and semi-arid areas, it permits higher crop yields, as it allows crops to be sown year-round due to higher water availability. • Due to the availability and reliability of water, crops with better profitability can be selected. • It avoids discharging pollutants to surface water bodies (which have a considerably lower treatment capability than soils). • It is possible to recharge certain type of aquifers through infiltration. • It can be part of a strategy to secure food and increase poor people’s income in water-scarce areas. To obtain all the advantages from reusing wastewater for agriculture in planned projects, it is important (1) to control possible negative effects (Jime´nez, 2006; WHO, 2006) such as those related to health; (2) to keep in mind that in many cases nontreated wastewater is being reused at low or even no cost by poor farmers and, hence, they will be unable to afford reuse costs; and (3) from the legal aspect, the historical use of nontreated wastewater by farmers confers riparian rights. 2. Reintegrating water for industrial reuse. Industrial reuse (reclamation of wastewater from a different use, i.e., reuse of a municipal effluent for industrial cooling) differs from municipal and agriculture reuse as it involves the private sector that has its own rules and well-defined needs driven
Safe Sanitation in Low Economic Development Areas
by economic factors (Jime´nez and Asano, 2008). Before reusing water, industries always prefer to implement watersaving projects as these immediately reflect on their budgets; for reusing water, investments to provide proper treatment and monitoring programs are needed. To promote industrial reuse, the best government strategy is to provide incentives rather than setting compulsory regulations (Jime´nez and Asano, 2008). Among the different industrial reuse options, cooling is the most popular due to its high water demand, and the possibility of using secondary-treated municipal effluents, sometimes coupled with filtration or softening processes. As a consequence, power plants located near urban areas are potential sites of industrial water reuse. 3. Reintegrating water to the environment. More than 1.4 billion people live in river basins where the intense use of water threatens freshwater ecosystems (Smakhtin et al., 2004). Reintegrating water to the environment is a practice that is gaining momentum, as it is being recognized that (1) the environment needs water and (2) the environment has the same entitlement to water as other uses. Unfortunately, these two aspects are better recognized by developed countries than developing ones. Overuse of water tends to occur in regions heavily dependent on irrigated agriculture or where there is rapid growth of densely populated areas (UNDP, 2006), two characteristics common in developing countries. Among the more prominent examples (UNDP, 2006) of water overuse, the exploitation of the Yellow River basin, in northern China, can be cited: Human withdrawal currently leaves less than 10% of the flow remaining in the river. The river ran dry 600 km inland for a record 226 days in 1997. The drying up of the river caused a drop in agricultural production averaging 2.7–8.5 million tons a year, with losses estimated at US$1.7 billion for 1997. The purified effluent from sewage treatment plants can be used for the augmentation of river flows, to raise the level of wetlands or lakes, to recover dried lakes, or even to create new lakes or wetlands. In doing so, biodiversity may recover. Care must be taken when restoring water into water bodies to preserve or improve the actual quality of water. Used water reclamation can be combined with rainwater reclamation. Water reuse with environmental restoration can be coupled with projects of urban image improvement or programs to provide better facilities at recreational areas. 4. Restoring water to aquifers. Aquifer recharge is not, itself, a use of reclaimed water but is often part of the pathway to reuse. It is a convenient way to reintegrate water into the environment but can be used only under certain circumstances related, in particular, to the type of soil and groundwater. Aquifer recharge can be performed to recover groundwater levels, to control saline intrusion, to augment drinking water sources, to protect and, in some cases, to improve underground water quality, to protect surface water bodies from contamination by effluents, to increase water availability for any use, and simply to store water for the future (Dillon and Jime´nez, 2008; Corrleje et al., 2008). Intentional recharge with reclaimed water can play a role in providing balanced storage and supplemental treatment for water (Bouwer, 2002; Dillon and Toze, 2005). It also provides low-cost storage that occupies a minimum of
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valuable urban land, while stored water is protected from pollution and evaporation. There are two methods to recharge aquifers. The first is known as land-spread infiltration where treated wastewater infiltrates through soil by gravity. This option has relatively low operating and maintenance costs. The second method for recharge is direct well injection. In this option, wells are used to convey a highly treated effluent directly to aquifers. Regulation to recharge aquifers are very different from one country to another; some are set at a national level while others are defined using a case-by-case approach (Jime´nez, 2003). Most of the projects to recharge aquifers are found in developed countries. In developing ones, some examples are found in Atlantis, South Africa (for drinking and agricultural purposes, using pond infiltration), in Windhoek, Namibia (for drinking purposes and using injection wells), in New Delhi, India (for irrigation using infiltration ponds for treated urban wastewater and stormwater), in Beijing, China (for drinking purposes using wells and recharge basins), and in Mexico City, Mexico (for drinking purposes on a limited scale and using infiltration ponds; Dillon and Jime´nez, 2008). In all these cases, wastewater is treated to at least at a secondary level (see section titled ‘Relevant websites’). 5. Reintegrating water for municipal use. In 20 years, 60% of the world’s population will be living in cities (UN, 2006). This being the case, more water will be needed for municipal use and, at the same time, more municipal wastewater will be produced. This situation, therefore, represents an opportunity to increase municipal wastewater reuse. Water reuse in cities represents an opportunity to conveniently treat wastewater, with environmental and even economic advantages. Opportunities to reuse wastewater in cities are classified into two groups: (1) those demanding relatively low-quality water and involving low health risks, and (2) those demanding high-quality water where health risks are high. In the first group, there are several types of uses, such as: (a) the filling of recreational lakes or the operation of fountains; (b) car, truck, or street washing; and (c) green area irrigation. Options demanding high water quality include reuse for drinking supply. Around the world, there are successful examples of both types of reuse, low risk options being the most common. Water reuse for human consumption, although less common, is no less important. Moreover, the only two examples of the reuse of water for human consumption in the world are notably from two countries from the developing world: Namibia and Singapore (Box 8).
4.06.7.3.4 Graywater reuse Graywater (i.e., domestic wastewater not containing toilet wastewater) is more accessible for reuse as it is less contaminated than wastewater, notably in terms of (but not limited to) pathogens. Typical sources of graywater are bathing, laundry, dishwashing, and food preparation. Due to its comparably low and easily degradable contamination, it can be relatively easily treated for reuse. Graywater reclamation entails the production of less wastewater to be treated in
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Safe Sanitation in Low Economic Development Areas
Box 8 Reuse of wastewater for human consumption in Namibia and Singapore Windhoek, Namibia, has been reusing wastewater for human consumption for more than 40 years (Van der Merwe et al., 2008) as result of an original idea in 1956. Since its operation, no measurable health risk has been observed and neither have people drinking reused water displayed associated health problems. The reclamation plant has undergone several modifications to improve the technology used. The quality of the water supplied can be consulted every day in the local newspaper. The amount of water reused is around 250 ls1, which is distributed after dilution by a factor of 1–3 with first-use water. The monitoring program for the facility represents 20% of the operating costs, and is performed by the wastewater treatment plant and also by three independent laboratories. The system is operated using a multiple barrier concept that goes beyond the wastewater treatment plant. The astute words ‘‘Water should be judged by its quality; not its history’’ are attributed to Dr. Lucas van Vuuren (van der Merwe et al., 2008), one of the pioneers of the Windhoek reclamation system. This refers to the fact that fear of reused water should be based on rational aspects. The other example of direct reuse of wastewater for human consumption comes from Singapore (Funamizu et al., 2008) and is known as the NEWater project. It started in 2003 and uses a secondary effluent that is further treated with a membrane system (microfiltration (MF) and reverse osmosis (RO)) and UV-light disinfection. The water produced is cleaner than tap water as it fulfills all the requirements set by US-EPA and WHO for drinking purposes. Treated water is channeled to a reservoir, from which it is taken as supply after dilution with first-use water. Water is distributed through the network for use for domestic and industrial purposes. When the NEWater project was launched, it operated at a rate of 870 l s1. This will be progressively increased to reach 2400 l s1 by 2011 (B0.5% and 2.5% of total water consumption, respectively). In both cases, Namibia and Singapore, before the implementation of the reuse programs, stringent industrial pre-treatment programs and segregation of industrial effluent from the sewer were put in place.
centralized plants. Graywater reuse is performed at the same facilities where it is produced and, as a result, a short storage time is needed (1 day retention time). Graywater reuse can be performed individually (for a single home) or collectively (several groups of houses or larger buildings). Treated graywater may be used for watering plants, kitchen gardens, and for the safe augmentation of ground- or surface water. Treatment can be very simple or highly sophisticated, ranging from simple manually operated sand filters to biomembrane reactors, hence, covering the needs for rural areas or buildings located in upmarket areas in megacities. Further details on design and operation can be found in Correlje and Schuetze (2008). Graywater reuse can be as well an important component for basic sanitation, as described in Section 4.06.6.1.
4.06.8 Sludge and Excreta Management As the quantum of wastewater treatment is still low in developing countries, little information is available concerning the actual situation. LeBlanc et al. (2008) performed a survey in some countries showing that the tendencies are the following: 1. For middle-income countries. From information coming from 10 middle-income countries, including Africa (Namibia and South Africa), the Middle East (Iran, Jordan and Turkey), Asia (China and Russian Federation), and Latin America (Brazil, Colombia and Mexico), it is shown that wastewater treatment facilities serve mostly urban areas using preliminary, primary, and, in some cases, secondary processes. For rural or poor periurban areas, basic sanitation facilities are provided. Although sludge is produced in these facilities, this is not always managed as part of the sanitation service. The disposal options for the sludge from wastewater treatment plants produced are landfill dumping, dumping into sewers, storage at wastewater treatment plants, land application, and agricultural reclamation. Land application and agricultural reclamation are options
limited by space problems, while the use of landfills is restricted in densely populated urban areas, where solid wastes compete for space with sludge. As sludge production is still low in the few wastewater treatment plants available, sludge management policies are novel, and are still in a maturation phase. Some of these policies offer new approaches different to those used in developed countries (LeBlanc et al., 2008). With regard to fecal sludge, the main constraint for their management is the cost to empty on-site sanitation systems as these are often located in inaccessible areas, are large in number, and are frequently highly dispersed. It is noted that the high cost of latrine emptying is not sustainable, even for large municipalities. Extracted fecal sludge is often buried on-site, dumped into landfills or sewers or sent to uncontrolled discharge sites. Discharge of sludge and fecal sludge in sewers often lead to surpass the wastewater treatment plants’ capacity when available. 2. For low-income countries. Data from different African countries (Burkina Faso, Cameroon, Coˆte D’Ivoire, Ethiopia, Mali, Mozambique, Namibia, Nigeria, Senegal, and South Africa) demonstrated a similar situation focused on the need to provide basic sanitation services either in rural or urban areas. Few cities have complete sewerage systems and, when available, sewers frequently feed into partially functioning wastewater treatment plants. In these countries, the use of on-site sanitation systems, such as septic tanks, bucket latrines, pit latrines, and dry latrines, produces fecal sludge, which is often ‘contaminated’ with domestic waste. In dense informal settlements, the challenges to properly handle fecal sludge are significant as besides the technical constraints other factors related to the social, political, and cultural aspects come into play. Fecal sludge handling includes the need to provide reliable and low-cost options to emptying the facilities, to provide proper and affordable treatment and transportation, and to have suitable sites for safe disposal. Literature exists concerning the alleviation of sludge and fecal sludge disposal and revalorization problems, not all of which is relevant for developing countries. Common issues in
Safe Sanitation in Low Economic Development Areas
properly managing sludge and excreta in developing countries are as follows (LeBlanc et al., 2008; Jime´nez, 2006, 2008):
• •
• •
Conventional sludge and excreta treatment options used in industrialized countries do not necessarily achieve the levels of pathogen inactivation required for its safe reuse. Nutrients, organic matter, and energy are resources available in fecal and wastewater sludge that should be utilized as best as possible. There are examples around the world showing the feasibility and convenience of reclaiming them. Applying properly treated excreta and biosolids to soils in a safe way can contribute to soil fertility and with it to food security; it can also raise income for poor farmers. Proper management of excreta and wastewater sludge can significantly reduce releases to the atmosphere of potent greenhouse gases such as methane and contribute to carbon sequestration in soils.
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4.06.9.1 Integrated Water Resources Management In order to consistently provide sustainable water services, it is recommended that an integrated water resources management (IWRM) approach is used. This approach is useful to analyze situations such as when
• • • •
4.06.9 Policy
•
The MDG Target 10 stating ‘‘Reduce by half the proportion of people without sustainable access to safe drinking water and basic sanitation is considered under Goal 7: Ensuring environmental sustainability’’ (Box 9). Therefore, sanitation is to be provided in a sustainable framework which, in practice, means to provide a service comprising much more than was expected in the past. To implement it, a proper policy is needed.
• • •
multiple barrier system comprising solutions that go beyond the construction of wastewater treatment plants need to be implemented to protect health and the environment; sanitation needs to be provided as a tool (sometimes indispensable) to have clean water supplies and to provide a safe water supply (Box 10); sanitation is coupled with projects contributing to food security, job opportunities, increases in exportation, soil erosion control, efficient use of water, etc.; sanitation needs to be provided over a wide area rather than to a single section of it to effectively control negative environmental impacts; sanitation needs to be part of a three R concept system (reduce, reuse, and recycle); sanitation is considered as part of a cycle in which wastewater is properly reintegrated to the environment; sanitation needs to consider the impacts caused by climate change; projects need to be designed, operated, and/or managed by different institutions, sectors, basin agencies, or even countries;
Box 9 What does sustainability mean? ‘‘A process that promotes the coordinated development and management of water, land and related resources, in order to maximize the resultant economic and social welfare in an equitable manner without compromising the sustainability of vital ecosystems’’, UN-Water, 2008 According to LeBlanc et al., 2008, elements defining sustainability are * * * * * * * * * * *
dealing transparently and systemically with risk, uncertainty, and irreversibility; ensuring appropriate valuation, appreciation, and restoration of nature; integrating environmental, social, human, and economic goals in policies and activities; providing equal opportunities and community participation; conservation of biodiversity and ecological integrity; ensuring inter-generational equity; recognizing the global integration of localities; a commitment to best practice; avoiding net losses of human or natural capital; implementing principles for continuous improvement; and providing good governance.
Box 10 The Bissau case, with information from Correlje AF and Schuetze T (2008). Every Drop Counts: Environmentally Sound Technologies for Urban and Domestic Water Use Efficiency. Division of Technology, Industry and Economics, TU Delft. India: United Nations Environment Programme. Bissau, Guinea, in West Africa is a city attracting huge numbers of people from the surrounding countryside. Most of them have settled in squatter new areas around the old colonial center. During a study performed in the 1990s, it was found that the newly piped water taps ran dry several times per day. As a result, many people returned to the old wells. These were often more contaminated than before because the new pit latrines installed close to the wells polluted the groundwater. Groundwater quality was also impacted by solid waste thrown into the pits dug for the production of adobe blocks to build new houses. Moreover, the new network of gutters was now efficiently removing most of the clean rainwater that used to recharge the groundwater. The gutters caused an extra problem. On the edge of the settlements, where the gutters ended, storm water peaks caused serious soil erosion. This created problems for a newly developed scheme of vegetable gardens on the urban fringe, and even threatened houses.The original problem – the lack of water in piped water taps – was related to electrical power failures causing water pumps to stop. Similar situations can be encountered in many developing countries and they cannot be easily solved as long as their roots are not properly and integrally tackled.
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• •
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good technical solutions needing proper social, economic, and political policies are to be put in place; and wastewater, treated or not, is being nonintentionally reused.
• • •
4.06.9.2 Need for an Own Policy for Developing Countries Developed countries, through experience, research, and technological innovations have progressively improved their sanitation services and have developed systems that are what they need. However, as described in this chapter, the problems they have faced and the problems they are now facing, although similar, are not the same as those confronted by developing countries. Thus, there is a need for low-income nations to develop their own processes using part of the developed countries’ experience. To contribute to this process, a definition of the issues to address and the challenges to face is provided in the following.
4.06.9.2.1 Issues to address The issues that need to be addressed are as follows:
• •
•
Low sanitation coverage lagging behind population growth, needing an intense effort in order to be tackled. Need/importance to couple sanitation programs with others addressing problems such as food security, low income, and soil erosion control. In practice, this requires increased efforts of coordination. Lack of sanitation as a component of poverty, and therefore, as a problem that cannot be completely solved if its roots are not properly addressed (Box 11).
•
•
Lack of sanitation, particularly in vulnerable groups that, due to their own characteristics, are often more difficult to provide services for. A growing population, notably in urban areas and, within them, in slums. Higher vulnerability to the negative impacts of economic and climatic change on sanitation needs. For low-income countries, lack of economic capacity to deal with the cost of covering the sanitation MDG targets and, for middle-income countries, the need to mobilize funding required to put sanitation above other needs. The proper management of sludge and excreta, two byproducts often not considered as part of sanitation targets of funding programs.
4.06.9.2.2 Challenges to face The challenges to be encountered are listed below: 1. The lack of political will and commitment at the highest level (WHO/UNICEF, 2000) is a barrier that is greater than, for instance, the lack of economic resources, the capacity for building, or the acquisition of appropriate technology, since all these may be overcome by a strong political support. In order to develop political will, politicians and society need to appreciate the value of sanitation. An understanding that it is through the provision of water supply and sanitation that industrialized countries build up strong societies with good health and good economic conditions is needed (Box 12). 2. The second challenge is to put in place accountability mechanisms to ensure that resources provided to fulfill
Box 11 The sanitation problem in Cameroon (with information from Mfoulu N (2008) Cameroon. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 169–179. Vienna: UN). In Cameroon, some houses are equipped with a 2 m-deep hole for a latrine, surrounded by pieces of timber. When the hole is full, it is covered with earth and medicinal or aromatic plants, and another facility is built. If the family has no land to dig another hole (as frequently happens), they call the tanker to empty it at a cost of US$120. Sometimes, while the family saves up the money, excreta overflows and pollutes the nearby area where wells and boreholes are located, threatening drinking water quality. When feces are removed by tanker trucks, they are often dumped into rivers or the forest, because there are no treatment facilities. Houses in modern residential areas have septic tanks, and their effluents are directed into wells for filtration. Often, this does not happen in the correct way because builders have not mastered the technology. Some collective residential areas, universities, and hospitals are connected to sewers that convey wastewater to a treatment plant, from where treated water is directed to a river. But still, there are people without access to any of the facilities described above who go into the bush to relieve themselves on the spot. Villagers continue to use this practice because they have no choice.
Box 12 Clean meansy yy healthy? Mexico City produces 21% of Mexico’s gross domestic product (GPD) (US$12 500 per capita). After the swine flu (H1N1 flu) outbreak in May 2009, a loss of US$144 million was experienced solely due to the shutting down of restaurants, and US$35.2 million were lost due to the closure of public transport for just 10 days. To allow the city to return to normal conditions, health experts advised constant handwashing and the disinfection of school toilets. At this point, politicians realized that 200 public schools had no water at all, 195 had malfunctioning toilets and 90 more had no facilities at all. Before the swine flu epidemic, politicians had not understood the link between water, sanitation, and health and had not addressed this problem, although on many occasions parents’ associations had requested the services. The president of one parents’ association commented on the news that, in contrast to most Mexicans, he believed that the swine flu had been a blessing as it was the only way to ensure proper sanitation facilities at schools. The Mexico City government invested US$56 million on the school program ‘Clean means healthy’.
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Box 13 Need for new type of institutions, with information from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan. Water and sanitation service agencies are typically modeled after utilities in industrial countries, and as such are organized around the goals of maximizing operational efficiency for public sanitation components (trunk sewers and treatment plants) rather than providing services to poor people, slums, disadvantaged groups, etc. As result, in, in developing countries, experience and institutional structures to provide the type of services needed is deficient. As a result, services are being provided by other means. Data from India indicate that as much as 8% of rural households across the country invest their own money and use small private providers to construct latrines. Self-provision accounts for about 1 million privately installed septic tanks in Manila and in Jakarta. Research in Africa confirms that the role of the small-scale private sector in sanitation provision is significant. These findings are further supported by data from the JMP (WHO-UNICEF Joint Monitoring Programme): between 1990 and 2000, the increase in the number of people served by sanitation reported by the JMP was much larger than the expected impacts of the public investment that occurred during this period. The reorientation of public programs to either modify their structures or to promote and assist the provision of sanitation services by small private and even familiar companies is needed. This does not currently occur in developed countries.
the MDGs (public and private from donors) will be used wisely and for what they were originally intended for. 3. The third challenge involves a broader aspect. Even if sanitation programs are put in place, if poverty is not properly addressed, most of the solutions provided will be unsustainable. This will possibly lead in the future to adding addressing poverty to the already lengthy list of reasons why sanitation has failed in developing countries (this list already comprises financing, institutions, education, the need for decentralization, and the need for private participation).
• • • • •
4.06.9.2.3 Strategies that can be used Although there is no recipe for success, strategies that can be considered when developing plans for sanitation include the following (Jime´nez and Gardun˜o, 2001; Jime´nez, 2003, 2006; Lenghton et al., 2005; UNDP, 2006; WHO, 2006; LeBlanc et al., 2008; Correlje and Schuetze, 2008): To develop policies:
•
•
•
•
Take time to perform proper planning in order to identify the resources (human and economic) needed to design, build, operate, and maintain facilities, and to develop policies and institutions. Do not initiate projects for which this has not been previously defined, otherwise there is a risk of losing any investments made (a case in point is the existence of many facilities installed around the world, which have been subsequently abandoned). Take time to define how much money is needed, supported by experts with no commercial interest, specifically not those from companies that are potential participants in bids. Define needs and priorities using the best available information even if it does not come from the water sector. Priorities can be set by using the methodology proposed by Lenghton et al. (2005), which considers actual water service coverage, and mortality due to gastrointestinal diseases and density of settlements, considering urban and rural areas. Evaluate risks using quantitative methodologies to properly identify and prioritize problems, and select solutions accordingly (in terms of size, and economic and human resource investments). As much as possible during the planning stage, involve sectors related to the solutions other than the water sector
• • •
(e.g., the federal, regional, and local governments, ministers of the environment, urbanism, agriculture, land use, transport, economic development, social development, finance, etc.). Couple sanitation programs with programs related to food security, soil remediation, and economic development. Produce efficient, affordable, and enforceable norms and set goals for them that are easy to understand. Promote innovation at all levels (institutional –Box 13–, financial, regulatory, and technological). Combine different intervention methods to control problems; consider not only of sewers, latrines, and wastewater treatment plants. Consider water reuse and the safe reintegration of sludge and fecal excreta as an important part of the overall sanitation program. Promote the management of the environment in an integrated way, even considering climate change effects. Design monitoring programs that wisely use resources by including information that WILL be used. Use the new information obtained to evaluate and improve the program. Review the program to ensure it covers the specific targeted population sectors (women, the poor, rural areas, etc.) and meet the defined goals.
For funding:
• • •
•
be creative in finding solutions to funding needs; extend financial support to the poorest households to ensure that sanitation is an affordable option; discern whether there is an absolute lack of resources for expanding water supply and sanitation coverage, or if there is a need to redistribute potentially sufficient existing resources; and develop and put into practice transparent mechanisms to easily and rapidly transfer monetary resources from central to local institutions.
For institutional design:
• • •
Develop national and local political institutions that reflect the importance of sanitation in terms of social and economic progress. Promote institutions throughout government that use or at least understand concepts of integrated management, not only for water. Develop institutions where innovation and solidarity are considered as a virtue.
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•
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Consider the need to have as part of the institutions welltrained and highly professional personnel.
progressively controlling drawbacks; this can be done by promoting controlled reuse rather than adopting vanishing current practices. Incorporate reuse as part of the sanitation standards.
For norms and regulations:
•
•
To set up programs:
•
•
• • • • •
Identify which problems should be addressed by using norms (compulsory), criteria (recommendations), or other type of tools (such as incentives and education). Set appropriate and affordable sanitation risk-based standards, designed to contribute to solving local problems that can be reviewed over time to integrate experience. These should be able to be adapted to new and better conditions in order to move progressively to an ideal situation. Allow the development of norms that are adapted to local needs and capabilities (Table 13). Sanitation systems are often adopted from other developed countries without sufficient adaptation and users tend to put in place an idealized solution in which a uniformly high level of service is provided and the technology to be used is already set. Set up regulations that combine different intervention methods to control risks that are not based only on wastewater treatment plants. Keep in mind that parameters selected are to be enforced and they will demand economic and human resources for. Review the whole legal framework related to the standard so they can fit in and be implementable. Set up standards using a participatory approach, which includes stakeholders and expert participation, notably coming from local universities. Where noncontrolled reuse is already in place, regulations need to maintain the benefits already obtained while
Table 13
•
•
• • • • • •
Perform a national inventory of the actual needs and solutions to be implemented to manage wastewater, excreta, and sludge, include a survey on water reuse possibilities to couple them with sanitation solutions when feasible. Implement policies by promoting incentives rather than imposing rules and fines; but when rules are to be observed, be firm on decisions, and inform society in order for it to be perceived that jeopardizing the health of others is important. As there is no universal solution, support a wide range of sanitation technologies and service levels that are technically, socially, environmentally, and financially appropriate. Promote innovation to have both technically and economically feasible technologies to deal with local pollutants, notably for the high and varied pathogen content. Implement pilot plant programs to test policies and use the information obtained to retrofit your program before scaling it up (Box 14; Spaliviero and Carimo, 2008). Empower local authorities and communities with the authority, resources, and professional capacity required. In order to fund the maintenance and expansion of services, local governments and utilities should ensure that users who can pay, do so. Carry out training programs addressing all stakeholders needs, from plumbers to politicians.
Some aspects to consider when setting regulations
Aspect
Advantages
Disadvantages
Definition of fixed treatment option(s) to use and inclusion of predefined treatment design and operating criteria.
– Reduces the need for monitoring and surveillance. – Renders project implementation easier.
Selection and use of the best indicators as parameters.
Reduces monitoring and surveillance cost.
Selection of normal monitoring parameters and establishment of limits for each one. Use of epidemiological local data.
– Facilitates surveillance.
– Limits innovation – Encourages bias in regulators who will be responsible for both selecting the method of control and meeting objectives. – May lead to nonviable schemes from an economic point of view. – Introduces the idea that indicators are the best and ideal parameters to define pollution. – Most of the current best indicators have been proven effective for developed countries but have not been tested for all conditions in developing countries. – May give a false impression of safety. – Cannot be universal or static over time. – Increases supervision costs.
– Introduce protection for local problems.
Use of toxicological tests.
– Data available internationally. – Helps to establish cause–effect relationship.
Use of risk evaluation models.
– Help governments to make rational decisions.
– Information not always available for all of the diseases currently present. – Often render norms too stringent. – For diseases originating from microbial pollution do not correspond to local conditions when diseases are endemic. – Difficult to explain their meaning to the population.
Adapted from Jime´nez B (2003) Health risks in aquifer recharge with recycle water. In: Aertgeerts R and Angelakis A (eds.). State of the Art Report Health Risk in Aquifer Recharge Using Reclaimed Water, pp. 54–172. Rome: WHO Regional Office for Europe.
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Box 14 Development of a stepwise program in Mozambique (with information from Spaliviero M and Carimo D (2008) Mozambique. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 431–437. Vienna: UN.) Following Mozambique’s independence in 1975, the government identified sanitation as one of the key components to improve health conditions. As such, in 1976, the Ministry of Health launched an intensive national campaign for the self-help construction of latrines. Many thousands of latrines were constructed during a relatively short period. However, there were numerous problems, including insufficient awareness about environmental conditions, a lack of technical guidance in latrine design and construction, and shortages of critical building materials. Consequently, many of the latrines became structurally unsafe and unusable. In response, a research project was initiated in 1979 to ‘‘identify and develop a suitable technology and method for large-scale implementation of improved sanitation in periurban areas.’’ The result was the development and successful pilot testing of an appropriate and cost-effective technology. From 1979 to 1994, around 135 000 improved latrines were produced. In addition, an awareness campaign was carried out on the use of the latrine, hygiene promotion, and capacity building. In 1996, the program was extended to the rural areas. Prior to 1998, more than 230 000 latrines were constructed and installed. In December 1998, the program was formally transferred to the National Directorate of Water Affairs. Overall, it has been a long and steady scaling-up process over more than 10 years that ended by ensuring a progressive withdrawal of the government from latrine production. The emphasis now is given to decentralization and privatization for the services, although the responsibility for the program remains with the government. From this experience, some lessons learned, are *
*
*
•
•
•
• •
Although technology must be simple, it is important for massive use to ensure its local production and commercialization. There must be several types of sanitation facilities with different prices in order to commercialize. A good network needs to be established between users (periurban communities, the government, nongovernmental organizations (NGOs), small private companies, and donors) to ensure that the program progressively developed its own dynamism. Latrines need to be emptied and the service needs to be provided.
Implement programs to segregate and/or pre-treat industrial discharges to sewers to render municipal wastewater treatment more affordable and to avoid the presence of noxious compounds in treated wastewater and sludge that will limit their revalorization options. As wastewater, sludge, and excreta management regulation compliance often depend on the work of different ministries, coordinate the work of such institutions taking care that the objectives of each are compatible. Develop public indicators to follow up progress globally and also consider the implementation of indicators to follow specific targets such as wastewater treatment coverage, safe reintegration of treated water to the environment, and sludge and fecal excreta management. Attention should also be provided to deprived sectors (women, poor people, slums, dispersed rural areas, etc.) Seek to validate your indicators by a third independent party such as a university or a non-governmental organization (NGO). Verify that the same information is provided international, nationally, and locally.
To raise support for the program:
•
•
Make it understandable to all that lack of sanitation means a barrier for economic development is an unsustainable way to manage the environment, is at the origin of local pollution problems, contributes to water scarcity as it reduces water availability, and increases vulnerability and reduces the capacity to adapt to climate change. All of these issues have broad support among society and different groups, not all of which are concerned by sanitation for the poor. Build community-level initiatives through government interventions aimed at scaling up best practice.
•
Create awareness of the nonplanned reuse of wastewater and the importance of investing in it as an option to make clean water accessible for any use.
4.06.10 Funding Figure 14 shows the investments made for water supply and sanitation from 1990 to 2000; it can be observed that, in the past, most efforts were orientated to water supply and cities, leaving sanitation (only about one-fourth of investments made for water supply) and rural areas far behind. Figure 15 shows the origin of investments. In the case of Asia and LatinAmerica, almost all the finances have come from governments, while, for Africa, it represented nearly a half. From the previous analysis, it is evident that there is need to invest money to catch up with the level of services needed. Before calling for funding, it is convenient to analyze (preferably only within each country, without the input of donors or enterprises) what the money should be used for. To sustainably increase sanitation coverage, economic resources are needed not only to build sanitation infrastructure, but also for planning according to local needs and possibilities, developing research and technology, and developing institutional capacity in a local context. Unfortunately, most of the time, funding is provided only for some of these activities (mostly for infrastructure); one major reason being that, often, this is the only type of funding that is sought.
4.06.10.1 Funding Options There are two funding options: public or private, each of which has different modalities. For public funding, the money comes from federal or local governments either directly from tax revenues or user charges, or, indirectly through crosssubsidies from users who can afford to pay, private-sector investment, or international and national loans. Private sector
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Latin America and The Caribbean
Asia
Africa
8
Billions USD
7 6 5 4 3 2 1 0 Urban water supply
Rural water supply
Urban sanitation
Rural sanitation
Figure 14 Investments made in billions of USD between 1990 and 2000 per region for rural and urban water supply and sanitation. Data from WHO/ UNICEF (2000) Global Water Supply and Sanitation Assessment Report, Joint Monitoring Programme for Water Supply and Sanitation. Geneva: WHO.
7 National investment
External support
6
Billion USD
5 4 3 2 1 0 Africa
Asia
Latin American and The Caribbean
Figure 15 Origin of the investments made in billion USD between 1990 and 2000 for sanitation per region. Data from WHO/UNICEF (2000) Global Water Supply and Sanitation Assessment Report, Joint Monitoring Programme for Water Supply and Sanitation. Geneva: WHO.
investments and national and international loans are to be paid from taxes, the difference is only that payments differ in time and are used simply because it is very difficult to finance sanitation projects directly from users. As a result, people who pay for the services are not always the same who will be using them. Private aid is made available by private enterprises or NGOs. Private funding is used simply because developing countries have greater needs than economic resources. The participation of private enterprise cannot be taken for granted as there are several factors that actually inhibit their participation. These include low accessibility to loans from towns and municipalities, the need to organize projects that have payback periods of 20 years, and the need to recover costs through water tariffs (Lenghton et al., 2005). Private funding includes not only international or national firms, but also self-
provision schemes provided by nonconventional private enterprise. These nonconventional private enterprises have been called by some ‘informal’ although for several developing countries, they have in many cases proven to be more formal, useful, and to provide more reliable services than formal ones. For example, in India, an NGO named Sulabh has installed 5500 pour-flush toilets that are operated on a fee-paying basis and are maintained by attendants who live at the facilities. Through providing good reliable service, Sulabh’s facilities have become a model for sustainable public sanitation services. This shows that there is growing knowledge and capacity provided by small and even family-run companies that are capable of producing significant and innovative improvements in access to sanitation. Financing strategies are specific for each country and situation and depend on the political will, the compatibility with
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existing institutional arrangements, the degree of community involvement in decision making, the available economic and financial resources, and the prevailing social and cultural preferences, among other aspects. When either private or public funding is used, some key elements to make a good use of it according to Lenghton et al. (2005) are
•
• • •
Maximum scalability. The selected financing strategy needs to be one that can be scaled up quickly and in a straightforward manner to allow for rapid increases in the population served. Minimal transaction costs. Full financial accountability. Closed revenue cycle, that is, financially viable in the sense that all capital and operating costs are fully covered – either through user fees, government subsidies, or external finance.
4.06.10.2 Why Sanitation Needs to be a Public Process Sanitation is of public interest (Box 15) and hence is a public process. In order to implement what needs to be provided is, for the governments, to identify the main requirements, the areas of responsibility, the risks associated, who is responsible for what, the different options to address needs, and the associated costs. Once this is performed, it is required to review, set up or adapt the legal and institutional framework, and to educate all the persons involved (from society to politicians, experts, regulators, private companies and functionaries,
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besides children and women). Sanitation management (basic sanitation facilities management, wastewater collection, treatment and reintegration, by-product management, and risk control) requires the coordination of different public institutions, society, academia, private enterprises, and in some cases, even different countries. Therefore, the government is needed to set up the programs.
4.06.10.3 Why Private Participation can be Involved Today, around the world, it is still mostly government agencies that construct and operate wastewater collection and treatment systems. However, private companies are contracted to conduct operations in many places, and all countries have significant commercial enterprises built around collecting excreta and septage and managing wastewater sludge and biosolids, mostly in cities. Theoretically, private companies, if well used by the government, could be useful to increase sanitation coverage if the level of society is raised and private companies are not used to increase the already-considerable differences existing between economic social classes. Nevertheless, private participation is not increasing in sanitation. After steadily increasing at a global level between 1990 and 1997, it began to decrease (Lenghton et al., 2005). There are many reasons for this, one of which is that it is not easy to build up successful schemes combining private and public interests.
Box 15 How industrialized governments approached funding for sanitation (with information from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan.) In general, in developed countries, public water infrastructure components have been highly subsidized by governments, reflecting an understanding that the public health benefits of sanitation generate substantial positive external gains that merit public investment. In Britain, for example, urban authorities borrowed more than d7.7 million for sewerage work during the period 1880–91. Eventually, the public provision of sanitation became an uncontroversial and indeed, an expected part of life. Similarly, for many municipalities in the United States, public financing of sanitation infrastructure was seen as the only option for ensuring investment adequate to protect public health. In the nineteenth century, Boston, for example, had lower-than-expected connection rates among households to the city’s new water and sewer network; this prompted the city to cover the cost of service pipes for all unconnected households. In 1850, an influential state sanitary survey concluded that governments must accept responsibility for financing public sanitation infrastructure because, left to their own devices, a large proportion of Massachusetts residents would be unable or unwilling to take on personal responsibility to conduct their lives in accord with recommended sanitary principles. Until recently, grants of up to 70% or more were provided for innovative sanitation technologies in the United States.
Table 14
Type of service and technology more suitable for private and public participation
Type
Modality
Type of service
Technology needed
Private
Public or private sector provision
Sewerage plus wastewater treatment plants
Self-provision
Septic tank systems
Low or normal volume flush water closets; house connections; sewers, biological or physicochemical treatment centralized or decentralized operated. Septic tanks; soakaway pits or absorption trenches; water closets or pour-flush toilets Squat slabs over pits or connected to offset pits
Public
Provision by public, private businesses or NGOs
Pour-flush toilets VIP latrines Nonventilated pit latrines Public latrines
Public water closets; public VIP latrines; public pour-flush toilets; public nonventilated latrines
Adapted from Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan.
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One aspect to keep in mind concerning public and private participation is that for the sanitation field, these funding options combine better with certain type of sanitation systems, characterized in terms of their size and used technology (Table 14).
4.06.10.4 Differences between Low- and Middle-Income Countries Low-income countries need to invest 10–30% of their GDP to fulfill their MDGs (Lenghton et al., 2005). For some, these are figures difficult to reach even if the use of loans is considered. For them, external donors can play an important role. Middleincome countries have fewer needs and more economical capacity to meet their MDGs. For some, it is estimated that they could use up to 15% of their GDP, and hence it is considered that no external finance is needed (Lenghton et al., 2005). Moreover, this situation, from the point of view of some authors, offers to inform the private sector of great opportunities to conduct a business and, as a result, in several middle-income countries private funding is being promoted. One possible risk, which needs to be considered by local government and known by society in general, is that through private participation and international loans, technology and sanitation schemes from other countries are promoted, which
Receipts of royalties and licence fees (USD/person) 2004 120
109.3
100 80
do not always effectively solve local problems in the cheapest and most efficient way. Another risk is the use of the money for additional purposes. To deal with this, it is important, on the one hand, for the government to be accountable and, on the other hand, for society to demand transparency. In any case, it is certain that developing countries need to be creative to raise funds for sanitation. One option is to raise them as part of other projects in which sanitation can be a component; these include those considering goals for food security, health, land remediation, environmental problems control, and adaptation to climate change, for which several donors may be available. As an example, carbon credits could be used to fund projects to manage sludge and fecal sludge.
4.06.11 Science and Innovation: Need to Develop Individual Knowledge In developed countries, a complex and complete system of public agencies, private companies, equipment vendors, consultants, scientists, engineers, operators, and supporting professional and educational organizations makes sanitation possible. Promoting this organizational and human capacity in developing countries is one of the challenges on the path to increasing adequate sanitation, wastewater reuse, and proper fecal sludge and wastewater sludge management. Science and innovation are needed in developing countries to reduce their intense dependence on developed countries. Unfortunately, in many situations, technology originating in high-income countries is still preferred and implemented. However, this may not match the actual needs or promote local Table 15 Information concerning the first three Prosab research phases (with information from Andreoli et al. (2008))
60 40
0.8
20 0
High inco
me
Middle income
Area
Number of projects
Public resources (million USD)
Water Wastewater Sludge Solid waste
12 30 16 13
2509.00 3931.00 1845.00 1548.00
Total
71
9833.00
17.3
0
Low inco
me
World
Figure 16 Receipts of royalties and license fees in countries with different income (with information from UNDP, 2006).
Box 16 Research program for sanitation in Brazil (with information from Andreoli et al., 2008, Garbossa LHP, Lupatini G, and Pegorini ES (2008) Brazil. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 131–146. Vienna: UN.). The Brazilian Sanitation Research Programme (Prosab) is a public program that has received financial support for different projects since 1996. Its goal is to develop and optimize existing technologies for water supply, wastewater treatment, and solid residues management. For that, its objectives are * *
* *
to establish the state of the art of technology; to adapt or develop technology to provide sanitation services in local and regional conditions, and to meet the different needs of all population sectors, preserving and restoring the environment; to make technology and knowledge part of the public domain; and to support participatory processes, creating cooperative research networks to discuss subjects.
The total investment for the three phases listed is around US$9 million distributed as shown in Table 15, in which investments made for salaries and scholarships are not considered. Both, research papers and technological innovation, were produced from this program.
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economic development. In some other cases, developing countries are even used as laboratory testing grounds for new magic solutions. In low- and middle-income countries, examples can be found where a significant part of the investment made for wastewater treatment plants is used to pay for the intellectual property rights of the processes, as happens with many other activities. In Figure 16, it is shown that royalties received because of patents in developing countries are nonexistent or low while those for developed countries are high; sanitation could be in the future another source of this dependency and inequity. On the top of this, some of these processes do not solve actual problems and, as a result, around the world, several places can be found where new solutions for providing sanitation to poor people have been installed in series unsuccessfully. This situation has two negative effects: first, it discourages donors from making further investments and, second, it makes local people wary of possible solutions. The only way to prudently overcome this is to promote the development of technology by people immersed in local problems. For this purpose, investment in education and local research is important (Box 16 and Table 15). As presented here, the solution to sanitation problems can be combined with the solutions to other problems. The possibility therefore exists to develop new and individual technologies, to adapt the existing ones, and even to rediscover ancient local solutions. In parallel, the same can be done with policies to manage water.
4.06.12 Conclusions At an international level, there is current mobilization to support and improve sanitation conditions in developing countries. This mobilization is being expressed in terms of donors, private participation, and international aid agencies support. From this chapter, it is concluded that there are many reasons explaining why providing sanitation in developing countries is different to the solutions implemented in developed ones; therefore, care must be taken to not to use the aid to implement projects, which may prove not successful. For this reason, it is important to promote that each country defines first its needs and works defining programs. As the challenges to provide sanitation are many and very complex (policy definition, technologies to be used, education and awareness programs implementation, development of adequate institutional capacity, finding new financing options, etc.) it is important for developing countries to share among them their knowledge and experiences in the framework of the so-called South–South cooperation. Sanitation is an important pillar to develop wealthy societies (in terms of health and economic capacity) and, for this reason, governments should promote investments in this field that are to be properly and responsible managed. The only way to assure this is to promote, allow, or to demand a participatory approach. Finally, the water situation in developing countries has some bright sides. The first consists in the fact that the wide divisions observed in developed countries within the water sector (water supply and wastewater experts) does not exist or is not so pronounced. This allows easier understanding and
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promotes the integrated management of the problem. The second has to do with the high degree of solidarity existing among the population, which may play an important role in speeding up a sanitation program proven successful and contributing to raising the quality of life.
References Andreoli CV, Garbossa LHP, Lupatini G, and Pegorini ES (2008) Brazil. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 131--146. Vienna: UN. Angoua KM (2008) Cote d’Ivoire. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 269--277. Vienna: UN. Asano T (1998) Wastewater Reclamation and Reuse, Vol. 10: Water Quality Management Library. Lancaster, PA: Technomic Publishing. Asano T (2009) The role of wastewater reuse in water resources management. In: Primer Simposio Internacional del Caalca. Monterrey, Mexico, 13–14 April (in Spanish). Ba S (2008) Senegal. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 487--494. Vienna: UN. Bahri A (2008) Water reuse situation on the Middle Eastern and North African countries. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs, pp. 27--48. London: IWA Publishing. Bouwer H (2002) Artificial recharge of groundwater: Hydrogeology and engineering. Hydrogeology Journal 10: 121--142. Brissaud F and Salgot M (1994) Infiltration percolation as a tertiary treatment. In: Colloque Scientifique Et Technique International, ‘‘Mieux Gerer L0 Eau’’, pp. 391–398. Marseilles, France. Campos JR (1999) Tratamento De Esgostos Sanitarios Por Processo Anaero´bio E Disposicao Controlad No Solo, 1st edn. Sa˜o Carlos, Brazil: Prosab. (in Portuguese). CONAGUA and WWC (2006) Regional Document for the Americas Prepared for the 4th World Water Forum. Ciudad de Me´xico, Mexico, 16–22 March. Correlje AF and Schuetze T (2008) Every Drop Counts: Environmentally Sound Technologies for Urban and Domestic Water Use Efficiency. Division of Technology, Industry and Economics, TU Delft. India: United Nations Environment Programme. Dillon P and Jime´nez B (2008) Water reuse via aquifer recharge: Intentional and unintentional practices. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA Publishing. Dillon P and Toze S (eds.) (2005) Water quality improvements during aquifer storage and recovery, Project #2618, AWWARF. Dobie P (2001) Poverty and the Drylands. Nairobi: United Nations Development Programme, Drylands Development Centre. Ensink J, Mahmood T, Van der Hoek W, Raschid-Sally L, and Amerasinghe F (2004) A nationwide assessment of wastewater use in Pakistan: An obscure activity or a vitally important one? Water Policy 6: 197--206. Feachem R, Bradley D, Garelick H, and Mara D (1983) Sanitation and Disease: Health. pp. 349–356. New York, NY: Wiley. Foster S, Gardun˜o H, Tuinhof A, Kemper K, and Nanni M (2003) Urban Wastewater as Groundwater Recharge Evaluating and Managing the Risks and Benefits, GWMATE Briefing Note Series No. 12. Oxford: World Bank. Funamizu N, Huang X, Chen GH, Jiangyong H, and Visvanathan C (2008) Water reuse in Asia. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues, and Needs. London: IWA Publishing. Godfrey S, Labhasetwar P, Swami A, Parihar G, and Dwivedi H (2007) Water safety plans for grey water in tribal schools. Waterlines 25(3): 8--10. Gray JL and Sedlak DL (2003) Removal of 17-b-estradiol and 17-a-ethinyl estradiol in engineered treatment wetlands. In: International Conference on Pharmaceuticals and Endocrine Disrupters. National Ground Water Association, Minneapolis, MN, 19–21 March. Hrudey SE and Hrudey EJ (eds.) (2004) Safe Drinking Water: Lessons from Recent Outbreaks in Affluent Nations. London: IWA Publishing.
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Hughes R, Ho G, and Kuruvilla M (2006) Conventional small and decentralized wastewater systems in developing countries. In: Ujang Z and Henze M (eds.) Principles and Engineering. Lyngby, Denmark: IWA Publishing. Jekel M and Gruenheid S (2008) Indirect water reuse for human consumption in Germany – the case of Berlin. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA Publishing. Jime´nez B (2003) Health risks in aquifer recharge with recycle water. In: Aertgeerts R and Angelakis A (eds.) State of the Art Report Health Risk in Aquifer Recharge Using Reclaimed Water, pp. 54--172. Rome: WHO Regional Office for Europe. Jime´nez B (2006) Irrigation in developing countries using wastewater. International Review for Environmental Strategies 6(2): 229--250. Jime´nez B (2008a) Helminth ova control in wastewater and sludge for agricultural reuse. Water reuse new paradigm towards integrated water resources management. In: Grabow WOK (ed.) Encyclopedia of Biological, Physiological and Health Sciences, Water and Health, Vol. II. Life Support System, pp. 429--449. Oxford: EOLSS Publishers/UNESCO. Jime´nez B (2008b) Unplanned reuse of wastewater for human consumption: The Tula valley, Mexico. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA Publishing. Jime´nez B (2009a) Coming to terms with nature: Water reuse new paradigm towards integrated water resources management Encyclopedia of Biological, Physiological and Health Sciences, Water and Health, Vol. II: Life Support System, pp. 398--428. Oxford: EOLSS Publishers/UNESCO. Jime´nez B (2009b) Wastewater risks in the urban water cycle. In: Jime´nez B and Rose J (eds.) Urban Water Security: Managing Risks, p. 324. Paris: UNESCO Leiden: Taylor and Francis Group. Jime´nez B and Asano T (eds.) (2008) Water reclamation and reuse around the world. In: Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA Publishing. Jime´nez B, Austin A, Cloete E, and Phasha C (2006) Using Ecosan sludge for crop production. Water Science and Technology 5(54): 169--176. Jime´nez B and Gardun˜o H (2001) Social, political and scientific dilemmas for massive wastewater reuse in the world. In: Davis C and McGinn RE (eds.) Navigating Rough Waters: Ethical Issues in the Water Industry. American Water Works Association Jime´nez B and Wang L (2006) Sludge treatment and management. In: Ujang Z and Henze M (eds.) Municipal Wastewater Management in Developing Countries: Principles and Engineering, pp. 237--292. London: IWA Publishing. Juanico´ M and Milstein A (2004) Semi-intensive treatment plants for wastewater reuse in irrigation. Water Science and Technology 50(2): 55--60. Keraita B, Jime´nez B, and Drechsel P (2008) Extent and implications of agricultural reuse of untreated, partly treated and diluted wastewater in developing countries. Cab Reviews: Perspectives in Agriculture, Veterinary Science, Nutrition and Natural Resources 3(58): 15. Kone´ D (2010) Making urban excreta and wastewater management contribute to cities’ economic development: A paradigm shift Water Policy, Vol. 12, No. 4, pp. 602–610. LeBlanc R, Matthews P, and Roland P (2008) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource UN-HABITAT, Vienna, 632pp. Lenghton L, Wright A, and Davis K (eds.) (2005) Health, Dignity and Development: What Will It Take? Millennium Development Goals. London: Earthscan. Mamadou SD (2008) Mali. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 413--418. Vienna: UN. Mara D (2004) Domestic Wastewater Treatment in Developing Countries. London: Earthscan. Maya C, Jime´nez B, and Schwartzbrod J (2006) Comparison of techniques for the detection of helminth ova in drinking water and wastewater. Water Environment Research 78(2): 118--124. Mfoulu N (2008) Cameroon. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 169--179. Vienna: UN. Murray C and Lo´pez A (1996) Global Health Statistics. Cambridge: Harvard University Press. Nelson K, Jime´nez B, Tchobanoglous G, and Darby J (2004) Sludge accumulation, characteristics, and pathogen inactivation in four primary waste stabilization ponds in central Mexico. Water Research 38(1): 111--127. Paskalev A (2008) Bulgaria. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 149--153. Vienna: UN.
Rusong W (2001) System consideration of eco-sanitation. In: China Proceedings of the First International Conference on Ecological Sanitation. Nanning, China, 5–8 November. Shuval HI, Adin A, Fattal B, Rawitz E, and Yekutiel P (1986) Wastewater irrigation in developing countries: Health effects and technical solutions. World Bank Technical Paper No. 51. The World Bank, Washington. Silva N, Chan M, and Bundy A (1997) Morbidity and mortality due to Ascariasis: Reestimation and sensitivity analysis of global numbers at risk. Tropical Medicine International Health 2(6): 19--28. SIWI-IMWI (2006) Water – more nutrition per drop. Towards sustainable food production and consumption patterns in a rapidly changing. In: World Stockholm International Water Institute (SIWI) and the International Water Management Institute, p. 36. Stockholm, Sweden. Smakhtin V, Carmen R, and Do¨ll P (2004) Taking into account environmental water requirements in global-scale water resources assessments. Comprehensive Assessment of Water Management in Agriculture, Research Report 2. Colombo: International Water Management Institute. Snelling W, Xiao L, Ortega-Pierres G, et al. (2007) Cryptosporidiosis in developing countries. Journal of Infection in Developing Countries 1(3): 242--256. Snyman F (2008) South Africa. Faecal sludge management. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 514--516. Vienna: UN. Spaliviero M and Carimo D (2008) Mozambique. In: LeBlanc RJ, Matthews P, and Richard RP (eds.) Global Atlas of Excreta, Wastewater Sludge, and Biosolids Management: Moving Forward the Sustainable and Welcome Uses of a Global Resource: UNHSP, pp. 431--437. Vienna: UN. UN (2003) Water for people, water for life. The United Nations World Water Development Report. Barcelona, Spain: UNESCO. UNDP (2006) United Nations Development Programme Human Development Report 2006 Beyond Scarcity: Power, Poverty and the Global Water Crisis. New York, NY: Palgrave Macmillan. UNEP (2002) International Source Book on Environmental Sound Technologies for Wastewater and Stormwater Management, United Nations Environment Programme, International Environmental Technology Centre, Osaka. pp. 319–398. London: IWA Publishing. UN-Habitat (2006) The State of the World’s Cities Report 2006/7; The Millennium Development Goals and Urban Sustainability: 30 Years of Shaping the Habitat Agenda. London: Earthscan. UN/WWAP (/WWAP, 2003) United Nations/World Water Assessment Programme. USA: UN. US-EPA (1992) Guidelines for Water Reuse. Washington, DC: Office of Wastewater Enforcement and Compliance. Van de Guchte C and Vandeweerd V (2004) Targeting sanitation. Our Planet 14(4): 19--21. Van der Merwe N, du Pisani P, Menge J, and Ko¨nig E (2008) Water reuse in Windhoek, Namibia: 40 years and still the only case of direct water reuse for human consumption. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA Publishing. WHO (1989) Guidelines of the Safe Use of Wastewater and Excreta in Agriculture and Aquaculture. Prepared by D. Mara and S. Cairncross: Geneva: WHO. WHO (2004) Guidelines For Drinking-Water Quality: Recommendations, 3rd edn., vol. 1. Hong Kong, China: WHO. WHO (2006) Guidelines for the Safe Use of Wastewater, Excreta and Greywater, Vol. 2: Wastewater Use in Agriculture. Geneva: WHO. WHO/UNICEF (2000) Global Water Supply and Sanitation Assessment Report, Joint Monitoring Programme for Water Supply and Sanitation. Geneva: WHO. WHO/UNICEF (2004) Meeting the MDG Drinking Water and Sanitation Target: A MidTerm Assessment of Progress. Geneva: WHO and UNICEF. WHO–UNICEF (2006) Meeting the MDG Drinking Water and Sanitation Target: The Urban and Rural Challenge of the Decade. Geneva: WHO and UNICEF. WHO–UNICEF (2008) Progress on Drinking Water and Sanitation: Special Focus on Sanitation. Geneva: WHO and UNICEF.
Relevant Websites http://www.windhoekcc.org.na City of Windhoek. http://earthtrends.wri.org Earth Trends: Environmental Information; Earthtrends 2009.
4.07 Source Separation and Decentralization TA Larsen and M Maurer, Eawag, Swiss Federal Institute of Aquatic Science and Technology, Du¨bendorf, Switzerland & 2011 Elsevier B.V. All rights reserved.
4.07.1 4.07.2 4.07.2.1 4.07.2.2 4.07.2.3 4.07.2.3.1 4.07.2.3.2 4.07.2.4 4.07.2.5 4.07.2.5.1 4.07.2.5.2 4.07.2.5.3 4.07.2.5.4 4.07.2.5.5 4.07.2.6 4.07.3 4.07.3.1 4.07.3.2 4.07.3.3 4.07.3.3.1 4.07.3.3.2 4.07.3.4 4.07.3.5 4.07.3.5.1 4.07.3.5.2 4.07.3.5.3 4.07.3.5.4 4.07.3.5.5 4.07.3.6 4.07.4 4.07.4.1 4.07.4.2 4.07.4.3 4.07.4.3.1 4.07.4.3.2 4.07.4.4 4.07.4.5 4.07.4.5.1 4.07.4.5.2 4.07.4.5.3 4.07.4.5.4 4.07.4.5.5 4.07.4.6 4.07.5 4.07.5.1 4.07.5.2 4.07.5.3 4.07.5.3.1 4.07.5.3.2 4.07.5.3.3 4.07.5.4 4.07.6 References
Introduction Gray Water Production Rate and Composition of Gray Water Reuse Purposes and Regulation The Risks of Gray Water Reuse Hygienic risks The risks of organic contaminants and salts Decision Making and Public Perception of Gray Water Recycling Treatment Technologies for Gray Water No treatment Storage Physical–chemical treatment Biological treatment Disinfection and removal of micropollutants Summary Urine Production Rate and Composition of Urine Reuse Purpose and Regulation Risks Associated with Source Separation of Urine Hygienic risks The risks of organic contaminants and salts Public Perception of Urine Source Separation Treatment Technologies for Urine No treatment Storage Physical–chemical treatment Biological treatment Hygienization and removal of micropollutants Summary Feces Production Rate and Composition of Feces Reuse Purposes and Regulation The Risks of Source Separation of Feces Hygienic risks The risks of organic contaminants Decision Making and Public Perception of Source Separation of Feces Treatment Technology for Feces No treatment Storage Physical–chemical treatment Biological treatment Hygienization and removal of micropollutants Summary Combined Domestic Wastewater Production Rate and Composition of Combined Domestic Wastewater The Risks of On-Site Treatment of Combined Wastewater Two Examples of On-Site Treatment Technologies for Combined Domestic Wastewater Septic tanks Johkasous On-site uncontrolled anaerobic digestion Summary Outlook
203 204 204 205 205 206 206 207 207 207 208 208 209 210 210 211 211 211 211 212 212 212 213 213 213 213 214 215 215 216 216 216 216 216 216 216 217 218 218 218 219 221 221 221 221 221 222 222 222 222 223 223 224
203
204
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4.07.1 Introduction Toward the end of the 1990s, a special issue on sustainable sanitation was published by Water Science and Technology, a journal of the International Water Association (IWA). In this issue, several authors discussed the issue of source-separation technologies in the area of wastewater management (Henze, 1997). All major sources of combined wastewater were covered: industrial wastewater (Visvanathan and Hufemia, 1997), storm water (Boller, 1997), gray water (domestic wastewater without toilet waste; Jeffrey et al., 1997), urine (Hana¨us et al., 1997; Larsen and Gujer, 1997), and black water (combined toilet waste; Otterpohl et al., 1997). Whereas separate discharge of storm water and separate treatment of industrial wastewater were already well-known concepts, source separation for domestic wastewater was at that time primarily recognized as an inexpensive sanitation technology for poor, rural areas. For urban areas, only sewerbased wastewater management was considered suitable. The above-mentioned special issue challenged this conventional wisdom with the idea that source separation could potentially compete with end-of-pipe technologies for all wastewaters, even in urban areas. The main concepts behind this idea were that, on the one hand, it would be more resource efficient to treat very concentrated solutions such as urine (Larsen and Gujer, 1997) and that, on the other hand, local recycling of (gray) water would increase flexibility of the urban water management system (Jeffrey et al., 1997). Since 1997, the acceptance of resource efficiency as a leading principle for sustainable urban water management has gained ground, recently resulting in the publication of a new paradigm for sustainable wastewater management based on resource recovery (Guest et al., 2009). Additionally, a number of studies have been published showing that source separation has the potential of being more resource efficient than end-ofpipe technology, depending on the specific choice of technology (e.g., Lundin et al., 2000; Hellstro¨m et al., 2008; Remy and Jekel, 2008). Moreover, it has also been shown that the evaluation of source-separating technologies is sensitive toward stakeholder preference (Borsuk et al., 2008), and also Guest et al. (2009) emphasize the importance of stakeholders for decision making. Additionally, emerging issues such as the problem of micropollutants may call for new approaches to urban water management and, again, we and other authors have argued that this problem may be best tackled close to the source (Larsen et al., 2004; Kujawa-Roeleveld and Zeeman, 2006; Joss et al., 2008). For source-separated waste streams, two different management approaches are possible: either on-site treatment or transport to a plant for (semi)centralized treatment. Most approaches to source separation are more decentralized than typical wastewater treatment plants; however, some authors explicitly foresee transport of the different source-separated waste streams to a (semi)centralized plant (see, e.g., Oldenburg et al., 2007). Others, for example, Larsen et al. (2009), argue that at least for urine, an on-site approach may be the more productive road to take. In this chapter, we explicitly look at decentralized treatment of source-separated waste streams, but we do not limit the scope to any specific size of the system. Besides, in most cases, the technologies in
question are still so new that issues that are more fundamental are of interest than the exact scale of application. A typical argument for centralized solutions is the possibility to achieve economies of scale in treatment plants (see compilation in Maurer (2009)), ignoring the tendency of conveyance systems to show a diseconomy of scale or to be scale neutral at best (Adams et al., 1972; Maurer et al., 2009). Only where sewers are deemed too expensive are decentralized technologies considered. With technical development, decentralized technologies may become more competitive, as exemplified by the progress in membrane technology (DiGiano et al., 2004). Membranes are becoming better and cheaper, due to technological development and mass production and, at the same time, the economic costs for sewers may increase due to growing pressures, such as the effects of climate change and increasing planning uncertainty, thereby reducing the useful functional life span of the sewer infrastructure (Maurer, 2009). As discussed by Larsen and Gujer (2001), decentralized treatment options could become more attractive if treatment technology for source-separated waste streams becomes integrated into household technology instead of the prototype wastewater treatment plants that we know today. For on-site treatment of urine, for example, we have calculated that about 260–440 USD/person would be available for investment in household technology for nutrient elimination directly from urine (Maurer et al., 2005), a benchmark which seems reachable with mass-produced devices. In the following sections, we give an overview of the fields of source separation for gray water, urine, and feces. Gray water refers to combined domestic wastewater without toilet waste. Source-separated urine may or may not include flush water. Feces can be collected with or without urine, and with or without flush water. Water-diluted feces including urine are conventionally known as black water, whereas water-diluted feces without urine have been termed brown water. Additionally, we added a short section on the on-site treatment of combined domestic wastewater (gray water þ black water). We base this overview on an extensive literature review, and it is no coincidence that the literature list contains very few items prior to 1996. The field of source separation for urban wastewater management is still very immature and this is why we deliberately refrain from drawing any conclusions with respect to the suitability of the different approaches. We intend to encourage the development of the entire field and the readers must draw their own conclusions. All the sections on source-separated waste streams have the same organization: after a short introduction of the field, we look at production rate and composition, reuse purposes and regulation, risk aspects, and public perception. At the end of each section, we review the literature on the different treatment options for the waste source in question. The literature list is long and we hope that this chapter will encourage interested readers to delve into the fascinating new field of source separation for sustainable urban water management.
4.07.2 Gray Water The most obvious target for source separation is gray water (combined wastewater without toilet waste) because of its
Source Separation and Decentralization
value as an alternative to drinking water, especially for nonpotable purposes. With the combined effect of increasing water demand from a growing world population and higher incidences of drought in many areas due to climate change, alternative water sources such as gray water will obviously gain importance in the decades to come. For logistic reasons, a decentralized approach to gray water recycling for nonpotable purposes is often taken. It is attractive to avoid a second distribution net, and for water, which is not of drinking water quality, a shorter residence time in the system is favorable. This approach will however also have to be considered from a risk perspective as discussed in Section 4.07.2.3. Gray water transport to a (semi)centralized treatment plant and back again to the households as nonpotable water as suggested by Bingley (1996) is not discussed here. An alternative to the decentralized recovery of gray water would be a decentralized treatment of toilet waste only, leaving the possibilities for a centralized treatment of gray water to drinking-water quality. This would even be possible via conventional wastewater treatment, infiltration, and drinking water treatment – an approach that is today often adopted more or less deliberately with combined wastewater (Wintgens et al., 2005). Without toilet waste, this approach could appear more appealing to the public, avoiding the toilet-to-tap notion. Moreover, the problems of accumulated particulate matter in sewers caused by steadily progressing water-saving measures (see Section 4.07.2.2 for a discussion) will be much less severe if the sewers are used only for gray water transport. However, such an approach would require centralized planning and would not allow the people directly concerned by water scarcity to solve their immediate problem. In this section, we confine our discussion to the decentralized approach to gray water treatment and reuse. In most cases, water reuse is the dominating reason for decentralized gray water treatment; however, where no sewers are available, urban hygiene or environmental protection may be in the foreground (see, e.g., Carden et al., 2007). Gray water, which refers to domestic wastewater without toilet waste, is generally considered more attractive for reuse than combined wastewater, from the point of view of both esthetics and pathogenic organisms. However, there are different types of gray water. Often, gray water from bath, shower, and washbasin is termed light gray water, whereas gray water from kitchen and laundry is termed dark gray water. It is disputed which type of gray water is best suited for reuse. Whereas Christova-Boal et al. (1996) do not recommend the reuse of kitchen gray water because they consider it as highly polluted and a source of many undesirable compounds, for example cooking oil, Li et al. (2009) suggested that for better
Table 1
205
treatability, kitchen gray water should always be included where biological treatment is foreseen.
4.07.2.1 Production Rate and Composition of Gray Water The amount of gray water produced greatly varies from 15 l/ person/day in rural areas of water-scarce countries such as Jordan (Halalsheh et al., 2008) to more than 100 l/person/day in many parts of Europe. Consequently, the strength of gray water also varies, even without accounting for the large differences between light and dark gray water. In different literature reviews, very different ranges of gray water concentrations are reported, often due to the fact that the type of gray water is not indicated. For instance, a thorough literature review by Eriksson et al. (2002) revealed concentrations of organic matter in gray water in the range from 13 to 8000 mgCOD l1, which is not really helpful for a characterization. A number of authors have estimated the typical daily load of organics to different types of gray water (Table 1). Since these estimates are already corrected for different water-consumption patterns, they should be more robust than the concentrations measured in different settings. However, these estimates stem from a limited number of European countries and care should be taken to extrapolate to other cultures. It is generally understood that the volume of kitchen gray water is small (e.g., 5% of total gray water (Christova-Boal et al., 1996); 20% (Almeida et al., 1999); or 30% (Friedler, 2004)), but containing a substantial part of the organic matter (expressed as chemical oxygen demand (COD); e.g., 40% (Almeida et al., 1999) or 42% (Friedler, 2004)). There are, however, authors, who see this differently. For instance, Henze and Ledin (2001) allocated 50% of gray water volume and 85% of gray water COD to kitchen gray water. According to Bester et al. (2008), household wastewater is today one of the most important sources of xenobiotics to the urban water cycle, whereas the relevance of industrial point sources is decreasing. The occurrence of xenobiotics in gray water has been documented by very few authors. Based on the information available on the composition of common Danish household products, Eriksson et al. (2002) identified at least 900 different organic chemical substances and compound groups in gray water. Based on an environmental hazard identification of 211 of these compounds, the authors categorized 66 compounds as priority pollutants; 34 of these were different types of surfactants. The remaining 700 compounds could not be evaluated. In an analytical gas chromatography–mass spectrometry (GC-MS) screening study, Eriksson et al. (2003) identified 191 different organic compounds in gray water from an apartment building in
Suggested values for typical gray water production (water and total organic matter)
Volume (l/cap/day) COD (gCOD/cap/day) N (gN/cap/day) P (gP/cap/day)
Henze and Ledin (2001)
Almeida et al. (1999)
DWA (2008)
Vinnera˚s et al. (2006a)
Average
99 54 1.9 0.47
71 63
108 47 1.0 0.50
100 52 1.4 0.52
94 54 1.4 0.5
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Copenhagen. The concentration of the compounds was semiquantitatively assessed and found to be within the range of 101–102 mg l1. Based on a questionnaire, it was possible to predict the presence of many, but far from all of these, compounds. The reasons for this were partially lack of detailed product information, and partially incomplete information from the tenants. In a search for hazardous substances, Palmquist and Hana¨us (2005) looked specifically for 81 organic substances, and found 46 of them in gray water from a Swedish housing area with separate collection of gray water. In addition, the content of microbial contamination in gray water greatly varies. Typical literature values are summarized in Table 2, and compared to typical values for combined wastewater. A more detailed discussion of microbial indicators of gray water quality can be found, for example, in Albrechtsen (2002), Winward et al. (2008a), and Gilboa and Friedler (2008).
4.07.2.2 Reuse Purposes and Regulation Reclaimed gray water is typically intended for toilet flushing, cleaning purposes, car washing, and irrigation. Toilet flushing is the typical example of indoor use, and irrigation the typical example of outdoor use, and most advantages and risks associated with decentralized reuse of gray water can be illustrated by these two applications. There are many different quality standards for recycled wastewater, with an emphasis on hygienic quality, biochemical oxygen demand (BOD), and turbidity. In Table 3, we list typical ranges of mandatory quality parameters that should not be surpassed, but for any practical purpose, one will of course have to adhere to local Table 2
regulations. Most noticeable is the large deviation between the requirements, even for the same reuse purpose. Li et al. (2009) made a suggestion for new guidelines, but without presenting any risk analysis. Examples of such risk analyses can be found in Ottoson and Stenstro¨m (2003) and Huertas et al. (2008). If gray water is only used for toilet flushing, there is a clear limit to the water savings that can be achieved. Today, it is often assumed that use of gray water for toilet flushing will save around 30% of the total drinking-water consumption. Modern toilets, however, use 6 l for a large flush and 3 l for a small flush, and even smaller flush volumes have been attempted (e.g., a 4/2-l toilet). A typical person gives rise to one large and four to five small flushes a day, which with a 6/3-l flush toilet amount to 18–21 l/person/day (see also Table 7 for an outlook on future water-saving toilets). From Table 1 it is thus obvious that in the long term only a smaller part of the total gray water production will be required for this special purpose. The requirement of water for irrigation can of course only be quantified for a specific setting. Water saving is generally considered environmentally friendly, but as shown by Parkinson et al. (2005), a reduction of water consumption in a conventional setting based on combined sewers is not without problems. Especially the high water-saving capacity of gray water reuse for flushing of oldfashioned toilets will typically lead to a higher rate of sedimentation in sewers, with higher emissions of pollutants when these sediments are mobilized during combined sewer overflow events. Furthermore, anaerobic degradation of the organic sediments may lead to methane emission and sewer corrosion. If gray water reuse (or alternatively water saving) is
Typical content of microorganisms in gray water compared to typical values in combined wastewater
Parameter
Unit
Gray water
Bathroom
Laundry
Combined wastewater
Total coliform Fecal coliform
# ml1 # ml1
105–107 101–106
100–105 ND*–103
100–104 101–102
109–1011
ND*, nondetectable in 100 ml. Based on data compiled by Henze M and Ledin A (2001) Waste and wastewater characteristics and its collection. In: Lens P, Zeeman G, and Lettinga G (eds.) Decentralised Sanitation and Reuse, Integrated Environmental Technology Series, pp. 56–72. London: IWA; Eriksson E, Auffarth K, Henze M, Ledin A (2002) Characteristics of grey wastewater. Urban Water 4(1):85–104, Ottoson J, Stenstro¨m T.A (3003) Faecal contamination of greywater and associated microbial risks. Water Research 37(3): 645–655; Friedler E, Kovalio R, Ben-Zvi A (2006) Comparative study of the microbial quality of greywater treated by three on-site treatment systems. Environmental Technology 27(6): 653–663; and Birks R and Hills S (2007) Characterisation of indicator organisms and pathogens in domestic greywater for recycling. Environmental Monitoring and Assessment 129(1–3): 61–69.
Table 3
Examples of regulation of gray water use (maximum values, unless otherwise stated)
Parameter
Unit
Domestic reuse
Toilet flushing
Irrigation
Li et al. (2009)a
Turbidity BOD Total coliform Fecal coliform Residual chlorine
NTU mg l1 # ml1 # ml1 mg l1
1–90 10–45 NDb–100 4–20 40.2 to 41
5 5–20 100–1000 0.03–10 40.2 to 41
20
2 10 100 10 Z1
a
50 0.03–10 40.2 to 40.4
Suggested guideline for unrestricted urban reuse. ND in 100 ml. Based on values compiled by Jefferson B, Laine A, Parsons S, Stephenson T, Judd S (2002) Technologies for domestic wastewater recycling. Urban Water 1(4): 285–292, Eriksson E, Auffarth K, Henze M, Ledin A (2002) Characteristics of grey wastewater. Urban Water 4(1): 85–104, Winward G.P, Avery L.M, Frazer-Williams R, et al. (2008a) A study of the microbial quality of grey water and an evaluation of treatment technologies for reuse. Ecological Engineering 32(2): 187–197, and Li F, Wichmann K, Otterpohl R (2009) Evaluation of appropriate technologies for grey water treatments and reuses. Water Science and Technology 59(2): 249–260. b
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the only decentralized measure taken in a catchment, this potential effect should be analyzed.
4.07.2.3 The Risks of Gray Water Reuse Gray water is contaminated by pathogens and carries a large amount of potentially problematic micropollutants (Eriksson et al., 2002, 2003; Palmquist and Hana¨us, 2005). In order to control the risks of gray water reuse, Salgot et al. (2006) recommended an elaborate scheme of surveillance based on a consideration of analytical costs. Obviously, with decreasing size of the recycling scheme, the possibilities of monitoring decrease. Even in a centralized setting, a dual pipe system with nonpotable water can be problematic. A recent large pilot study in the Netherlands (Oesterholt et al., 2007) with production of nonpotable water from surface, rain, and groundwater led to health problems. The main problems were linked to cross-connections in the dual pipe system, drinking of tap water intended for outdoor use, and the production of aerosols during toilet flushing. It was also observed that biological instability of the water leads to growth of Legionella. As a consequence, all pilot studies were canceled and larger settings with dual pipe systems forbidden in the Netherlands. Smaller settings are still allowed, but only with rainwater or groundwater as source. Although the Dutch experience was gained in a centralized and not a decentralized setting, some conclusions may still be drawn:
• • • •
the more complex the dual system is, the larger the risk of cross-connections; outdoor taps lead to a larger risk of drinking nonpotable water than toilet flushing; biological stability to prevent growth of Legionella must be provided; and the justification of any wastewater reuse for nonpotable purposes must be carefully examined.
4.07.2.3.1 Hygienic risks Generally, fecal contamination is considered the major hygienic risk from recycling of gray water. This is reflected in the use of fecal coliforms as an indicator organism in gray water (see Table 3). Although virus may be more critical than bacteria, many authors still favor fecal coliforms as an indicator (see, e.g., Dixon et al., 1999b). For a discussion of more advanced risk models, see, for example, Ottoson and Stenstro¨m (2003) or Huertas et al. (2008). The main discussion of hygienic risks connected to decentralized gray water reuse is the question of scale. Dixon et al. (1999b) stated the importance of population size for the range of risk, and distinguished between the multiuser and single family water reuse, where for the latter the hygienic risks are obviously much lower. This is also reflected in public perception as reported by Jeffrey and Jefferson (2001): people are generally more favorable with respect to recycling their own gray water for toilet flushing than using gray water from somewhere else. The authors contrast this with a possible lower health risk from more centralized schemes, where
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authorities can control the quality of the recycled gray water. The most critical situation, however, may occur in the middle situation, as illustrated by Albrechtsen (2002): a gray water recycling scheme in a multistory building is large enough that one family with a contagious illness can put all the other families at risk, but still too small for efficient monitoring of treatment efficiency. For pathogens such as Legionella, which may proliferate in a technical system for nonpotable water, especially if the water has not been biologically treated (Oesterholt et al., 2007), the risk would however also be present at a very small scale. Although we have found no examples in the literature of problems occurring due to Legionella in an existing gray water system, gray water recycling systems offer favorable conditions for their growth (biofilms and high temperatures), which must therefore be prevented (Dixon et al., 1999b). Consequently, for single family recycling of gray water, these authors suggest focusing more on system design (e.g., treatment of gray water and prevention of biofilm) than on measuring the microbial quality.
4.07.2.3.2 The risks of organic contaminants and salts Nonpathogenic pollutants from gray water are primarily considered a potential problem for the environment, and not for public health (Oesterholt et al., 2007). The major contributions to xenobiotics in the urban water cycle stem from household and service applications, including personal care products, detergents, etc., which will be present in gray water (Bester et al., 2008). Although the anthropogenic origin of organic compounds in gray water may lead to problems, the same anthropogenic origin also holds a potential for improvements at the source. Since most substances in gray water are not related to the improvement of human health, there may be more scope for their re-engineering than in the case of pharmaceuticals (see Sections 4.07.3.3.2 and 4.07.4.3.2). Besides salt in arid climate, organic compounds leading to a major change in soil structure may be the most serious threat to the sustainability of using gray water for irrigation. Two types of organics have been shown to increase water repellency of soils: surfactants, with a hydrophobic and a hydrophilic end (Wiel-Shafran et al., 2006), and oil and grease, which are strongly hydrophobic compounds (Travis et al., 2008). The former compounds originate mainly from bathroom and laundry gray water, whereas oil and grease are mainly found in kitchen gray water. In general, salts could be a major problem, as well as specific toxic inorganic compounds such as boron (Gross et al., 2005). Whereas the general problem of salt can only be solved by reducing the amount of salt-containing gray water used for irrigation or by separating salt and water by reverse osmosis (RO) or evaporation, a specific problem such as boron plant toxicity may be solved at the source. In Israel, for instance, the concentration of boron in detergents has been limited by regulation (Gross et al., 2005). For decentralized use of gray water for irrigation in the own garden, the household itself has a certain amount of control over the contaminations in the gray water. As discussed in Section 4.07.2.5.1, this may however not be of much use, if the gray water is not treated. Moreover, the risks are rather long term, which may be less important to a household that
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has to optimize its living on a short-term basis. Furthermore, through leakage of micropollutants to groundwater, the risks may spread to the larger community.
4.07.2.4 Decision Making and Public Perception of Gray Water Recycling Decision making in the area of decentralized gray water recycling is not easy. Scha¨fer and Beder (2006) analyzed the complexity of the issue within the context of the precautionary principle and discussed a number of uncertainties concerning human health, environmental quality, technology, politics, and socioeconomic issues. As a consequence of these uncertainties, the authors conclude that a precautionary approach requires transparency and public participation. However, only little work on public participation in the planning of gray water reuse has been internationally published. Based on a survey in the Melbourne area, ChristovaBoal et al. (1996) reported that people are generally interested in reusing gray water from the bathroom and from laundry, with a preference for using the recuperated water in the garden. However, people are interested to invest in gray water recycling only if the payback period is very short (2–4 years). Similarly, Neal (1996) found a very high approval rate of using gray water for irrigation (79% total support, 17% support a little, and nobody against). As discussed above, Jeffrey and Jefferson (2001) found that people prefer recycling their own gray water for toilet flushing rather than using gray water from somewhere else. In a hotel in Spain, however, guests were generally satisfied with using recycled gray water for toilet flushing, possibly due to the acute water scarcity experienced in the area (March et al., 2004). In a survey in Oman involving 1365 people, Jamrah et al. (2008) reported that 76% of the respondents accept recuperated gray water for irrigation, 66% for toilet flushing, and 53% for car washing. From a larger study in Australia, where the different perception of desalinated and recuperated gray water was compared, Dolnicar and Scha¨fer (2009) concluded that people generally accept recycled gray water as the more environmentally friendly option; however, from the point of view of public health, they consider desalinated water the safer alternative. People are also generally more disgusted about the idea of recycled gray water, despite an assumption of identical water quality. However, desalinated water is not generally preferred over recycled gray water, only for close-to-the-body applications. For irrigation, for instance, recycled gray water is preferred. This is in accordance with the older study by Christova-Boal et al. (1996) cited above.
4.07.2.5 Treatment Technologies for Gray Water From the above discussion, it is clear that gray water recycling will generally demand some sort of treatment, and, in this chapter, we briefly review the possible technologies. Most of them are well known from mainstream wastewater treatment and, for a more general introduction to the technologies, we refer to the standard wastewater treatment literature. As compared to treatment of urine and feces, the technology applications are very close to conventional treatment of combined wastewater.
The question preceding any discussion on treatment options is: What are the objectives of the treatment? For reuse of gray water, the treatment objectives are basically different from the aims of wastewater treatment. Whereas the question of receiving water quality dominates the discussion of conventional wastewater treatment, the risk in connection with reuse is the dominating issue of gray water treatment. The purpose of reuse determines the quality criteria; however, in nearly all cases, hygienic quality is of high importance. Since some storage and transport of the treated gray water before reuse will nearly always be necessary, the stability of the gray water is thus essential. Furthermore, for irrigation purposes, the risk of not only pathogens but also organic compounds is important (Section 4.07.2.3.2). The quality parameters are highly disputed, but we refer to the original literature for a discussion. The most important uncertainties are connected to the choice of indicator organisms for pathogens (see Section 4.07.2.3.1) and the potential risks of xenobiotics and surfactants (see Section 4.07.2.3.2). Furthermore, the possibilities of quality monitoring in decentralized settings are very limited, and robustness of treatment will therefore be one of the most important criteria for technology choice.
4.07.2.5.1 No treatment In many cases, gray water (or even combined wastewater) is reused without any treatment at all. Although this is not to be recommended, we have found one apparently successful example of intended gray water reuse without any treatment at all. This is the so-called hand basin toilet, where a small washbasin is placed above the flush water reservoir of the toilet (commercially available). The water used for hand washing thus flows directly into the reservoir. Apparently, the short transport and residence time prevent the buildup of odor, and the very local recycling without any risk of crossconnections limits the risk of distributing pathogens. We have found no information on the risk of Legionella, which could, in principle, grow on surfaces in the reservoir. The hand basin, however, is used in Australia, where drinking water is typically chlorinated; in countries without chlorination, problems even with this simple system could be larger. In all other cases, reuse of untreated gray water must be assumed to hold very large risks. Pathogens are a risk for all applications and, for irrigation, surfactants, oil, and grease represent a major risk to soil structure (see Section 4.07.2.3.2). It is interesting to note that some authors (e.g., Krishnan et al., 2008) assume that gray water from kitchen and laundry (dark gray water) is unlikely to cause severe pollution, because it is easily degraded in nature. For irrigation purposes, however, even degradability may not avoid problems. As shown by Gross et al. (2005) and Wiel-Shafran et al. (2006), surfactants may adsorb to soil particles and be resistant to biological degradation.
4.07.2.5.2 Storage If gray water is to be reused within the general water cycle of the household, some sort of storage will be necessary in order to balance production and demand. Based on experimental evidence and modeling, Dixon et al. (1999a) suggested that storage before treatment would be beneficial because
Source Separation and Decentralization
sedimentation in the storage tank would lead to a reduction of the organic load to the treatment stage. Depletion of oxygen and development of unpleasant odors due to anaerobic degradation of organic matter in the sediments would have to be counteracted through some sort of aeration, and, obviously, sludge should not be allowed to accumulate in the storage tank. Nolde (2000) presented an example of such a storage tank, but for practical reasons also recommended storage after treatment. One main purpose of the treatment technologies discussed subsequently is to stabilize the treated water for storage and prevent regrowth of microorganisms in the storage tank.
4.07.2.5.3 Physical–chemical treatment Although several authors state that for on-site treatment, only biological treatment will be able to stabilize gray water in order to prevent regrowth (e.g., Jefferson et al., 2000) and produce a stable sludge that will not give rise to odors (e.g., Abu Ghunmi et al., 2008), there are many attempts to treat gray water with physical–chemical methods, also in decentralized settings. The resulting gray water is stabilized by disinfection. From a process engineering point of view, there are good reasons for selecting a physical–chemical treatment method in a decentralized setting. The obvious advantage of nonbiological technologies as compared to biological technologies for on-site application is the higher resistance toward toxic chemicals and long absences. Moreover, the highly variable organic loads experienced in gray water, the large amount of nondegradable COD, and the shorter hydraulic residence times (HRTs) that can be obtained in a chemical– physical system would favor physical–chemical treatment (Rivero et al., 2006). Physical treatment. In a recent review, Li et al. (2009) compared the effectiveness of different physical treatments (different types of filtration, mainly sand and membrane filtration) and concluded that physical treatment alone is not sufficient to reach good gray water quality, except perhaps for very low loaded gray waters. It should, however, be noted that, for example, for toilet flushing, the amount of light gray water produced in a household would be sufficient (see Section 4.07.2.2). Nghiem et al. (2006) considered the use of on-site ultrafiltration (UF) for gray water recycling promising and also Friedler et al. (2008a, 2008b) suggested UF either alone or followed by RO as an attractive technology for decentralized gray water treatment. Direct membrane filtration takes up little space and delivers water of excellent quality, despite large fluctuations of gray water production rate and quality. The negative aspects of direct membrane filtration are connected to organic and biological fouling and inorganic scaling of the membranes, leading to rapid decline of flux. There are different ways of responding to this challenge. Friedler et al. (2008a, 2008b) tested two types of pretreatment: chlorination or coagulation with ferric chloride. Coagulation was found to be the better treatment method for reducing fouling in the UF membrane, and, in contrast to chlorination, it did not increase inorganic scaling of the RO membrane. Oschmann et al. (2005) and Nghiem et al. (2006), in contrast, investigated in detail the mechanisms leading to fouling in membrane
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reactors treating gray water. Especially calcium in combination with humic acids and particulate matter seems to play a major role. Unfortunately, typical drinking-water calcium concentrations around 0.5 mM present the worst case for fouling, whereas at concentrations above 3 mM no fouling was observed. As a consequence, Nghiem and Scha¨fer (2006b) suggested pretreatment to reduce the amount of particulate matter (similar to the approach taken by Friedler et al. (2008a, 2008b)) and/or chemical cleaning and backwashing to control fouling. Nghiem et al. (2006) showed that ordinary household bleach would probably be suitable for chemical control, which would be favorable for decentralized applications. We have found no recommendations of the simpler filter systems such as sand filtration as the sole treatment method for gray water. On the contrary, a number of authors discourage such applications due to their low efficiency with respect to removal of organic matter and microorganisms (e.g., Li et al., 2009). Chemical treatment. With respect to the chemical treatment systems, Li et al. (2009) reviewed work on coagulation, ion exchange, granular activated carbon, and photocatalytic oxidation. Not much work has been performed in this area, and it is therefore difficult to judge the potential of these technologies. It seems that coagulation and ion exchange are best suited for low-strength gray water (Lin et al., 2005; Pidou et al., 2008), whereas photocatalytic oxidation combined with microfiltration also delivers good results for higher-strength gray water (Rivero et al., 2006; Li et al., 2009). Like in the case of UF, however, solving the problems of membrane fouling is essential for achieving economic operation. Physical–chemical pre- and posttreatment. Sedimentation is a common pretreatment step for any gray water treatment. It occurs naturally in storage tanks if gray water is stored before treatment (see Section 4.07.2.5.2). Other common pretreatments are coagulation or filtration through sand or soil (see, e.g., Gross et al., 2007a, 2007b). The main object of pretreatment is a rapid removal of organic matter in order to reduce the load to a biological reactor or prevent fouling of membranes. Posttreatment is a polishing step intended to remove organics and microorganisms, which were not removed in the main treatment step – examples include sand filters (Friedler et al., 2006), solar photocatalytic oxidation (Gulyas et al., 2005), and RO (Friedler et al., 2008a, 2008b).
4.07.2.5.4 Biological treatment Biological treatment is highly favored for gray water treatment because it leads to stabilization of the organic material and thereby lowers the risk of microbial regrowth during storage. In a more centralized, nonbiological production plant of nonpotable water from mainly surface water, Oesterholt et al. (2007) reported on regrowth of Legionalla, which would of course be a serious health risk also in decentralized settings. For the effectiveness of biological treatment, the source of gray water is highly important. Degradability and nutrient availability depend primarily on the sources of gray water with kitchen gray water as the primary source of BOD (Friedler, 2004) and nutrients (Li et al., 2009). One of the main questions
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when evaluating biological degradability of gray water is the question whether the organic compounds are not a priori biodegradable, or whether nutrients limit degradability. Obviously, the nutrient requirements for degradation of organic matter will depend on sludge production and, consequently, on the solids retention time (SRT) in the biological reactor. The longer the SRT, the lower the sludge production and, as a result, the lower the nutrient requirements will be (see any standard textbook on biological treatment of wastewater). Equally, anaerobic treatment will require fewer nutrients than aerobic treatment, due to the lower yield coefficient of anaerobic organisms. Jefferson et al. (2001) cited literature values of required COD:N:P ratios of 100:20:1, 250:7:1, and 100:10:1 for aerobic treatment. For nutrient-deficient light gray water (from the bathroom), Jefferson et al. (2001) found that balancing the COD:N:P ratio certainly increased biodegradability, whereas the effect of adding micronutrients was more complex to evaluate. Krishnan et al. (2008) dealt with nutrient-deficient dark gray water (from kitchen and laundry; in a Malaysian setting) with a surprisingly high COD:N:P ratio of 100:1.82:0.76, and found a clear improvement of biodegradability after balancing the nutrients. The optimal COD:N:P ratio was determined to be 100:5:1 for an SRT of 13 days. Obviously, one has to be careful when comparing gray water from different cultural settings. Whereas European authors consider kitchen gray water to be rich in nutrients (see, e.g., Li et al., 2009), this may not be true in other cultures. Aerobic treatment. Aerobic treatment is generally considered the most promising technology for treatment of gray water for reuse, because it stabilizes the gray water with respect to organic matter. Typical reactors for decentralized treatment of gray water include biofilm reactors, sequencing batch reactors, and membrane bioreactors (MBRs). A special case is the recycled vertical flow bioreactor, a combination of a small wetland with a trickling filter, presented by Gross et al. (2007a, 2007b). In principle, all biological reactors can be used, provided it is possible to run them in a decentralized setting. It is no coincidence that we have found no examples in the literature of conventional activated sludge reactors, which are less suitable for decentralized settings. In older (Jefferson et al., 2000) as well as more recent reviews (Li et al., 2009), the MBR is consistently highlighted as the most successful technology for gray water treatment, not only as compared to other aerobic reactors, but also as compared to nonbiological treatments. The reasons are obvious: MBRs combine the advantages of biological treatment (stabilization of organic matter) with the advantages of membrane filtration (removal of suspended matter and microorganisms). The disadvantages are costs, energy consumption, and fouling. At a total water price of 1.46 USD m3, Friedler and Hadari (2006) showed that MBRs were only economically feasible in very large buildings (larger than 160 flats). In comparison to the more conventional technologies, however, there is still room for technical improvement of MBRs. It is, of course, also possible to add a filtration unit after a biological treatment, for example, a sand filter or a membrane. For all biological technologies, with the possible exception of the MBR, a subsequent disinfection step is necessary to render the water safe from a hygienic point of view (Jefferson et al., 2000).
Constructed wetlands are more natural systems, and will normally not be termed reactors. For water reuse, they offer some advantages. First of all, they are effective, relatively cheap, and have low environmental impact (Memon et al., 2007). Furthermore, for outdoor applications, there is virtually no risk of cross-connections with drinking-water pipes. However, they are space intensive, and thus in competition with other space-consuming activities. For areas with little available space, the green roof water recycling system has been developed (Frazer-Williams et al., 2008). Constructed wetlands are more advantageous in warmer climates, where biological activity can be maintained over the entire year. In very cold climates, they may be possible, but at the expense of very high HRTs. Gu¨nther (2000), for instance, described a constructed wetland in Sweden with an HRT of 1 year in order to bridge the winter, whereas Dallas et al. (2004) presented a case study of a reed bed in Costa Rica with an HRT of 7.9 days that achieved sufficient reduction of pathogens. Without getting into a detailed discussion of wetland technology for gray water treatment (that can be found elsewhere, see, e.g., Frazer-Williams, 2007), the latter example from Costa Rica presents an interesting concept for out-contracting of the maintenance of the reed bed and a very impressive cost reduction by replacing the expensive crushed rock carrier material with a waste product (used polyethylene terephthalate bottles). Anaerobic treatment. Anaerobic treatment of municipal wastewater is a relatively new area. Typically, one would consider this type of treatment only for very concentrated wastewater flows. However, as discussed in Section 4.07.2.1, in some countries with a very low per capita gray water production, gray water is a high-strength wastewater with concentrations of several thousand gCOD m3. Equally problematic, however, is the low degradability of many gray water organic compounds under anaerobic conditions, including some surfactants. In a recent review on gray water treatment, Li et al. (2009), therefore, concluded that anaerobic treatment is not suitable for gray water treatment. Leal et al. (2007) hypothesized that either inhibitory substances or a lack of trace elements may be responsible for poor COD removal from gray water under anaerobic conditions and suggested to look into the potential of a combined anaerobic and aerobic treatment. The problem of methane emissions from anaerobic treatment is discussed in Section 4.07.4.
4.07.2.5.5 Disinfection and removal of micropollutants Depending on the strength of the gray water, disinfection is normally only considered as a last step of treatment, because of the interference of the disinfection method with organic compounds. As described in Section 4.07.2.3.1, the risk from pathogens in gray water is high, and in many cases, disinfection is a necessary last step of a treatment sequence (see, e.g., Nolde, 2000). Disinfection. Chlorination with hypochlorite is a common method for disinfection of wastewater; besides, for reuse of gray water, chlorine is the most prevalent disinfectant (Winward et al., 2008b). Chlorination is generally considered effective and residual products prevent regrowth of microorganisms (March et al., 2005). In many cases, a certain
Source Separation and Decentralization
residual concentration of chlorine is required by regulation (Table 3). A main problem of the decentralized use of chlorine is the variable production and quality of gray water, leading to a variation in the chlorine demand (March et al., 2002). If too much hypochlorite is dosed, this is wasteful and leads to the typical odor of chlorinated water, whereas if too little is used, the disinfection purpose is not reached. Furthermore, many viruses and protozoans are resistant to chlorination, and in a decentralized setting it may be inconvenient for households to add the necessary chemicals (Fenner and Komvuschara, 2005). Chlorination also gives rise to toxic by-products, which are unwanted (March et al., 2004). The most critical issue of chlorination is the lack of effect on particle-associated microorganisms. Winward et al. (2008b) showed that such microorganisms were resistant to chlorination and thus concluded that removal of larger aggregates of microorganisms is essential when chlorination is used for disinfection of gray water. Ultraviolet (UV) disinfection would be a more elegant and potentially cheaper method of providing hygienic safety in decentralized settings, without any addition of chemicals. However, also for UV disinfection, shielding of microorganisms by particles occurs. Lack of disinfecting residuals may be considered a drawback of this technology, although Nolde (2005) found evidence that for properly treated gray water, this is not problematic. Fenner and Komvuschara (2005) developed and verified a theoretical model of UV disinfection of gray water and concluded that the practical limits of gray water disinfection (defined as a log 4 reduction of fecal coliforms) are found at a concentration of suspended solids above 60 mg l1 and a turbidity of 125 NTU. For biologically treated gray water with a turbidity of 1.5 NTU, Gilboa and Friedler (2008) found total removal of a number of pathogens at a UV dose of 69 mJ cm2, but they also found heterotrophic bacteria, which apparently were resistant even to high UV doses of 439 mJ cm2. These bacteria showed regrowth within 6 h of UV treatment. Winward et al. (2008d) showed that larger particles shield the microorganisms more from UV light than smaller particles and suggested the combination of biological treatment with filtration in order to improve the efficiency of UV disinfection. A well-accepted method of disinfection in households would be the use of plant essential oils, and an obvious application would be the use as a regrowth inhibitor after UV treatment. Although origanum oil has proved effective for this purpose in well-treated gray water (turbidity 2 NTU, total suspended solids o1 mg l1), the amount of oil required seems to make this approach too expensive (Winward et al., 2008c). Removal of micropollutants. Removal of micropollutants from gray water in decentralized settings is not a main focus of research. Obviously, the same methods as used for centralized treatment (e.g., chemical oxidation, activated carbon, and membrane processes; see Bolong et al. (2009) for a recent review) could also be used for gray water recycling. Micropollutants can be oxidized with chlorine, chlorine dioxide, ozone, and other chemical oxidants (Lee et al., 2008). Andersen et al. (2007) investigated the oxidation of parabens by chlorine dioxide in biologically treated gray water. Since already the biological treatment removed more than 97% of
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the parabens in the gray water, the effluent was spiked with 5 or 10 mg l1 of parabens. The result of the chemical oxidation was good (498% removal of parabens), but there is no discussion of interference with organic matter, should this technology be used in gray water that has not been biologically treated. A general discussion of the removal of micropollutants from wastewater by membrane processes is found, for example, in Kim et al. (2007). Scha¨fer et al. (2006) have looked specifically at retention of biphenyl A in the process of direct UF of gray water in a decentralized setting. It appears, however, that the major effect of biphenyl A retention in these experiments was adsorption onto the membrane and organic matter. The problem of resuspension of micropollutants during backwash of the membrane was looked into by Nghiem and Scha¨fer (2006a).
4.07.2.6 Summary The main purpose of decentralized gray water treatment is the reuse of water within the household or for irrigation. The most severe problem of gray water reuse in the household is linked to hygiene, whereas the problem of irrigation is mainly surfactants, which may render soils water repellent. Biological, especially MBR, technology is the gold standard of gray water treatment, because of the stabilizing effect on the treated water. In order to prevent bacterial regrowth and Legionella, chlorine is often used for disinfection. However, the right choice of treatment technology and disinfection method is largely disputed as discussed in this section. A main nontechnical issue is the lack of acceptance for gray water reuse, especially for close-to-body applications. The most-accepted indoor use of recycled gray water for toilet flushing may be out-competed by water-saving toilets, and more general reuse of gray water may be necessary in areas with severe water scarcity.
4.07.3 Urine In this chapter, we define urine very broadly, as fresh or stored urine, and with or without flush water. In domestic wastewater, around 80% of the nitrogen and 50% of the phosphorus stem from urine (DWA, 2008). Separating these nutrients efficiently at the source would result in a fairly balanced C:N:P ratio at the treatment plant and thus eliminate the need of advanced nutrient elimination (Larsen and Gujer, 1996; Wilsenach and van Loosdrecht, 2006). Furthermore, where treatment plants do not exist and nutrient removal is required for water pollution control (e.g., due to eutrophication of coastal areas), urine source separation would be the technology of choice (Larsen et al., 2007). Compared to a typical denitrifying treatment plant with a nitrogen-removing capacity of 50–60%, urine separation is attractive. At the same time, the nutrients in urine could be used beneficially in agriculture as a fertilizer. This is favorable especially in view of the limited phosphorus resources (see, e.g., Cordell et al., 2009), and where farmers consider commercial fertilizers too expensive (Medilanski et al., 2006).
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Urine also contains a good part of the micropollutants in wastewater stemming from the human metabolism (Lienert et al., 2007a), and separating and treating it at the source could be an efficient way of reducing the load of micropollutants to the aquatic environment (Larsen et al., 2004). Some urineseparating toilets also give rise to water saving and can therefore be seen as an alternative to gray water recycling for toilet flushing. Source separation of urine comes in two fundamentally different technical variations. The more important variation from a quantitative point of view is the traditional urineseparating dry toilet. A newer version of this type of toilet has been available for at least 30 years (Winblad, 1994), and drastically improves the management of dry feces. It is installed in large numbers in many areas without access to flush toilets, for instance, in China, where nearly 700 000 such toilets were installed by 2003 (Kva¨rnstro¨m et al., 2006; Figure 1). The European type of urine-separating toilet, often termed the NoMix toilet, was invented in Sweden in the early 1990s (Hellstro¨m and Johansson, 1999). It has been installed in very large numbers in pilot projects in Sweden, mainly motivated by the depletion of the phosphorus resources. The NoMix flush toilet (Figure 2) has also been used in most of the other European pilot projects and in recent years also in some Swedish municipalities, where the technology is subsidized for environmental reasons (Kva¨rnstro¨m et al., 2006). Obviously, urine is collected undiluted in the dry toilets, whereas it may be more or less diluted when collected in a flush toilet. Different NoMix flush toilets are available in the market. The main problems of these toilets are linked to clogging of pipes, as reported by Hellstro¨m and Johansson (1999) and Udert et al. (2003c). For a comprehensive discussion of the rationale for introducing urine source separation, the reader may refer to Larsen et al. (2001), Lienert and Larsen (2007), and Larsen et al. (2009).
Figure 1 A dry urine-separating toilet of the type that is often installed in China. & Edi Medilanski 2008, Eawag.
4.07.3.1 Production Rate and Composition of Urine Based on several years of international experience, new data for the production rate and composition of urine have recently become available (DWA, 2008; Vinnera˚s et al., 2006a). These are reported in Table 4 and compared to literature data for fresh urine compiled by Udert et al. (2006). As compared to the data for gray water, it is striking how close these values are. However, it should be noted that the data stem from European countries and, that in an international context, they are much more variable. For example, van Drecht et al. (2003) predicted a variation in human nitrogen excretion of a factor of 4, depending on diet.
4.07.3.2 Reuse Purpose and Regulation In contrast to gray water recycling, we will look at urine source separation not only from the point of view of (nutrient) reuse, but also from the point of view of treatment for water pollution control. Whereas treatment of gray water and conventional combined wastewater only differs very slightly, treatment of urine is closer to the treatment of concentrated streams such as supernatant from sludge treatment or industrial wastewater. To our knowledge, however, currently no regulations for this purpose exist. From a resource point of view, it would be of advantage to reuse the nutrients in urine. This is especially true for phosphorus, normally produced from phosphate rock, which is a limited resource. Many authors discuss the availability of phosphate (e.g., Driver et al., 1999; Zapata and Roy, 2004), and depending on the assumptions, known reserves will meet the requirements for the next 50–300 years. However, of more immediate concern may be the quality of the available phosphate rock, especially their content of cadmium (Driver et al., 1999; Smil, 2000; Isherwood, 2000). For nitrogen, the main
Figure 2 A typical flush urine-separating (NoMix) toilet. & Ruedi Keller 2008, Eawag.
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Table 4 Suggested values for typical urine production (volume, nitrogen and phosphorus) compared to literature data on fresh urine compiled by Udert et al. (2006)
Volume (l/cap/day) Organic matter (gCOD/cap/day) Nitrogen (gN/cap/day) Phosphorus (gP/cap/day) Potassium (gK/cap/day)
Udert et al. (2006)
DWA (2008)
Vinnera˚s et al. (2006a)
Values used in this work
1.3 13 12 0.93 2.8
1.4 10 10.4 1.0 2.5
1.5
1.4 10 11 1 2.7
concern is energy and price, since nitrogen is abundantly available in the atmosphere. Maurer et al. (2003) discussed the energy issues in detail, and showed that it is possible, but challenging, to compete with industrial ammonia production. The reuse of urine in agriculture is in many European countries basically forbidden, because it is not specifically allowed, for example in Austria (Starkl et al., 2007) and in Switzerland (Pronk et al., 2007). One exception is Sweden, where urine from many pilot projects or even from full-scale urine source separation in rural areas is applied to agricultural fields without previous treatment, but after storage of several months (Kva¨rnstro¨m et al., 2006). In Switzerland, a temporal permission to use treated urine for research purposes could be obtained, under the condition that the product would be hygienically safe and free from micropollutants (Pronk et al., 2007).
4.07.3.3 Risks Associated with Source Separation of Urine Risks of urine source separation are normally considered only when urine is used as a fertilizer in agriculture. As for gray water reuse, the risks are associated with hygiene and other organic and inorganic contaminants in urine. However, for the handling of source-separated urine in the household, other issues may be of importance. Due to the risk of explosion, for example, the production of dry ammonium nitrate calls for a proper risk analysis before being introduced at the household level, and since ammonia is toxic, proper ventilation is essential (Udert et al., 2006).
4.07.3.3.1 Hygienic risks Normally, urine is considered sterile, when it leaves the human body, but certain diseases do lead to excretion of pathogenic organisms (Santos et al., 2004; Vanchiere et al., 2005) and prions (Reichl et al., 2002) via urine. Of more practical importance, however, is the fecal contamination occurring at the toilet. In one system of urine source separation, Scho¨nning et al. (2002) found that urine was contaminated with approximately 9 mg feces per liter. A detailed risk assessment is found in Ho¨glund et al. (2002b), concluding that virus must be considered the most critical parameter. It is very difficult to estimate typical concentrations of microorganisms in source-separated urine, since, for instance, the typical indicator organism Escherichia coli rapidly dies of in a urine storage tank (Ho¨glund et al., 2000). From the limited information that is available in literature, Enterococci seem to be present at the highest levels, reaching in one case a level of 105 ml1 at the bottom of the urine storage tank (where particles and microorganisms accumulate; Ho¨glund et al., 2000).
11 1.0 2.7
4.07.3.3.2 The risks of organic contaminants and salts As compared to gray water, urine is much less problematic in agriculture. Due to the proximity to animal urine, it must be taken for granted that urine will normally not cause any damage to soil structure. However, it is unclear whether salt could be a problem under very dry conditions (Lienert and Larsen, 2007). It is generally recognized that urine contains a significant amount of excreted micropollutants. Based on an extensive literature study, Lienert et al. (2007a) found that on average two-thirds of the micropollutants from the human metabolism are excreted via urine, and one-third via feces. However, due to the more lipophilic character of the compounds excreted in feces, Lienert et al. (2007b) estimated that the ecotoxic potential from these two excretion pathways is about equal. In addition, Winker et al. (2008) concluded on the basis of a literature review that the excretion of micropollutants via urine is relevant. Since pharmaceuticals are produced with the explicit purpose of biological action, source control is generally difficult.
4.07.3.4 Public Perception of Urine Source Separation It is frequently argued that people will never accept urine source separation, because they have no personal interest in changing a well-functioning system. However, many research results show the opposite: most people actually find the idea interesting and support it, provided there would be environmental advantages of changing the system. Currently available NoMix toilets are acceptable to most users if they are not responsible for maintenance themselves (e.g., in public buildings), and many people can even envisage living in a home equipped with NoMix toilets (Lienert and Larsen, 2006). In a comparison of pilot projects on urine separation in different European countries, urine source separation appeared to be generally well accepted, although blockages, smell, or more time-consuming cleaning were unpleasant for the users (Lienert and Larsen, 2010). Based on interviews in a housing estate in Austria equipped with NoMix toilets, Starkl et al. (2007) reported that about half of the tenants would prefer to change back to a normal toilet again, due to these problems. However, in view of the present state of technology, it is more surprising that half of the tenants would actually prefer to keep these toilets. To be widely accepted, it is obvious that NoMix toilets must evolve to approximately the same comfort standard as experienced with other modern toilets. In addition, the use of a urine-based fertilizer in agriculture is generally seen positively by the public (Pahl-Wostl et al., 2003; Lienert and Larsen, 2010), as well as by farmers (Lienert et al., 2003). However, for the recycling to agriculture, the risk
214
Source Separation and Decentralization
aspect dominates and, for the farmers, obviously convenience and costs are equally important.
4.07.3.5 Treatment Technologies for Urine As compared to gray water, the goals of urine treatment are more diverse, and, for reuse, nutrients and not water are of interest. Like in the case of gray water, stabilization, hygienic quality, and removal of any undesirable organic matter are important, and also the concentration of one or more of the important nutrients in a smaller volume of water or as dry matter could be an important goal (Maurer et al., 2006). Moreover, nutrient removal for water pollution control can be a goal if no reuse is possible or desired.
adding a urease inhibitor. To our knowledge, there has been no attempt to keep urine sterile, but urease has been successfully inhibited. Specific urease inhibitors are available, but they have low efficiencies and negative side effects (Benini et al., 1999). Immediate addition of a strong acid, however, can keep pH below 4 for more than 250 days by inhibiting urea hydrolysis and at the same time improving the inactivation of pathogens (Hellstro¨m et al., 1999). Reducing pH after hydrolysis of urea can also be achieved by acid, but about 4 times more acid would be required. An alternative would be partial biological nitrification, discussed in Section 4.07.3.5.4.
4.07.3.5.3 Physical–chemical treatment 4.07.3.5.1 No treatment The traditional use of source-separated urine is the direct use of fresh urine in agriculture, and, even today, this is common practice in some areas (Kva¨rnstro¨m et al., 2006). From the point of view of hygiene, this is only advisable if special care is taken to avoid contamination of eatable products. Much more common, however, is the infiltration of source-separated urine from dry toilets (Anonymous, 2005), which obviously is not a good idea from the point of view of environmental protection. The motivation for this practice is the improved stabilization of feces that is obtained in the urine-separating dry toilets.
4.07.3.5.2 Storage The effect of storage on the chemical composition of urine has been discussed in detail by Udert et al. (2006). Storage is the only process that has been tested in depth for its ability to reduce potential health risks from microbial contamination of urine. A large number of experiments on the inactivation rates of selected bacteria and virus in source-separated urine were conducted by Ho¨glund et al. (1998, 1999, 2000, 2002a, 2002b). It was shown that three parameters determine the success of the process: temperature, storage time, and pH. Temperature was the most important parameter: whereas storage for 35 days at 20 1C and a pH larger than 9 removed 90% of the activity of rhesus rotavirus, no significant decrease in activity was observed at 4 1C. Based on this broad experience, Ho¨glund et al. (2002b) concluded that a storage time of at least 6 months at 20 1C without pH control would be sufficient to produce a safe fertilizer. Storage of urine, however, may also have negative effects. Precipitation of phosphorus will result in the accumulation of a nutrient-rich sludge at the bottom of the storage tank (Udert et al., 2003c), which is at the same time the most critical part of urine from a hygienic point of view (Ho¨glund et al., 2000). Perhaps more seriously, due to the high pH value in the tank, ammonia may evaporate from tanks through ventilation (Udert et al., 2006; Rossi et al., 2009). Stabilizing urine for storage would therefore be of advantage. Since precipitation of phosphorus and evaporation of ammonia are both due to a high pH, lowering the pH in the urine solution is the most important measure suggested. There are basically two possibilities: either preventing a pH increase through inhibition of urease activity or subsequently reducing pH. Preventing a pH increase would be possible by keeping the solution sterile (e.g., by applying a membrane process) or by
Apart from just spreading urine as a liquid on the fields, the simplest imaginable recovery method for the nutrients in urine would be concentration by evaporation. However, evaporation of urine is limited by two main factors: ammonia volatilization and energy consumption. Loss of ammonia only occurs at a high pH and the same pH-reducing strategies as for storage could be adopted (see Section 4.07.3.5.2). Energy recovery or energy cascading, for example, recovering the energy in the vapor for heating warm water in the household, can reduce energy consumption. Energy recovery in decentralized settings is of course a challenge. However, a small-scale vapor compression distillation system developed for space applications operates with 85% energy recovery (Wieland, 1994). An alternative would be the evaporation from nonhydrolyzed urine. Mayer (2002), cited in Maurer et al., (2006), evaporated nonhydrolyzed urine at 78 1C and 200 mbar, producing a viscous solution containing about 10% nitrogen. Moreover, freezing could be used for producing a concentrated solution. Repeated freezing of urine at –14 1C resulted in the concentration of 80% of the nutrients in 25% of the original volume (Lind et al., 2001), a result that was generally confirmed by Gulyas et al. (2004). However, the energy consumption of the chosen technology was relatively high. Volume reduction can, in principle, also be obtained by RO. The efficiency of RO for ammonia retention depends heavily on pH because the retention of the charged molecules (NHþ 4 ) is much better than the retention of the uncharged NH3. Dalhammar (1997, Behandling och koncentrering av humanurin, Royal Institute of Technology, Stockholm, Department of Biochemistry and Biochemical Technology, report, personal communication, cited in Maurer et al. (2006)) performed experiments with hydrolyzed, but acidified, urine at a pH of 7.1. At a concentration factor of 5 (at a pressure of 50 bar), around 70% of all three main nutrients (N, P, and K) were recovered in the concentrate. Thorneby et al. (1999) obtained similar results at a pressure of 30 bar with liquid manure, but with a much higher nutrient recovery (490%). In decentralized settings, problems of energy recovery and fouling and/or scaling must be expected (Maurer et al., 2006). The most widely applied physical–chemical treatment of urine is the precipitation of phosphorus. This is an attractive technology for decentralized recovery of phosphorus because the residuals produced in the reaction are small. Consequently, the precipitation product can be easily collected, for instance, with the normal solid waste from households.
Source Separation and Decentralization
Furthermore, technologies of phosphorus precipitation are known from conventional wastewater treatment, especially from the treatment of concentrated liquids, for example digester supernatant. For an overview of existing technologies, see, for example, Wilsenach and Van Loosdrecht (2002) and De-Bashan and Bashan (2004). Udert et al. (2003c) investigated the naturally occurring precipitation in urine conducting pipes and found mainly struvite (magnesium ammonium phosphate, MAP), hydroxyapatite (a calcium phosphate), and calcite (a carbonate mineral). For technical recovery of phosphorus from urine, we have only found examples of struvite precipitation. Struvite is an attractive precipitate because it is rapidly formed on the addition of magnesium (Ronteltap et al., 2007a), for example as magnesium oxide, magnesium hydroxide, or magnesium chloride. As an additional advantage, struvite is a good slow-release fertilizer (Johnston and Richards, 2003). Models for the thermodynamic equilibrium and the kinetics of phosphorus precipitation in urine have been set up by Udert et al. (2003b, 2003d). An overview of solubility products for struvite is presented by Ronteltap et al. (2003) and a simplified solubility product determined by Ronteltap et al. (2007a). It would be equally attractive if nitrogen could also be precipitated to form a solid product. Unfortunately, there are only two possible nitrogen precipitates, and there are problems with both. One is the commercially available slowrelease fertilizer isobutylaldehyde-diurea (IBDU), a complex of urea, and isobutyraldehyde (IBU), which is however not very suitable for the recovery of nitrogen from urine. Behrendt et al. (2002) have shown experimentally that high urea concentrations and excess IBU are necessary. Applying 5 times more IBU than should be necessary from the stoichiometry of the process only resulted in the precipitation of 75% of the urea as IBDU. The other possibility is the precipitation of nitrogen as struvite, at the cost of adding about 25 times more phosphorus than available from urine. The process is perfectly possible (Tu¨rker and C¸elen, 2007), but only sustainable on a large scale if phosphorus is recovered from the precipitate. Note that the degree of recovery must be very high if the process should not be wasteful in phosphorus. In niche applications, where there is a market for the struvite produced, this technology may be applied. For the recovery of nitrogen, a number of other physical– chemical processes are available, a promising one being ammonia stripping. This process is well known from the treatment of digester supernatant, but energy consumption is relatively high (7 kWh m3 treated liquid; Siegrist, 1996). Ammonia can be adsorbed either in water under pressure (resulting in a 10% ammonia solution at a pressure of 5 bar; Behrendt et al., 2002) or in sulfuric acid (also resulting in a 10% ammonium sulfate solution; Siegrist, 1996). As for struvite precipitation, the advantage of urine over digester supernatant for this process is the naturally high pH of urine, avoiding the large dosage of chemicals for pH regulation. Ion exchange is often suggested as a possible method of producing an attractive fertilizer product from urine. Naturally occurring zeolites have high affinities for ammonium and have been used for recovery of ammonia from urine (Lind et al., 2000; Ba´n and Dave, 2004). Combined with struvite precipitation, Ba´n and Dave (2004) obtained a relatively high
215
recovery rate dosing 15 g l1 of zeolite, however, at a very high effluent concentration of 1000 gN m3. Furthermore, it has not been possible to reproduce these results with such low concentrations of zeolite and, with more typical concentrations, the process becomes unattractive due to the large amount of residuals produced.
4.07.3.5.4 Biological treatment For removal of nitrogen from wastewater, the efficiency of biological treatment is well proven. In concentrated solutions, however, it is still a matter of dispute whether biological removal or physical–chemical recovery would be the best solution (see Maurer et al. (2003) for a discussion of the energy issues of nitrogen recovery). Biological recovery is possible in the form of stabilization of ammonia, with or without physical–chemical volume reduction. Biological recovery of phosphate with the Bio-P process known from conventional wastewater treatment plants has, to our knowledge, never been attempted from concentrated solutions. Nitrification is very suitable for decreasing pH of urine, thereby achieving a stabilization of ammonia. Stabilization is relevant for storage (Section 4.07.3.5.2), for transport and spreading, and as a pretreatment for evaporation (Section 4.07.3.5.3). Nitrification, either to nitrite or to nitrate, produces acid that will eventually stop the biological process when a pH around 6 is achieved. Based on the chemistry of urine, one would expect a partial nitrification of urine to lead to a ratio of ammonia to nitrite or nitrate of approximately 1:1. Experimentally, this was confirmed by Johansson and Hellstro¨m (1999) and Udert et al. (2003a). The latter authors ran different reactor systems, leading to either ammonium nitrate or ammonium nitrite. Nitrite is toxic, but at low pH, chemical oxidation with oxygen will easily convert it to nitrate (Udert et al., 2005). A welcome side effect of biological treatment is the removal of the typical urine odor and more than 80% of COD (Udert et al., 2008). For removal of nitrogen, either heterotrophic or autotrophic biological denitrification can be applied. For heterotrophic denitrification, the process used for the treatment of combined wastewater, an organic substrate is needed to reduce nitrate or nitrite to nitrogen gas. Autotrophic denitrification, or the anammox process (Strous et al., 1998), is applied, for instance, for the treatment of supernatant from sludge dewatering. In this process, ammonia is oxidized under anaerobic conditions with nitrite by autotrophic microorganisms. Udert et al. (2003a) showed that the ammonium nitrite solution produced from urine is suitable for autotrophic denitrification, and Udert et al. (2008) found that ammonia removal from urine by partial nitrification and a combination of heterotrophic and autotrophic denitrification is possible in a single reactor with a removal efficiency of approximately 80%. However, urine is a complex solution (see Udert et al., 2006) with high concentrations of ammonia and salt, and research is still necessary before stable biological nitrogenconverting processes can be run in a decentralized setting. Inhibition of nitrifiers is still not completely understood and the process is sensitive especially to auto-inhibition. Larsen et al. (2009) suggested that the use of genetic methods for
216
Source Separation and Decentralization
identification of microorganisms may be a promising route for developing stable biological treatment of source-separated waste streams such as urine in decentralized settings.
4.07.3.5.5 Hygienization and removal of micropollutants Hygienic parameters in urine are primarily of importance for reuse purposes, whereas micropollutants may be of importance in the receiving waters and for reuse. Technically, there are many ways of disinfection or sterilization of any solution (heat, pressure, UV, etc.), but none of these methods have been tested for urine. Only storage (see Section 4.07.3.5.2) has been explicitly investigated for its ability to reduce the amount of pathogens in source-separated urine, although many of the treatment steps discussed in this chapter are expected to have an influence on its hygienic properties (see Maurer et al. (2006) for an overview). When a fertilizer based on urine is produced, the separation of nutrients and micropollutants is important, whereas elimination of micropollutants is relevant for water-pollution control. Separation processes are normally based on membranes or precipitation, whereas removal processes rely on oxidation or adsorption (Larsen et al., 2004). Fortunately, struvite precipitation is an efficient way of separating phosphorus and micropollutants (Ronteltap et al., 2007b). For an overview of the efficiency of different urine treatment methods with respect to the removal of micropollutants, the reader is referred to Escher et al. (2006). The same separation methods have a certain influence on hygienic parameters; however, as stated above, we have found no systematic research on this topic. In contrast to gray water, where RO can be expected to deliver water of the highest quality, also with respect to microorganisms and micropollutants, this is obviously not true when the goal is to produce a safe fertilizer product from urine. For this purpose, micropollutants and microorganisms must take a different route than the nutrients. Other membrane processes have been tested for their ability to separate nutrients and micropollutants. Electrodialysis is based on ionexchange membranes, with an apparent pore size of approximately 200 Da (Kim et al., 2003), which could be expected to retain micropollutants. Pronk et al. (2006a) showed that a separation of nutrients and micropollutants is in fact possible. The efficiency of nutrient recovery was 90%, with a 90% separation of micropollutants, and the concentration factor was around 3. In order to prevent a pH increase in the concentrate containing the nutrients, Pronk et al. (2006b) tested the use of bipolar membranes, which were effective in batch experiments. However, for real-life applications, improvements of the system are necessary (discussed in detail by Pronk et al. (2006b)). Nanofiltration is a well-known possibility of retaining different micropollutants (see, e.g., Kimura et al., 2004; Nghiem et al., 2004). Pronk et al. (2006c) have shown that it is possible to separate nutrients and micropollutants in source-separated urine with microfiltration, provided urea is not hydrolyzed. At optimal conditions, more than 90% removal of micropollutants was observed. Obviously, micropollutants can also be chemically oxidized in urine. However, organic matter and ammonia will compete with the oxidation of micropollutants. Pronk et al.
(2007) tested the use of ozone for the oxidation of micropollutants in untreated urine and found between 80% and complete removal of different compounds at an ozone dose of around 1 g l1. Although this seems high as compared to conventional wastewater treatment, urine is of course produced at a much lower rate and contains a substantial amount of micropollutants (see Section 4.07.3.1). In addition to the processes presented here, it is possible, in principle, to remove micropollutants by adsorption on active carbon or other adsorbents. It can be expected that the presence of high amounts of COD in urine strongly interferes with the adsorption process. However, a biological process can remove more than 80% of the organic matter (Udert et al., 2008), which will drastically improve the efficiency not only of oxidation, but also of adsorption processes.
4.07.3.6 Summary Urine source separation is publicly well accepted, given that toilet comfort is not compromised. Due to the dominating role of urine with respect to nutrients, the technology has a high potential for efficient nutrient recovery and removal. However, there is still much scope for further technical development before urine source separation can compete with existing wastewater treatment technologies. Socioeconomic issues have not been discussed in this chapter, but play a major role, for the successful introduction in industrialized and fast-industrializing countries. In dry sanitation systems in developing countries, urine source separation is already standard technology, but there is a lack of suitable technologies for further processing of urine.
4.07.4 Feces In this chapter, we consider feces with or without urine, and with or without flush water. Diluted combined toilet waste (feces þ urine) is known as black water, whereas diluted feces alone (without urine) are termed brown water. Dry as well as diluted feces may contain toilet paper. From the point of view of urban water management, feces compose the single most important source of pathogens, and it is difficult to find a satisfying solution to this problem. Sewer-based wastewater management effectively removes feces from the immediate urban environment, but the actual removal of microorganisms depends on how advanced the treatment plants and how leak-proof the sewers are. Really effective removal is only achieved with membrane reactors, with disinfection of the effluent, and with perfect sanitary sewers. At the same time, and for obvious reasons, this fraction seems to be the most critical one to handle in a decentralized setting. However, doing so in a safe way can result in a dramatic improvement of urban hygiene and is thus an urgent matter where sewers are not appropriate. Feces, however, also contain important nutrients and where agricultural production is limited by the availability of fertilizer, it seems worthwhile to reuse feces – or alternatively the nutrients contained in feces – as an important resource. It should be noted that the main contribution of feces to agriculture is phosphorus, whereas nitrogen and potassium are
Source Separation and Decentralization
mainly contained in urine. The importance of the organic content of feces for soil improvement is disputed, but according to Jo¨nsson et al. (2004), the improvement is hard to distinguish if feces are added based on the phosphorus requirements of the soil.
4.07.4.1 Production Rate and Composition of Feces Until recently, the production rate and composition of feces were more relevant to medical doctors than to wastewater professionals. However, this has changed during the last decade and some of the data reported in Table 5 are based on extensive experimental research in Sweden (Vinnera˚s et al., 2006a) and Germany (DWA, 2008).
4.07.4.2 Reuse Purposes and Regulation Reuse of feces is relevant for agriculture and aquaculture. Like for sewage sludge, reuse is relevant not only from a resource point of view, but also as a disposal option. In densely populated areas, the only sustainable options for feces are reuse or total oxidation (e.g., incineration), either directly or via sludge production in a treatment plant. Nutrient recycling and total oxidation may be combined, for example, by using ashes from incinerated feces as a fertilizer – an option that would allow recycling of phosphorus without hygienic risks (Scho¨nning and Stenstro¨m, 2004). In areas where reuse of nutrients is essential, recycling of feces is desirable, under the condition that hygienic risks are minimized. WHO (2006) has set up new guidelines for this approach, replacing the older guidelines from 1973 and 1989. Although these guidelines are only recommendations and regulations only available nationally, we here present the recommendation of the World Health Organization (WHO) guidelines as one example of how regulation could look (Table 6). Since monitoring is difficult, regulation based on best technical practice may be the more successful approach.
217
organic and inorganic pollutants, these aspects are normally overshadowed by the higher risks of pathogens.
4.07.4.3.1 Hygienic risks Feces contain around 1011–1013 cells g1 and are the main source of microorganisms in wastewater. These microorganisms are however only critical when the individual excreting them is infected. In a comprehensive review, Scho¨nning and Stenstro¨m (2004) discussed the prevalence of different critical bacteria, virus, protozoan, and helmints in different environments. They generally define a conservative approach to risk analysis, where the most resistant organism is chosen as indicator organism. It is not possible to define one specific indicator organisms that will always be appropriate, but some guidance with respect to the choice in specific situations will be found in the review cited above. The main conclusion to bear in mind is that there is a major difference between detecting fecal contamination (which may be done by looking for fecal coliforms) and controlling whether fecal matter has been treated to a hygienically safe level.
4.07.4.3.2 The risks of organic contaminants There is not much work available on organic contaminants in feces. In a literature study, Lienert et al. (2007a) found that about one-third of the pharmaceuticals and hormones excreted by humans end up in feces, whereas the rest is excreted via urine. However, since the compounds excreted via feces tend to be more problematic than those excreted via urine, there is an estimated 50–50 distribution of the ecotoxicological risk potential between urine and feces (Lienert et al., 2007b). Since pharmaceuticals are produced with the explicit Table 6
WHO guidelines for reuse of feces in agriculture
Verification monitoring Recommendation for storage
o1 g1 total solids of Helminth eggs o1000 g1 total solids of E. coli 2–20 1C: 1.5–2 years 420–35 1C: 41 year pH49: 6 months
4.07.4.3 The Risks of Source Separation of Feces Feces are the main source of hygienic risks from wastewater and any discussion on risk connected to feces is thus dominated by the hygienic aspect. Although feces also contain
Table 5
Suggested values for typical production of feces Ciba-Geigy (1977)
Wet mass (kg/cap/day) COD (gCOD/cap/day) Dry mass (g/cap/day) Nitrogen (gN/cap/day) Phosphorus (gP/cap/day) Potassium (gK/cap/day) Toilet paper (g/cap/day) a
Data compiled from WHO (2006) Excreta and greywater use in agriculture. In: WHO Guidelines for the Safe Use of Wastewater, Excreta and Greywater, vol. 4, ISBN 92 4 154685 9. http://www.who.int/water_sanitation_health/wastewater/gsuww/en/ (accessed March 2010).
Henze and Ledin (2001)
a
0.1–0.23
Vinnera˚s et al. (2006a) 0.14
60 21; 34b 1.1 0.55 1.1
30 1.5 0.50 1.0 8.5
DWA (2008)
Values used in this work
0.14 60
0.14 60 30 1.5 0.5 0.9 8.5
1.5 0.5 0.7
Mean values for adults for different diets; the lower value corresponds with a ‘European’ diet, while the higher value corresponds with a diet ‘rich in fibers’. Mean values from two different samples of adults (7 and 24 people). Since the relative density of feces (compared to water) is close to 1, data on volume are reported as wet mass. b
218
Source Separation and Decentralization
purpose of biological action, source control is generally difficult.
4.07.4.4 Decision Making and Public Perception of Source Separation of Feces Since many diseases are transmitted via feces-to-mouth contact, there are good reasons why feces are considered repulsive. However, as described by Avvannavar and Mani (2008), the societal context will determine the ‘‘ystrict and unwritten rules and taboo of how to behave when excreting.’’ These taboos also lead to the fact that hardly any scientific literature on the acceptance of different types of feces handling is available, despite the high importance of this issue for urban hygiene and public health. For a detailed overview of societal attitudes to the handling of feces and urine in different contexts, the reader is referred to Avvannavar and Mani (2008). In the literature, we have found only very few practical acceptance studies connected to feces, and all these focus on the acceptance of dry sanitation. In a study on large-scale dry sanitation in an urban area in Mexico, Cordova and Knuth (2005) found that when the toilets were well functioning (which was the case in four out of five sites), users were very satisfied. Similarly, our own experience with pilot projects in China indicates that the proper technical implementation is essential for the success of a pilot project and the acceptance of the technology (Medilanski et al., 2007). It is generally known that it is difficult to assess questions of technology acceptance if the respondents have no access to the technology in question. However, in a study on rural wastewater management in Austria (Starkl et al., 2007), such hypothetical options were discussed in focus groups. In this study, focus-group participants strongly opposed compost toilets because they anticipated a higher demand for space and maintenance, and feared that the toilets would smell. Experience from a rural area in Sweden (Kva¨rnstro¨m et al., 2006) indicates that the Swedish rural population is much more open toward dry sanitation systems. For environmental reasons, about 200 urine sourceseparating flush and dry systems were installed in a rural municipality and user satisfaction was high. Interestingly, in a follow-up investigation, 87% of the users with a dry system said that they would have installed this system even without the 50% subsidy that they received from the municipality, whereas only 23% of the people with a urine-separating flush toilet said the same. Experience from pilot projects indicate that the high acceptance of dry urine-separating toilets is connected with the fact that these toilets are completely odorless, because they are much better ventilated than conventional and urine-separating flush toilets (Kva¨rnstro¨m et al., 2006). Besides, note that conventional bathrooms are heavily ventilated (Tung et al., 2010), and that the ventilation of dry toilets will not automatically have negative effects on energy consumption. All studies emphasize that successful technology implementation and comfort for the user are the most important issues for acceptance. This corresponds with the results found for urine source-separating toilets (see Section 4.07.3.4): if source separation makes sense from an economic and/or ecological point of view, people are generally ready to accept such technologies, but they do require a level of bathroom
comfort, which is at least comparable to the level that they are used to. It is interesting to realize that people often associate dry toilets with increased problems of smell, whereas practical experience shows that in fact the opposite may be the case.
4.07.4.5 Treatment Technology for Feces The main goal of decentralized feces treatment is to improve the hygienic and esthetic quality (including also the removal of odor), and, in some cases, also to gain energy and reduce the amount of material to be transported. Hygiene is characterized by pathogen die-off, which is considered a first-order reaction (at least during shorter periods), where the organismspecific rate constant is considered a function of temperature, pH, and water activity (Scho¨nning and Stenstro¨m, 2004). Esthetic qualities are related to a change in physical appearance, where color and odor are significantly altered (Avvannavar and Mani, 2008). The main decisions with respect to the options of practical treatment technology for feces are taken very early in the system, namely at the point of the toilet. The most important distinctions are between dry and flush toilets and between toilets with and without urine source separation. However, Table 7 toilet type
Concentration of organic matter in feces as a function of
Flush volume (l/flush)
Concentration (gCOD l1)
Comments
Typical older American toilet Typical older European toilet Typical modern European toilet Possible development of modern toilet NoMix toilet, type Dubletten/ Gustavsberg Possible development of Dubletten/Gustavsberg Existing vacuum toilets Possible development of vacuum toilet NoMix vacuum toilet Possible development of NoMix vacuum toilet
Feces related
Urine only
13 6 6
13 6 3
0.92 2.0 3.3
4
2
5.0
6
0.2
8.8
4
0.2
13
1 0.5
1 0.5
12 24
1 0.5
0.2 0.1
33 67
Dry toilets With urine Without urine (100% urine separation)
% dry matter 6 23
Assumptions (Ciba-Geigy, 1977): 140 g feces/cap/ day with a dry matter content of 23% and 1.5 l urine/cap/day (dry matter B65 g/cap/day). Additives and toilet paper are not considered
Calculations based on 5 flushes/cap/day, one of those being feces related (Friedler et al., 1996), and a feces production of 60 gCOD/cap/day (Table 4). 100% compliance of the toilet user with respect to flushing behavior is assumed. Toilet paper is not accounted for.
Source Separation and Decentralization
with respect to the concentration of organic matter from flush toilets, there is a gradient from high-flush toilets without urine separation, resulting in highly diluted feces, to very waterefficient devices, where organic matter ends up in a very concentrated form. In Table 7, we give some examples of the resulting concentration of organic matter in separately collected feces depending on the type of toilet. Particle separation after flushing, for example with a whirlpool surface tension separator (Vinnera˚s, 2004), can result in very high feces concentrations similar to the ones found in 1-l vacuum toilets. For toilets with particle separation as well as for watersaving NoMix toilets, the mode of dealing with toilet paper is essential for the resulting concentration of feces dry matter. In both cases, flushing away toilet paper used in the urine-only mode (terminology from Friedler et al. (1996)) will lead to considerably lower concentrations of dry matter. For all other toilets, the addition of toilet paper and/or additives will increase the dry matter concentration. As seen from Table 7, the concentration of COD in separately collected feces from flush toilets varies from typical wastewater (1000 mgCOD l1) to very concentrated activated sludge (up to nearly 7% COD). The dry matter content of feces from dry toilets varies from 6% (without urine separation; more than half of the dry matter is made up by salt from urine) to 23% (with 100% urine separation), without accounting for toilet paper and/or additives.
4.07.4.5.1 No treatment The no-treatment option for source-separated feces is frequent in many developing countries, either in the form of open defecation (Avvannavar and Mani, 2008) or in the form of disposal of toilet waste from on-site sanitation in agriculture, aquaculture, receiving waters, or just simply in the urban or peri-urban environment (Ingallinella et al., 2002). It is needless to discuss that in more densely populated areas, this option is unacceptable for the purpose of urban hygiene, but it should be kept in mind as a base scenario when evaluating the alternatives presented in the following.
4.07.4.5.2 Storage In Section 4.07.3, we introduced storage as a treatment method, because this is a relevant option for urine. However, for feces, there are two distinct types of storage: those which are only intended to keep feces contained, until they can be transported to a central treatment (most often in the form of fecal sludge (FS), see Section 4.07.4.5.1), and those which are intended to stabilize feces. Here, we only discuss the latter type of storage. For feces, it is not very clear where storage ends and physical–chemical treatment begins. When feces are stored, most often at least some drying will take place, and with the addition of conditioners, for example ash for increasing the pH value of the fecal matter, also hygienic issues are pursued. However, pure storage would have a certain effect on the microbial quality, as seen for urine (Section 4.07.3.5.2); but with the higher content of pathogens and the lower pH values, the required storage time is considerably higher (Scho¨nning and Stenstro¨m, 2004). Obviously, there are many problems of pure storage of feces, mainly caused by smell and flies, and,
219
where possible, some type of conditioning is preferred (see Section 4.07.4.5.3).
4.07.4.5.3 Physical–chemical treatment Solid–liquid separation and drying. For FS, the feces–liquid mix arising from decentralized storing of toilet waste, solid–liquid separation is obtained in settling–thickening tanks or on sludge drying beds (planted or unplanted; Strauss and Montangero, 2002). Whereas the settling–thickening tank is a pretreatment, which should be followed by further drying or biological treatment, the drying bed combines percolation of water with drying to produce a product that can be used in agriculture. Ingallinella et al. (2002) reported good results of a planted drying bed, with a count of live helminth eggs of 2 g1 TS, which is in accordance with the older WHO recommendations, and only a factor of 2 higher than the 2006 recommendations referred to in Section 4.07.4.2. Although the percolate from drying beds is of better quality than from the settling tanks, it still requires treatment (Strauss and Montangero, 2002). In dry toilets, feces are in many cases further dried by the addition of structural material such as saw dust, not only decreasing the relative humidity by adding more dry mass, but also allowing for better evaporation of water. Ventilation can be passive or forced. Decreasing humidity helps kill pathogens (Scho¨nning and Stenstro¨m, 2004) and reduces nuisance such as smell and flies. A special type of drying takes place in a precomposting tank, often described by the German word Rottebehaelter (Gajurel et al., 2007, 2004, 2003), which consists of a simple filter bag, where feces from a water-flushed toilet are collected and drained. This is a further development of the septic tank, which is widespread also in rural areas of industrialized countries (Section 4.07.5.3.1). According to the articles cited above, this filter bag provides an (outdoor) odor-free storage and drying of fecal matter obtained from flush toilets. After 3 months of storage, a product is obtained with a suitable water content for the sludge to be treated by vermicomposting (Gajurel et al., 2007; see Section 4.07.4.5.4, for a discussion of vermicomposting). Additives. A classical way of stabilizing feces is the addition of ash, which will increase pH and decrease humidity, and thereby increase the rate of pathogen die-off (Scho¨nning and Stenstro¨m, 2004). At the same time, the covering of feces with ash improves the esthetics in a dry toilet. A newer technology is the addition of either urea or ammonia, which has proved to be effective for all types of fecal pathogens. For details on dosage, temperature, and storage requirement, the reader is referred to Nordin et al. (2009), Ottoson et al. (2008), and Vinnera˚s et al. (2003). Physical–chemical oxidation. We have found no recent technical papers on physical–chemical oxidation of feces, but Scho¨nning and Stenstro¨m (2004) suggest that incineration would be the optimal treatment method for this type of bio waste, eliminating all hygienic risks and allowing for the recycling of phosphorus from the ashes. An on-site approach is suggested (i.e., incineration directly in the toilet) in order to avoid hygienic risks associated with transport. These same authors also note that heating feces sufficiently would have the
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same result on pathogen elimination – an option, which may be easier to realize in practice and which is already under development in on-site settings for thermophilic composting (see Section 4.07.4.5.4). For closed life-support systems, as they may be considered for long-term space missions, physical–chemical oxidation of human waste, including feces, has been considered (Upadhye et al., 1993), but we have found no literature on actual experiments. Upadhye et al. (1993) calculated a heat of combustion of about 22 kJ g1 dry mass of feces (lower heating value at 25 1C), which in principle would be available to evaporate the water content of feces. Based on an energy requirement for evaporating water of approximately 2.6 kJ g1, feces with up to 88% water content (see also Table 7) can be burned energy neutral. It is obviously attractive to imagine a toilet with internal combustion of feces, where the resulting ash contains many of the nutrients contained in feces (including phosphorus), presenting the minimum amount of residues possible (B7 g/cap/day; Ciba-Geigy, 1977), and eliminating effectively all fecal pathogens and organic micropollutants. However, there are equally obvious reasons, why nobody (to our knowledge) has tried out this process in practice: drying and burning feces (and toilet paper) in a bathroom is a complex and potentially risky undertaking. Some experience exists with respect to the combustion of manure, on the household scale in a number of developing countries and, on a larger, more technical scale, in industrialized countries with an overproduction of animal manure. From the latter experience, it is known that potential emissions of air-polluting compounds such as CO, NOx, and SOx must be taken into account (Florin et al., 2009; Lundgren and Pettersson, 2009) and that the amount and type of inorganic elements in different types of bio waste increase the risk of slagging and ash deposition in the combustion system (Miller and Miller, 2007). From this very limited experience, it seems that it would be rewarding, but far from trivial, to develop incineration technology for feces, especially in an on-site setting.
4.07.4.5.4 Biological treatment Aerobic treatment. Composting is the most widespread method of aerobic treatment of feces, with co-composting with organic solids being one of the standard treatment methods for FS in developing countries (Strauss and Montangero, 2002). These and other authors (e.g., Scho¨nning and Stenstro¨m, 2004), however, are of the opinion that most of the composting takes place at ambient temperatures with insufficient pathogen elimination. This is supported by a study of the survival of fecal coliforms in dry solar composting toilets in Mexico, where most of these toilets did not comply with the existing regulation (Redlinger et al., 2001). However, of those that did comply, nearly all were composting toilets with optimal solar exposure, confirming the importance of correct technical application of these single compost toilets. Thermophilic composting or co-composting at temperatures higher than 50 1C is given much attention in the scientific literature because a hygienic end product can be obtained with this technology (Ingallinella et al., 2002). From a technical point of view, there are several ways of obtaining the
necessary temperatures. Due to the heat-producing aerobic processes, sufficiently high temperatures can be obtained through co-composting with organic solid waste either in a compost heap (Kone´ et al., 2007) or in a single well-insulated toilet (Niwagaba et al., 2009). Solar heating of the toilets, as discussed above, may also be effective. In the specific compost heap referred to earlier, feces and organic solid waste were mixed in the ratio of 1:2, leading to a reduction of helminth eggs below the WHO guidelines of o1 viable egg/g TS (WHO, 2006). In the single toilet, a feces to solid waste ratio of 3:1 was apparently sufficient to obtain a reduction of E. coli 43 log10 and a reduction of Enterococcus spp. of 44 log10 units. However, based on experience, the authors conclude that mixing of the material is crucial in order to kill pathogens in the entire sample. Without addition of solid waste, no bacterial inactivation was observed. In the compost heap as well as in the single toilet, regular turning of the compost (approximately once a week) was required. In a single toilet, high temperatures can also be obtained through external heating. Since the temperature in such a setting can be chosen more freely, it makes sense to determine the optimal operational temperature. Lopez Zavala et al. (2004) showed that the rate of aerobic degradation increased with temperature between 20 and 60 1C, whereas no reaction took place at 70 1C. Using sawdust as a matrix and assuring that oxygen was not limiting for the process, stabilization of feces was obtained within 24 h at a temperature of 60 1C, whereas 72 h were required at 50 1C. According to the safety zone diagram set up by Feachem et al. (1983), the 24 h at 60 1C would be sufficient to ensure a safe end product, whereas at least 96 h would be necessary at 50 1C. However, for practical applications, the WHO guidelines from 1989 (cited in Scho¨nning and Stenstro¨m (2004)) recommend at least 1 month aerobic composting in piles at 55–60 1C; presumably because, in practice, the fecal matter will never be homogeneously mixed. The principle of a newer Japanese composting toilet (called the Bio-toilet) may help overcome some of the problems of the more conventional compost toilets. This toilet is based on automatic dosing of sawdust, an external heating system, and mechanical mixing (Nakagawa et al., 2006), and the authors model the die-off of virus as a function of temperature and water content. It should be noted that although from a hygienic point of view, heating alone would be sufficient, the esthetic quality of the fecal matter will profit from a biological process (Avvannavar and Mani, 2008). Vermicomposting. If FS is collected as dry matter, earthworms may be used for aerobic degradation, in a process called vermicomposting. For treatment of feces, this process has been studied far less than the more conventional composting discussed previously. However, some experience exists for treatment of sewage sludge as well as for on-site treatment of feces in Australia (Bajsa et al., 2003), and for on-site treatment of feces in Indonesia (Malisie et al., 2007) and Germany (Gajurel et al., 2007). Since vermicomposting is only possible within the relatively narrow limits of 70–90% water content (Edwards, 1995), feces from flush toilets can only be vermicomposted after dewatering, whereas feces from dry urine-separating toilets would fall within this limit (see Table 7). Vermicomposting will decrease the number of
Source Separation and Decentralization
pathogens (Bajsa et al., 2003), but there is still no conclusive evidence how effective this process is for feces (Gajurel et al., 2007). Treatment of black water in an MBR is not a typical technology choice, but it has been tested in laboratory scale (Atasoy et al., 2007) with very good results for COD and nitrogen removal (96% and 89%, respectively) and 100% removal of total coliforms. The specific energy demand was 2.30 kWh m3 at a concentration of 1218 gCOD m3 and 188 gTKN m3, which would correspond to approximately 5.75 W/person in this special case. However, based on Tables 4 and 5, we had to make these calculations based on the assumption that the black water production in this special investigation was very high (around 60 l/cap/day), which in the case of MBRs is relevant for energy consumption. Biological removal of odor has not been discussed in connection with source separation of feces, but in densely populated areas, good ventilation of the bathroom (see Section 4.07.4.4) must be followed by effective odor removal from the off-gas. Sato et al. (2001) suggested that the (very small amount of) malodorous compounds stemming from feces and urine could be efficiently degraded in a biological process. The authors identified that 90% of the substances responsible for the smell of stored wastewater are fatty acids (acetic, propionic, and butyric acid), which are well degradable in biological systems. Anaerobic treatment. Anaerobic digestion is a standard treatment method for sewage sludge, and also for the treatment of black water, anaerobic treatment is very well documented. The main arguments for anaerobic treatment of black or brown water as opposed to aerobic treatment are the possible recovery of energy (most often as methane) and the preservation of nutrients under anaerobic conditions (KujawaRoeleveld and Zeeman, 2006). The lower heating value of organic matter in the form of methane is 12.5 kJ g1 COD (Rittmann, 2006), resulting in a maximal energy content of feces of 750 kJ/cap/day or 8.7 W/cap (based on specific feces production as reported in Table 4). Up to 65% of the organic matter in feces can be transformed into methane (Feng et al., 2006), resulting in a potential maximal energy production of 5.7 W/cap/day. Anaerobic digestion of feces can be enhanced by the addition of kitchen refuse, a typical way of increasing COD concentration (Otterpohl et al., 1999). Kujawa-Roeleveld and Zeeman (2006) reviewed the more general knowledge about anaerobic treatment of black water, and the following summary is based on this article. A number of technical applications have been developed, mainly the completely stirred reactor (CSTR), the accumulation (AC) system, and the up-flow anaerobic sludge blanket (UASB) reactor. For nutrient recovery from the effluent, principally, the same technologies are available as for urine (see also Section 4.07.3.5). The main factors of importance for reactor performance are SRT and temperature. For energy balances, also the COD concentration of the influent is important: the higher the concentration, the less energy will be necessary for any heating anticipated and the lower the amount of methane lost to the atmosphere. Posttreatment of the anaerobic effluent will often be necessary in order to comply with effluent quality either to the receiving water (COD and nutrients) or to agriculture (pathogens and micropollutants).
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Like for other treatment methods, higher temperature and higher residence time will lead to higher pathogen removal. For a CSTR treating source-separated feces, a mathematical model for the influence of HRT and temperature was developed and calibrated at 55 1C, using fecal coliforms and intestinal enterococci as indicator organisms (Lu¨bken et al., 2007). In accordance with the general understanding of pathogen die-off, this model predicts that pathogen inactivation in such a reactor is only possible at temperatures beyond 50 1C. It is well known that ammonia is inhibitory to methanogenesis, which is one of the reasons that anaerobic digestion of brown water (liquid feces without urine) is sometimes preferred to anaerobic digestion of black water (Otterpohl et al., 1999). In an experimental investigation of the influence of free ammonia (NH3) and total ammonia concentration (NH3 þ NHþ 4 ; Lay et al., 1998), it was found that first signs of inhibition of methanogenesis occur at a total ammonia concentration of 1.7 gN l1, which would occur at a toilet flush volume of around 6–7 l/cap/day. Existing vacuum toilets would result in a toilet flush volume of around 5 l/cap/day (Table 7), and thus be just below this critical volume. The same authors also found that the lag phase in a batch experiment depends on the free ammonia level, identifying a shock level of 0.5 gN l1 of free ammonia (NH3). In dry systems without urine separation, one must expect ammonia inhibition of methanogenesis, which is also found experimentally (Chaggu et al., 2007). It should be noted that with a urine-separating toilet of the type Dubbletten/Gustavsberg, the same COD concentration can be obtained as with a conventional vacuum toilet (Table 7), with much simpler technology. As a further advantage, separate treatment of urine will avoid any risk of ammonia inhibition in an anaerobic treatment process. Anaerobic treatment is typically not foreseen for a single household and many projects operate on a scale from about 100 persons and upwards (see, e.g., Zeeman et al., 2008; Otterpohl et al., 1997). It has been shown that no microbial risk should arise from gas usage (Vinnera˚s et al., 2006b), but we have found no other risk analyses of decentralized gas production, neither with respect to possible greenhouse gas emissions nor with respect to any possible danger of explosions. However, in order to reduce the loss of methane, it is important to maximize COD concentration, for example, by using a urine-separating vacuum toilet, an approach that would also increase process stability (Elmitwalli et al., 2006).
4.07.4.5.5 Hygienization and removal of micropollutants Since one of the main objectives of feces treatment is hygienization, we have discussed this topic in nearly all the paragraphs of Section 4.07.4.5. Shortly summarized, pH, time, and temperature are the main parameters determining whether hygienic requirements can be reached. In addition, ammonia has proved to be effective for the removal of microorganisms. Since ammonia is an important plant nutrient, where some surplus is required for efficient plant uptake, this technology has a great potential for recycling of feces to agriculture as illustrated by the Peepoo bag, the self-sanitizing single-use biodegradable toilet (Vinnera˚s et al., 2009).
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Due to the organic nature of feces, there is no explicit way of removing only micropollutants from feces. From conventional wastewater treatment, it is known that a number of micropollutants are degraded in aerobic treatment, and there is no reason to believe that this would not also be the case for aerobic treatment of feces. We refer to the standard wastewater treatment literature on this topic. However, the typical treatment method for black and brown water is anaerobic treatment, and the few studies on anaerobic degradation of pharmaceuticals, which we have found, do not show much degradation (Mes et al., 2008). With aerobic treatment of the effluent from anaerobic treatment of black water, however, at least the soluble micropollutants may be degraded (de Mes et al., 2007). It is obvious that processes such as total oxidation and incineration will also eliminate the problem of micropollutants.
4.07.4.6 Summary With respect to urban hygiene, feces make up the most important part of wastewater. Separate feces handling is not only challenging, but also rewarding. The most important decisions concerning technology choice are taken with the choice of toilet. Two main factors are decisive: water consumption and separation of urine. A dry toilet allows for fundamentally different treatment technologies as compared to a flush toilet, but most conveyance systems require a certain amount of flush water. Urine source separation is essential for dry toilets and advantageous for water-saving toilets with anaerobic treatment. For reasons of energy conservation, anaerobic treatment of feces from flush toilets is often preferred, whereas drying or biological aerobic processes are common for feces from dry toilets. It is no surprise that for the products, esthetics and hygiene are the most important parameters.
4.07.5 Combined Domestic Wastewater There is a long tradition for decentralized treatment of combined wastewater in sparsely populated areas where conveyance systems are either too expensive or technically not feasible. There is a large wealth of technologies and options available to treat combined wastewater and it would exceed the scope of this chapter to review the entire literature on these small and well-known treatment plants. However, we briefly discuss the main trends of newer research into onsite treatment of domestic wastewater and thereby illustrate the problems of this approach. The advantages, however, are evident: treating combined wastewater on-site avoids the problems of reinventing household devices, and concentrates the efforts on established and tested technologies that wastewater treatment professionals feel comfortable with. Additionally, on-site treatment enables the simplification or even the abolishment of large and investment-intensive conveyance networks. We exemplify our main points on on-site treatment of combined wastewater with two types of common treatment technology that represent the two sides of the spectrum: (1) the septic tank, which is the simplest of all possible systems and abundant in many rural areas all over the world, and (2) the Japanese johkasou as a standardized mass-produced enhanced treatment unit.
Table 8 Suggested values for typical production of combined purely domestic wastewater (water, total organic matter, nitrogen and phosphorus) based on Tables 1, 4, and 5 and the assumption of an average 6-l toilet standards resulting in 30 l/cap/day of water for toilet flushing Volume (l/cap/day) COD (gCOD/cap/day) Nitrogen (gN/cap/day) Phosphorus (gP/cap/day)
130 120 14 2.0
4.07.5.1 Production Rate and Composition of Combined Domestic Wastewater Numbers for the production rate and composition of combined wastewater often include other types of wastewater than domestic. The numbers indicated in Table 8, in contrast are ‘estimated’ instead of ‘intended’ (the program does not allow me to erase the word ‘intended’) intended for purely domestic wastewater. Like for gray water, large variations of the water production rate must be expected.
4.07.5.2 The Risks of On-Site Treatment of Combined Wastewater Traditionally, the hygienic risks of on-site treatment have been avoided by subsurface evacuation through a septic tank. Since the household does not get into contact with the on-site treatment plant, the same direct safety level as for centralized treatment is obtained. For aboveground on-site treatment, there is a potential hygienic risk, especially if the household itself is involved in maintenance and sludge removal. In Japan, where such systems are frequent, this risk is minimized by an elaborate scheme for control, maintenance, and sludge handling by professional companies (Yang et al., 2001). More difficult to avoid is the risk of contamination of receiving waters, that is, groundwater (e.g., in the case of septic tanks) or surface water. The main difference to central treatment is the generally uncontrolled discharge of wastewater, due to low treatment efficiencies, malfunctioning, or leaks. The risk of groundwater contamination from septic tanks is well known and widespread. Wistrom et al. (2003) referred to extensive literature on this problem in the United States, explicitly mentioning contamination of drinking water wells with fecal coliforms, nitrate, and phosphorus from septic tank treatment of combined wastewater. The same problems are observed with the johkasous in Japan (Yang et al., 2001; see also Sections 4.07.5.3.1 and 4.07.5.3.2). The risks from reusing treated wastewater are well known and the discussion is not much different from the discussion on reuse of gray water (see Sections 4.07.2.2, 4.07.2.3, and 4.07.2.4). An obvious difference of treated combined wastewater is the potential to return nutrients to agriculture.
4.07.5.3 Two Examples of On-Site Treatment Technologies for Combined Domestic Wastewater 4.07.5.3.1 Septic tanks The traditional Western on-site wastewater treatment process for areas with low population density is an underground
Source Separation and Decentralization
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septic tank. Not only in the US, but also in other countries such as Germany and France, this is the most widespread onsite treatment technology for domestic wastewater (EPA, 2002). The technology is based on primary sedimentation and HRT to reduce pathogens. The removal of particulate matter reduces clogging in the subsurface outflow. Sludge reduction through anaerobic digestion is an important issue in order to reduce the frequency of septic tank emptying (e.g., Fleckseder and Krejci, 1982). However, due to low subsurface temperature, only acidification may take place, resulting in the transformation of sludge into soluble COD instead of methane production (EPA, 2002). As discussed in Section 4.07.5.3.3, uncontrolled methane production may not be the process of choice from a global-warming point of view, but it is obvious that mobilized organic COD can also cause problems in some settings. There are several other reasons that traditional septic tanks are not an adequate solution for today’s problems, even in rural areas: poor nutrient removal (e.g., Gill et al., 2008), risk of microbial contamination of aquifers (e.g., Scandura and Sobsey, 1997), and contamination of aquifers with micropollutants (e.g., Godfrey et al., 2007). Since septic tanks are widespread, and the problems are recognized, there are many attempts to improve the technology. In some cases, septic tank effluent is further treated to avoid contamination of receiving water (e.g., Hu et al., 2007); in other cases, the septic tank itself is upgraded for enhanced biological treatment (e.g., in the form of an UASB septic tank (Al-Shayah and Mahmoud, 2008) or simply with filter material (Zhang et al., 2009)). There is no doubt that both approaches are possible, and with enough effort, they will both lead to success. However, if the simplicity of the septic tank approach shall be retained and treatment requirements are high, such solutions tend to be very space demanding, moving toward systems based on natural self-purification. For more compact solutions, simplicity tends to get lost. It is consequently a question, whether it will not be cheaper to go for an on-site reactor in the first place, which can be mass-produced. The principles of such on-site reactors can be well discussed based on the Japanese example of the johkasou (Section 4.07.5.3.2).
by Yang et al. (2001). There are two types of johkasous: the tandoku-shori johkasou for treatment of toilet waste (not discussed in Section 4.07.4.5 because there is no information in English on this technology) and the gappei-shori johkasou for combined wastewater. With the latter type, the wastewater of about 10 million people in rural Japan is treated. The johkasous are mass-produced and typically consist of a sedimentation chamber, where primary sludge is anaerobically degraded like in a septic tank, followed by biological aerobic treatment. The biological compartment is often based on filter and contact media, but activated sludge systems also exist (Nakajima et al., 1999). Similar to the situation of septic tanks, there is a trend toward improvement of the johkasous, because they do not sufficiently address today’s problems of environmental pollution, even in rural areas. In contrast to the septic tanks, however, the improvements of johkasou technology more closely mirror the developments in central treatment: improved nutrient removal (nitrification/denitrification), reduction of tank volume with membrane technology, improved disinfection of the effluent, and reuse of sludge for agriculture. Similar systems are also in use in Europe, with or without membranes. An interesting feature is the thermal destruction of sludge, reducing the need for emptying and maintenance (Wistrom et al., 2003). In some cases, the sedimentation step is omitted in order to avoid odor problems from primary sludge, thus allowing for in-house installations and the use of filter bags for sludge dewatering (Abegglen et al., 2008). In Table 9, we report some results on volume, BOD and nitrogen removal, and energy consumption of different types of gappei-shori johkasous, all from Yang et al. (2001). It is obvious that the use of membrane johkasou drastically improve the quality of wastewater treatment (at a similar or lower specific tank volume), but at the expense of higher energy consumption. Denitrification in a membrane johkasou is obtained either in a separate denitrification tank (Ohmori et al., 2000) or by using the anaerobic first sedimentation reactor for denitrification (Yang et al., 2001). In both cases, denitrification relies on recirculation of nitrified wastewater.
4.07.5.3.2 Johkasous
Considering the high global warming potential (GWP) of methane, it is surprising that widespread on-site technologies for disposal of domestic wastewater still rely on uncontrolled
4.07.5.3.3 On-site uncontrolled anaerobic digestion Due to a pronounced lack of information on johkasous in English, the following summary is mainly based on an article
Table 9
Volume, BOD, and nitrogen removal and energy consumption of different types of gappei-shori johkasous
Specific volume Energy consumption Percentage of reactors keeping the following targets: BODo5 mg l1 and T-No10 mg l1 BOD and T-No20 mg l1 BODo20 mg l1 and T-N420 mg l1 BOD 4 20 mg l1 and T-No20 mg l1 BOD 4 20 mg l1 and T-N420 mg l1
Type A
Type B
Type C
m3/cap W/cap
0.6–0.7 16–18
0.5 18
0.6 14–21
% % % % %
64 18 11 7
81 13 3 3
83 15 2
Membrane 0.5 32 100
Type A: with contact aeration tank; type B: with biofilm filtration tank; type C: with moving bed biofilm tank; membrane: membrane johkasou, only lab results. Based on Yang XM, Yahashi T, Kuniyasu K, and Ohmori H (2001) On-site systems for domestic wastewater treatment (johkasous) in Japan. In: Lens P, Zeeman G, and Lettinga G (eds.) Decentralised Sanitation and Reuse, Integrated Environmental Technology Series, pp. 256–280. London: IWA.
224 Table 10
Source Separation and Decentralization The global warming potential (GWP) of uncontrolled anaerobic degradation of COD from combined domestic wastewater
GWP(20) of CH4 (timescale 20 years) GWP(100) of CH4 (timescale 100 years) Specific CH4 production from COD Specific GWP(20) of methane production from COD Specific GWP(100) of methane production from COD Specific production of COD Assumed primary sludge production (25% of COD) Assumed primary sludge degradation (50% of sludge) GWP(20) based on assumptions above GWP(100) based on assumptions above Specific CO2 production from electricity production CO2 production from typical 10 W/cap WWTP
gCO2 =gCH4 gCO2 =gCH4 gCH4 =gCOD gCO2 =gCOD gCO2 =gCOD gCOD/cap/day gCOD/cap/day gCOD/cap/day gCO2 =cap=day gCO2 =cap=day gCO2 =Wh gCO2 =cap=day
72 25 0.25 18 6 120 30 15 270 90 0.8 192
a a b
a a d
e f
a
IPPC (2007: ch. 2, p. 212). Based on a mass conservation of theoretical COD (Gujer and Larsen, 1995). c From Table 8. d Typical values from centralized treatment of municipal wastewater (Gujer, 2007). Note that the anaerobic degradability is for primary and secondary sludge; for primary sludge it will be higher. e Based on EU 15 electricity mix from 1997 (European Environmental Agency, 2002). f A typical advanced wastewater treatment plant (WWTP) consumes around 10 W/person of electrical power (1 W/cap ¼ 24 Wh/cap/day). b
anaerobic digestion. For septic tanks, the methane production depends on the subsurface temperature (see Section 4.07.5.3.1), whereas the better-controlled johkasous are more likely to give rise to methane gas production. We therefore briefly illustrate the implications of uncontrolled methane production from on-site reactors. Since it is difficult to obtain any exact data on sedimentation in septic tanks and johkasous, let alone data on anaerobic degradability, we only present specific data and one example based on primary sedimentation in treatment plants (Table 10). Methane production from COD elimination is based on the concept of mass conservation of theoretical COD (Gujer and Larsen, 1995). The GWP is presented for a time period of 20 and 100 years (the latter timescale normally being considered of relevance). For comparison, we also present the GWP of electricity production, because electricity is the dominant energy form used for wastewater treatment technologies. Obviously, the net greenhouse gas emission from electricity production greatly varies (from close to zero for renewable energy sources to about 1 g CO2/Wh for small coal-fired power plants; Bettle et al., 2006); however, for simplicity, we use a value corresponding approximately to the European electricity mix (0.8 g CO2/Wh; European Environmental Agency, 2002). From Table 10, we can get an idea whether methane production from on-site anaerobic degradation of primary sludge is relevant or not. A typical modern wastewater treatment plant has an electricity consumption of about 10 W/person (corresponding to 240 Wh/cap/day), whereas the membrane johkasous are about 3 times as energy intensive (Table 9). We thus see that in the short term (20-year timescale), the GWP of methane production from primary sludge is similar to the GWP arising from electricity use in a modern centralized treatment plant. Due to the much shorter lifetime of methane in the atmosphere, it is obvious that the longer the time horizon, the smaller the contribution will appear. In conclusion, from a global warming perspective, the energy use for enhanced treatment is justified, especially if the electricity used has a small carbon dioxide footprint.
One way of reducing anaerobic digestion of primary sludge in johkasous, for example, is denitrification in the sludge compartment as described in Section 4.07.5.3.2. In septic tanks, at least, where nitrification and recycling of nitrified wastewater are difficult, separate nitrification of urine could be a good alternative. From Table 9 it can be concluded that about 10 g of NO 3 N/cap/day can be removed by denitrification, corresponding to around 29 g/cap/day of COD removal (based on mass conservation of theoretical COD; Gujer and Larsen, 1995). In principle, this would be more than enough to suppress anaerobic degradation of primary sludge, even if the assumptions made in Table 10 are conservative. More detailed investigations would however be necessary to test these assumptions, and obviously denitrification processes may also give rise to global warming, if not properly controlled.
4.07.5.4 Summary Decentralized treatment of combined wastewater is a typical application in rural areas of industrialized countries. In this section, we have discussed two exemplary decentralized technologies: (1) the septic tank, which is the simplest of all possible systems and abundant in many rural areas all over the world, and (2) the Japanese johkasous as a standardized massproduced enhanced treatment unit. Where the septic tank is an underground technology, with only periodical emptying, the johkousa requires an institutional setup for emptying, maintenance, and control. Today, it is recognized that both types of reactors do not fulfill modern requirements for waterpollution control, even in rural areas and they are both further developed. Whereas the septic tanks are developed more in direction of systems relying on natural self-purification, the johkasous are developed along the lines of conventional central wastewater treatment. From a global warming perspective, the energy use for enhanced treatment is justified, especially if the electricity used has a small carbon dioxide footprint.
Source Separation and Decentralization
4.07.6 Outlook We do not draw special conclusions from this chapter. As illustrated in Section 4.07.1, there are good reasons to work on the development of decentralized technologies based on source separation, but it is way too early to conclude which technology or technologies will be the choice of the future. However, also decentralization without source separation holds promise for a more resource-efficient future: without the large expenditures for long-lived sewer systems, flexibility is retained for introducing source separation later on, when such technologies have matured. We hope that the overview in this chapter will inspire more scientists to leave the well-thread path of centralized wastewater treatment and get involved in decentralization and source separation.
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Relevant Websites http://www.novaquatis.eawag.ch Novaquatis: ein Querprojekt der Eawag. www.roevac.com Roediger Vacuum: World Leading Systems for the Collection of Wastewater Using Vacuum.
4.08 Modeling of Biological Systems M Wichern, T Gehring, and M Lu¨bken, Institute of Environmental Engineering, Bochum, Germany & 2011 Elsevier B.V. All rights reserved.
4.08.1 4.08.2 4.08.2.1 4.08.2.1.1 4.08.2.1.2 4.08.2.1.3 4.08.2.1.4 4.08.2.2 4.08.2.3 4.08.3 4.08.3.1 4.08.3.1.1 4.08.3.1.2 4.08.3.2 4.08.3.3 4.08.3.4 4.08.3.5 4.08.4 4.08.4.1 4.08.4.2 4.08.4.2.1 4.08.4.2.2 4.08.4.2.3 4.08.4.2.4 4.08.4.2.5 4.08.4.3 4.08.4.3.1 4.08.4.3.2 4.08.5 4.08.5.1 4.08.5.2 4.08.5.2.1 4.08.5.2.2 4.08.5.2.3 4.08.5.2.4 4.08.5.2.5 4.08.5.3 4.08.5.3.1 4.08.6 4.08.6.1 4.08.6.2 4.08.6.2.1 4.08.6.2.2 4.08.6.2.3 4.08.6.3 4.08.6.3.1 4.08.6.3.2 4.08.6.3.3 4.08.6.3.4 References
Introduction Mathematical Modeling of Biochemical Processes Biological Processes Carbon removal Nitrification Denitrification Biological P-elimination Modeling of Hydraulic Conditions in Activated Sludge Plants Modeling of Biochemical Processes Modeling of Biological Processes in Activated Sludge Systems Introduction Activated Sludge Model No. 3 EAWAG BIOP Module WWTP at Koblenz WWTP at Hildesheim WWTP at Duderstadt Calibrated Biochemical Parameters and COD Influent Fractionation Soil Filters Introduction Material and Methods Pilot-scale sand filter Analytical methods Mathematical model Biological processes Biofilm modeling Results and Discussion Model calibration and simulation results Sensitivity analysis Waste Stabilization Ponds Introduction Material and Methods Description of the pilot pond Mathematical model Hydraulic concept Algal processes Physico-chemical processes Results and Discussion Model calibration and simulation results Anaerobic Treatment Introduction Material and Methods Analytical methods Reactor operation Mathematical model and sensitivity functions Results and Discussion Calibration of the ADM 1 Modeling reactor performance Sensitivity analysis for the biochemical parameters and the inflow fractioning Simulation of the energy balance
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4.08.1 Introduction
4.08.2 Mathematical Modeling of Biochemical Processes
Sanitary engineering is an extremely complex field of work. Practical experience and understanding of basic principles of engineering sciences, biology, hydrology, and computer science, as well as of social science and economics are necessary to deal with water issues in an efficient and sustainable manner. Sanitary engineering covers the specific fields of drinking water, water supply, sewage disposal, wastewater treatment, and river water quality. In many cases, there are overlaps and intersections between individual areas of expertise. In this chapter, the focus is on the application and development of mathematical models for wastewater treatment. Mathematical models can supply important information for the basic understanding of biochemical microbial conversion in technical systems, solve questions of optimal plant operation to save energy or investment costs, and contribute to the management of highly complex systems. Especially in developing countries with their demands for new and sustainable water solutions, integral approaches have to be developed to bring together technical, economic, and social objectives. To achieve these goals, mathematical modeling is an essential tool. Section 4.08.2 provides a short introduction to the mathematical modeling of biochemical processes. In Section 4.08.3, results from the calibration and validation of the Activated Sludge Model No. 3 (Gujer et al., 1999) for German wastewater are presented, including detailed information on how to deal with nitrogen and enhanced biological phosphorus removal in the mathematical model. Section 4.08.4 discusses the dynamic modeling of biochemical processes in soil filters. Apart from issues like how to calculate substrate conversion and how to anticipate soil clogging, this section informs about the modeling of biofilms, discusses how the growth of the microbial community which is formed on the single grains is limited by diffusion, and points out how the detachment of cells is one of the most sensible factors influencing the results of biofilm modeling. Section 4.08.5 describes the simulation of facultative lagoons and maturation ponds. The newly developed model expands the ASM 3 by adding algae growth and decay, sun radiation, ammonia stripping, wind impact, and ionic equilibrium as variables. Section 4.08.6 then presents the mathematical modeling of anaerobic processes. The Anaerobic Digestion Model No. 1 (Batstone et al., 2002) is calibrated for agricultural substrates and extended by equations to calculate energy balances.
Biological processes in wastewater-treatment plants (WWTPs) are based on natural processes in waterways. In waterways, microorganisms oxidize carbon into CO2, as well as water and ammonia nitrogen into nitrate nitrogen and elementary gaseous nitrogen. These conversion processes take place with the help of both suspended and attached biomass, with the latter finding its growth areas on stones and plants. Although conversion processes in attached and suspended biomass are similar, the mathematical modeling of attached biomass is considerably more difficult. There, substance conversion can be restricted by diffusion processes of the substrate to the cells. For suspended biomass, it is assumed that the diffusion of substrate into the interior of the floc is no rate-limiting step. Although substance conversion processes in waterways and WWTPs are based on the same principles, in WWTPs considerably higher conversion rates are achieved. This is due to the amount of provided substrate and the quantity of active biomass which is retained in the system through sedimentation tanks. The increased demands on environmental protection and further development of the process technology are met through the dimensioning of WWTPs with complex biochemical models. Apart from the dimensioning, mathematical simulation is used for the optimization of existing treatment plants, to achieve better effluent concentrations, lower sludge production, and oxygen input and to minimize operational costs. The following sections serve to explain the basic principles of the mathematical modeling of biochemical processes. Thus, this chapter deals with the hydraulic presentation of the flow stream, biological conversion processes, and sedimentation of biomass in the secondary clarifier tank. The aim of biological wastewater treatment is to remove those substances contained in the wastewater which are hazardous for humans and/or nature, or to change them in such a way as to render them harmless. Biological wastewater treatment realizes and technically utilizes the natural processes of transforming pollutants into inorganic and organic final products by microorganisms. In WWTPs, these conversion processes take place in the activated sludge tank, where wastewater and activated sludge are mixed and aerated. The oxygen necessary for the biological degradation is provided by aeration implements. From the activated sludge tank, the wastewater–sludge mixture flows into the secondary clarifier tank, where the activated sludge separates from the treated wastewater (see Figure 1). The sludge settling in the secondary
Clarifier Inflow
Effluent Aerated tank Surplus sludge
Return sludge Figure 1 Activated sludge plant consisting of aerated tank and secondary clarifier tank.
Modeling of Biological Systems
clarifier tank is conveyed back into the activated sludge tank as recirculation sludge, while the treated wastewater flows off. During the biological processes, the amount of activated sludge increases and is then removed as surplus sludge. Activated sludge processes are characterized by aerobic environment (dissolved oxygen available), anoxic zones (no dissolved oxygen available, but bound oxygen in the form of nitrate), and anaerobic zones (neither dissolved nor bound oxygen available). While aerobic and anoxic zones are essential to remove chemical oxygen demand (COD), nitrogen, and phosphorus, anaerobic zones in activated sludge processes are mainly established in order to facilitate enhanced biological phosphorus elimination.
4.08.2.1 Biological Processes 4.08.2.1.1 Carbon removal One crucial aspect for the dimensioning of the aeration and for the amount of produced surplus sludge in municipal WWTPs is the elimination of carbon. The oxidizable carbon can approximately be described as biochemical oxygen demand (BOD) or COD. Nowadays, COD is increasingly used to describe the organic load of a WWTP. COD represents the oxygen amount necessary to oxidize organic carbon into CO2 and water. The major advantage of COD is that it is easy to create the COD balance. The principle of this balancing for an activated sludge plant is presented in Figure 2 as an example. It becomes obvious that the influent COD must be equal to the sum of effluent COD, oxygen demand, and surplus sludge.
4.08.2.1.2 Nitrification During nitrification, ammonium nitrogen (NH4–N) is biochemically converted into nitrate nitrogen. This process is affected by bacteria which oxidize inorganic nitrogen into nitrate nitrogen (NO3–N), and which commonly are referred to as nitrifiers. These are obligatorily autotrophic microorganisms, which require CO2 and thus do not need any further organic carbon for the growth of their biomass. Nitrification takes place in two stages: first nitritation, then nitratation, both with different microorganisms.
4.08.2.1.3 Denitrification Denitrification is the conversion (reduction) of nitrate into gaseous nitrogen (N2). Many heterotrophic bacteria (denitrifiers) are able to consume the nitrate oxygen instead of
bound oxygen. If oxygen is available (aerobic milieu), the microorganisms will, as a rule, always prefer the O2 respiration; only in times of oxygen shortage and presence of nitrate and/or nitrite (anoxic milieu) will they switch to denitrification. Thus, nitrate is removed from wastewater by denitrification only if there is a shortage of dissolved O2.
4.08.2.1.4 Biological P-elimination Next to the elimination of nitrogen, the removal of phosphorus from the wastewater is of major importance. Phosphorus removal is achieved by bacteria which, under certain conditions, take up an increased amount of phosphate. By extracting these bacteria with the surplus sludge, phosphorus can be removed. The increased phosphate incorporation of the bacteria occurs if the organisms are first subjected to oxygen-free (anaerobic) zones, where they deliver phosphate into the wastewater (release), and then are moved to aerated (aerobic) zones, where they increasingly incorporate phosphate. Generally, enhanced biological P-elimination is thus coupled to a switching between anaerobic and aerobic or anoxic conditions.
4.08.2.2 Modeling of Hydraulic Conditions in Activated Sludge Plants Today, one-dimensional (1D) or multi-dimensional hydrodynamic models are used for complex simulations of river water quality. Hydrodynamic models are particularly recommended if highly unsteady effluent conditions must be expected. These models are based on the St.-Venant equations, which fulfill continuity conditions, and on equations of motion. 1D models with rectangular profiles are used for waterways if measuring data are not available in sufficient quantity and quality, or if an integrated approach is needed, where sewer system, WWTP, and river have to be considered together. Plug-flow reactors, which imitate the situations in actual waterways, are a sensible solution if strongly unsteady waterway conditions appear more rarely, if the target is to make statements about the capacity of the respective river in regard to the substance degradation of stationary loads, or if annual balances of substance loads should be determined. This reactor type should also be used for the simulation of activated sludge plants. Strong plug-flow is observed, for instance, in recirculation ditches. Figure 3 depicts the substrate degradation in a plug-flow reactor.
2500 kgO2 d−1 demand Clarifier Inflow Aerated tank
Effluent 500 kgCODd−1 Surplus sludge
5000 kgCODd−1
2000 kgCODd−1
Return sludge Figure 2 Exemplary COD balance for an activated sludge plant.
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c(t)
c (t)
Reactor length Time
Figure 3 Simplified substrate degradation in a plug-flow reactor.
Figure 5 Substrate degradation in a batch reactor over time.
substrate concentration C (if the reaction rate is assumed as k*C):
c (t)
C¼
C0 ðg m 3 Þ 1 þ ðk V=QÞ
ð3Þ
where k is the reaction rate constant (d1). This also corresponds to the equation for the first CSTR reactor:
C1 ¼
Time Figure 4 Substrate degradation in the CSTR reactor.
ðg m 3 d 1 Þ
C2 ¼
ð5Þ
Thus, the following is a general rule:
Cn ¼ where C is the substrate concentration (g m ), t the time (d1), Q the flow rate (m3 d1), A the reactor cross-section area (m2), z the space (m), and r the substrate conversion rate (g m3 d1). Often recirculation ditches are modeled not as a plug-flow reactor, but as a series of fully mixed reactors. Based on the hydraulic regime, it should be determined with how many fully mixed reactors (completely stirred tank reactors, CSTRs) the plug-flow reactor can be approximated. Mathematically, a cascade of an infinite number of CSTRs in line corresponds to a plug-flow reactor. Substrate degradation in a CSTR can be approximated as shown in Figure 4. One can assume that the substrate concentration in the reactor is equal to that of the reactor effluent. The corresponding differential equation is as follows (a constant reactor volume provided):
ðg m 3 d 1 Þ
C1 ðg m3 Þ 1 þ ðk V=QÞ
ð1Þ
3
dC Q ¼ ðC0 CÞ r dt V
ð4Þ
with C1 being the effluent concentration of the reactor (g m3). The following applies to the second reactor, if its influent concentration is the effluent concentration of reactor 1:
The corresponding differential equation expresses the dependence of concentration on both flowing distance and time:
dC Q dC ¼ r dt A dz
C0 ðg m3 Þ 1 þ ðk V=QÞ
ð2Þ
with V being the volume (m3) and C0 the inflow substrate concentration (g m3). For steady-state conditions (substrate concentrations do not change with time), it can be easily explained how a series of fully mixed reactors will provide approximately same results as a CSTR cascade. In steady states, the following equation results, broken down for
Cn1 ðg m3 Þ 1 þ ðk V=QÞ
ð6Þ
If all previous equations are considered, for a reactor n the following can be concluded:
Cn ¼
1 1 þ ðk V=QÞ
n C0
ðg m3 Þ
ð7Þ
where n is the number of reactors in serie. In a batch reactor with a fixed volume, biochemical reactions are investigated without reactor input or output (Figure 5). With wastewater treatment, for instance, respiration experiments are run to determine the respiration activity of the sludge with different substrates. The mathematical description is quite simple:
dc ¼r dt
ðg m 3 d 1 Þ
ð8Þ
with r being the conversion rate in (g m3 d1).
4.08.2.3 Modeling of Biochemical Processes Based on the previous considerations, the following section serves to explain how biological degradation processes are described mathematically. The following considerations will be based on the CSTR reactor. Conversion rates (kinetics) of biological processes are often described using Monod kinetics, with
Modeling of Biological Systems
differentiation into growth kinetics and inhibition kinetics. With growth kinetics, the conversion rate rises with an increased substrate supply; with the inhibition kinetics, the conversion rate decreases due to inhibition of biochemical processes. A typical example of growth kinetics can be found in nitrification. The conversion increases with the presence of ammonium. At WWTPs, inhibition kinetics gain importance, for instance, with denitrification, which may eventually come to a standstill by oxygen. Figure 6 shows a typical example to describe the dependence of the conversion rate on a substrate – growth kinetics at lower substrate concentrations and inhibition kinetics at higher ones. The kinetics of the process illustrated in the diagram then can be described:
m ¼ mmax
S KI ðd1 Þ S þ KS KI þ S
ð9Þ
where m is the growth rate (d1), mmax the maximal growth rate (d1), S the dissolved substrate (g m3), and KS and KI the substrate half-saturation and inhibition concentrations (g m3), respectively. Monod kinetics can be further simplified depending on substrate concentrations (Figure 7). If it can be assumed that substrate is always available in more than sufficient amounts (SbKS), the growth term becomes 1. Then, if inhibition is not considered, the following applies: m ¼ mmax, resulting in an equation of zero order. In cases when substrate is available only at very low concentrations, for instance below the half-saturation constant KS, the reaction rate can be reduced to r ¼ m*S (equation of first order). Apart from the degradation rate of a process, the substance conversion as such must be described. This happens on the basis of chemical equations. A typical example is the degradation of carbohydrates into biomass, CO2, and water (Gujer, et al. 1999):
C6 H12 O6 þ 2:45O2 þ 0:71NH3 -0:71C5 H7 NO2 þ 2:45CO2 þ 4:58H2 O 0.6
ð10Þ
This equation is often used to determine the yield rate (Y). The yield states how many grams of biomass develop from 1 g of influent substrate. The yield rate is commonly rendered in COD units. For the equation mentioned above, the following applies:
CODðC5 H7 NO2 Þ 160 ¼ 0:71 Y ¼ 0:71 CODðC6 H12 O6 Þ 192 ¼ 0:59 ðgcod;biomass g1 cod;
dXB V ¼ Q XB;o XB rV dt
rV ¼ 1mmax
/max
ðg d1 Þ
S So XB V 1bXB V S þ KS So þ Ko
ð12Þ
ðg d1 Þ
ð13Þ
where Ko is the oxygen haf-saturation concentration (g m3) and b the biomass decay rate (d1). For each process, the stoichiometric factor and kinetics are multiplied. Then, the single processes relevant for one parameter (here: S, So, and XB) are summarized. Stoichiometric factors are determined on the basis
0.4
0.3
0.2
0.1
0 10 S/Ks Figure 6 Growth and substrate inhibition with one substrate.
ð11Þ
with XB,0 being the inflow biomass concentration (g m3). Substrate conversion is described by the term r*V, which results from the matrix. It is comprised of the stoichiometric factor and kinetics, with both factors being multiplied by each other. For the biomass XB, the following applies:
0.5
5
inflow Þ
In a simplified manner, both the chemical equation and degradation kinetics are summarized in a matrix (see Table 1). The matrix contains two processes (biomass growth and decay) and three substrates (COD: S, XB: biomass, oxygen: SO). The matrix serves to present processes in a clearly structured manner and can be translated into the differential equations for any substrate. This is explained below for a CSTR reactor, using the biomass as example. Generally, the differential equation for each substrate comprises influent load per time, effluent load per time, and substrate conversion per time. Thus, for the biomass XB in a CSTR reactor, the following applies:
Growth inhibition
0
235
15
20
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1
0th Order
1st Order
0.9 0.8
/max
0.7 0.6 0.5 0.4 0.3 0.2 0.1 0 0
5
10
15
20
S/Ks Figure 7 Reduction of the Monod kinetics according to the frame conditions in equations of 0th and 1st order.
Table 1
Matrix notation
Description
Stoichiometry S
Biomass growth
1 Y
Biomass decay
Kinetic: process rate p XB
S0 1
1
of chemical equations. Each process is comprised of a chemical equation and kinetics. If all substrates are regarded in the balance, the COD of the products must be equal to the COD of the reactants, which means that the COD balance is closed. In the present case, the following applies for the biomass growth:
SðsubstrateÞ þ So ðoxygenÞ-XB ðbiomassÞ þ H2 O þ CO2 ð14Þ As the COD of H2O and CO2 is zero, the equation can be reduced, and the following results:
SðsubstrateÞ þ So ðoxygenÞ-XB ðbiomassÞ
ð15Þ
If this equation is expressed using the yield coefficient Y, the following applies:
1 ð1 YÞ ¼ Yð2Þ
ð16Þ
This means that from one unit of substrate, there results Y units of biomass under reduction of 1 Y units of oxygen. It must be considered here that oxygen has a negative COD, so that a negative sign before (1 Y) needs to be included into the equation. Two transformations then yield the stoichiometric factors, which can also be found in the matrix:
1 1Y ¼ 1 ð2Þ Y Y 1 1Y þ1 ¼0 ðÞ Y Y
ð17Þ
ð18Þ
ð1 Y Þ Y
mmax
1
bX B
S So XB S þ K S So þ K o
4.08.3 Modeling of Biological Processes in Activated Sludge Systems 4.08.3.1 Introduction In sanitary engineering, the most extensive experience has been made with the modeling of activated sludge systems. The first models were developed as early as at the beginning of the 1980s and led to the development of the Activated Sludge Model No. 1 (Henze et al., 1987). Nowadays, processes such as carbon removal, nitrification, and denitrification as well as the biological and chemical phosphorus removal are considered in the models. The focus of later models is on the description of enhanced biological phosphorus removal (EBPR). In an anaerobic environment, phosphorus is released when substrate is stored into cell biomass. In the following aerobic and anoxic conditions, polyphosphate accumulating organisms (PAOs) grow on stored substrate and are simultaneously take up phosphate to larger extent than they released before. To describe EBPR mathematically, models such as the EAWAG-BioP-Module (Rieger et al., 2001) in connection with the Activated Sludge Model No. 3 (ASM3; Gujer et al., 1999), the TUD model from the University of Delft (Murnleitner et al., 1997), the model from Barker and Dold (1997), and the Activated the Sludge Model No. 2d (Henze et al., 1999) are widely used. It is common to all models that Petersen matrix (Petersen et al., 2000) is used for the description of the biochemical processes. In this chapter, the focus is on investigations that were accomplished by ASM 3 to
Modeling of Biological Systems
model COD and nitrogen removal, and on the EAWAG BioP module to describe EBPR. The computations were accomplished by using software Simba 4.2 (2005) and Matlab/ Simulink (2005).
4.08.3.1.1 Activated Sludge Model No. 3 The Activated Sludge Model No. 3 was published in 1999 by the IWA task Group on Mathematical Modelling for Design and Operation of Biological Wastewater Treatment (Gujer et al., 1999). Better possibilities to identify biological processes today have resulted in the development of ASM 3 to simulate nitrification, denitrification, and degradation of COD. ASM 3 describes the storage of organic substrates, and the decay of heterotrophic organisms is modeled by the endogenous respiration. Heterotrophic yields are considered separately for aerobic and anoxic environments. The effect of redox conditions on the hydrolysis was assumed to be negligible. The hydrolysis of nitrogen was combined with the process of COD hydrolysis. Decay rates for endogenous respiration of heterotrophic and autotrophic organisms are reduced under anoxic conditions (Nowak, 1996; Siegrist et al., 1999). The following simulations were based on findings of Koch et al. (2000).
4.08.3.1.2 EAWAG BIOP Module The EAWAG Bio-P Module (Rieger et al., 2001) extends ASM 3 by processes which describe biological phosphorus removal. General principles of ASM 3 can be found again in EAWAG BioP Module. Endogenous respiration replaces lysis, different yields for aerobic and anoxic conditions are used, and the anaerobic decay was neglected for the PAOs. Eleven processes
237
were implemented additionally into ASM 3 to describe EBPR. Fermentation of COD is not modeled as a single process, nor was glycogen considered as an additional substrate pool for growth. Processes of chemical precipitation were integrated additionally in the same way as they were used in ASM 2d (Henze et al., 1999).
4.08.3.2 WWTP at Koblenz The wastewater treatment plant Koblenz was extended until 1992 to a population equivalent of PE ¼ 320000. The plant was designed for pre-denitrification and consists of a primary treatment unit, a first biological stage using trickling filter (only operated in the case of storm water flow), and a second biological stage built as activated sludge system. Rectangular tanks with horizontal flow are used as sludge sedimentation tanks. The WWTP is characterized by two different lines with different tanks in series and different operational conditions. In both lines, strong substrate gradients can be found, such that the plug-flow characteristic needs to be considered in the models. Figure 8 shows the treatment system. Both lines, consisting of activated sludge tanks and secondary settling tank were modeled in detail. Measured data was available in the inflow for total COD, dissolved COD, and easily degradable COD. The influent load of the new line was different from the old one, as wastewater from paper production was dumped into the new line. The influent COD was fractioned as follows: 61.5% XS, 10% XI, 16% XH, 3% SI. Measured easily degradable COD was 10% of total COD after pretreatment in the trickling filters. Data were achieved by respiration measurements in batch reactors. In the activated
Old line
TF Denitrification reactor
Nitrification reactor
SST Outflow
Bypass RS
Inflow
SS
TF New line
Denitrification reactor
Nitrification reactor
SST Outflow
TF
RS
SS
Bypass Figure 8 Wastewater treatment plant at Koblenz. TF, trickling filter; SST, secondary settling tank; RS, return sludge; SS, surplus sludge.
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Modeling of Biological Systems
sludge tank total suspended solids (TSSs) and the effluent concentrations of ammonia and nitrate were measured. To consider the plug-flow character of the systems, measured data of oxygen at different points were modeled. Sludge retention times were 11.6 and 10.4 days, respectively. Phosphorus removal was not modeled for the WWTP at Koblenz. Chemical precipitation was only integrated in the simulation as precipitation sludge to consider its effect on sludge age. Precipitated sludge was considered as inert fraction of influent TSS. Because of the plug-flow character of the system (e.g., 1:8 width:length at the new line), the plant was modeled with several tanks in line. Measured oxygen data were available at six points of the old line and four points of the new line. The average oxygen concentrations were quite high (3.2 gO2 m3), especially in the old line. Measured oxygen concentrations are used in the simulation by reading the data from files and comparing and adapting the oxygen values in the simulation with the measured data by proportional-integral-derivative (PID) controllers. The results of ammonia nitrogen effluent simulation are depicted in Figure 9. During the calibration, the nitrification capacity was raised, as with the standard parameters of ASM 3 no sufficient nitrification could be realized. Therefore, the maximum autotrophic growth rate mN was increased to 1.5 d1. Average ammonia nitrogen concentrations in the effluent of the aerated tanks were simulated well. The dynamic behavior shows sufficient results. Modeling of denitrification was easier at the WWTP Koblenz. Despite a low BOD/TKN inflow ratio of 3, low average nitrate nitrogen effluent concentrations of 6 g m3 could be modelled well (TKN, total Kjeldahl nitrogen). The ratio of anoxic heterotrophic decay to aerobic decay was 50%
4.0
4.08.3.3 WWTP at Hildesheim The WWTP at Hildesheim was equipped with a biological stage that allows for nitrification, denitrification, and EBPR. The plant is operated according to the ISAH process. The main asset is that the return sludge is denitrified in a separate anoxic tank in order to prevent any possible impairing of phosphate release in the anaerobic tank through recirculated nitrate. Of four lines planned for the activated sludge plant, two were built and started in July 1987 (Figure 11). The other two lines have both been running since 1997. Operational experiences from the old lines were taken into consideration for the upgrade of the plant. The WWTP at Hildesheim is operated with simultaneous denitrification. The circulation ditch is modeled with two to three layers representing different substrate concentrations over the depth (Figure 12). The circulation ditch is 2 m deep while the upper layer is 71 cm. The lower layer was devided in two layers (65 cm each). One cannot assume that the installed surface aerators (rotors) accomplish complete mixing of the wastewater over the complete depth. In contrast, strong oxygen gradients can be measured. Data showed that the ratio of velocities of the upper and lower part of the tank was 1.8. Modeling of three layers also allows for simulation of anaerobic zones near the bottom of the tank. Oxygen calibration was done based on oxygen sensors in a depth of 30 cm below surface. For simulation of the plant at Hildesheim, measured online-data of total COD, filtrated COD, TKN, and ammonia nitrogen were used.
NH4−Neff, AST, sim
25
NH4−Neff, AST, meas. old line
NO3−Neff, AST, sim
20
NO3−Neff, AST, meas. old line
NO3−N (mg I−1)
NH4−N (mg I−1)
5.0
(ZNO3,end,H ¼ 0.5). This value was recommended by Koch et al. (2000). The results are shown in Figure 10.
3.0 2.0 1.0 0.0 1.0
2.0
3.0 4.0 Time (days)
5.0
6.0
10 5 0
7.0
0
5.0
NH4−Neff, AST, sim
25
4.0
NH4−Neff, AST, meas. new line
20
NO3−N (mg I−1)
NH4−N (mg I−1)
0.0
15
3.0 2.0 1.0 0.0
1
2
3 4 Time (days)
5
6
7
NO3−Neff, AST, sim NO3−Neff, AST, meas. new line
15 10 5 0
0.0
1.0
2.0
3.0 4.0 Time (days)
5.0
6.0
7.0
Figure 9 Simulation results compared with measured data of ammonia nitrogen (g m3), old line (left), new line (right). Gray line: simulated values, black line: measured data.
0
1
2
3 4 Time (days)
5
6
7
Figure 10 Simulation results compared with measured data of nitrate nitrogen (g m3), old line (left), new line (right). Gray line: simulated values, black line: measured data.
Modeling of Biological Systems
Inflow
SST
Denitrification reactor Nitrification reactor
Anaerobic reactor
239
Outflow
Surplus sludge
DNRS reactor
Figure 11 Flow sheet of the WWTP at Hildesheim. SST, secondary settling tank; DNRS, denitrification–recirculation sludge.
D
A
D
A
A
A/D
A/D
D/P
D/P
Inflow
Outflow A/P D
D
Internal recirculation Figure 12 Simplified scheme of model for the recirculation ditch with different layers at the WWTP at Hildesheim (D: mainly anoxic, A: mainly aerobic, P: mainly anaerobic).
Respirations tests revealed a readily degradable COD to be 16% of total COD. Additional online data were available in the effluent for NH4–N, NO3–N, and PO4–P. For the calibration of the TSS in the aerated tank, we used a measured factor of 1:1 gCOD gTS 1 . Simulation results (ASM 3 with EAWAG BioP module) of nitrogen species showed excellent correspondence with measured data, as shown in Figure 13. The influent fractioning in % of total COD was as follows: 49% XS, 15% XI, 9% XH, 11% SI and 16% SS. The results are presented in Figures 13 and 14. Nitrogen effluent values at Hildesheim were simulated without further calibration of biological parameters. Nitrification was increased only slightly by setting the maximum autotrophic growth rate to 1.1 d1. The half-saturation coefficient of heterotrophic biomass for oxygen was calibrated to 0.5 gO2 m3, which resulted in better denitrification. Figure 14 shows the simulated and measured PO4–P data. As can be seen, the PO4–P peak was modeled very well. The polyphosphate storage rate qPP was increased from 1.5 to 2.3 d1 and the maximum storage of polyphosphate Kmax was changed from 0.2 to 0.25 g m3.
4.08.3.4 WWTP at Duderstadt The WWTP at Duderstadt was designed for intermittent denitrification with an integrated pre-anaerobic volume in a
round aerated tank (Figure 15). Enhanced biological phosphorus removal could be established. The plant is operated with a high sludge retention time of about 25 days and without a primary settling tank. Thus, higher concentrations of particulate components in the influent of the AST can be expected. The influent fractioning in % of total COD was as follows: 63% XS, 10% XI, 14% XH, 3% SI and 20% SS. The intermittently aerated tanks are operated through oxygen and ammonia control. Influent concentrations of TSS calculated by ASM 3 were increased by an additional inert fraction to model measured TSS data. For Duderstadt, nitrification and phosphorus removal was increased again. The highly dynamic situation in the effluent of the activated sludge tank is quite difficult to model. The saturation coefficient of nitrogen for autotrophic biomass KNH,N was set to the value of 0.5 gN m3 to model the effluent ammonia nitrogen concentrations. As can be seen in Figure 16 (left), the dynamic simulated curve of ammonia nitrogen still did not always reach minimum effluent ammonia nitrogen concentrations of 0 mg l1. However, the simulation was improved considerably compared with results achieved when modeling biomass decay is independent of the redox milieu. Furthermore, the maximum autotrophic growth was changed to a typical value of 1.4 d1. The NO3–N and PO4–P effluent concentrations showed good results, as can be seen in Figure 16 (right). The
240
Modeling of Biological Systems
NH4−N (g m−3)
20.0
0.0
1.0
2.0
3.0 4.0 Time (days)
5.0
6.0
denitrification capacity was increased with the saturation coefficient of oxygen for the heterotrophic organisms KO,H ¼ 0.5 g m3. At Duderstadt, the EBPR was modeled with a polyphosphate storage rate qPP increased from 1.5 to 1.7 d1. Remarkable, however, is the PO4–P peak on three days of simulation time. Differently than measured, the model showed increased concentrations in the effluent of the secondary settling tank. PO4–P effluent concentrations are found in situations with high phosphorus inflow or in cases with suddenly high COD input. This can result in a high PO4–P release in the anaerobic tank without phosphorus being taken up that fast in the aerobic and anoxic zones. Yet, data analysis showed that here discrepancies can be found because of a rainwater event with approx. 80% increased inflow in between day 2.5 and 3. Further simulations showed that measuring data in the case of rainy weather could be modeled better with a PHA uptake rate of qPHA ¼ 12 d1. Calibrated in this way, greater discrepancies in dry weather situation occurred. Thus, in the end the PHA uptake rate was not changed.
0.0
1.0
2.0
3.0 4.0 Time (days)
5.0
6.0
4.08.3.5 Calibrated Biochemical Parameters and COD Influent Fractionation
15.0 10.0 5.0 0.0
NH3−N (g m−3)
25.0 20.0 15.0 10.0 5.0 0.0
Figure 13 Simulation results compared with measured data of NH4–N (left) and NO3–N (right) (g m3) for a representative load at the WWTP at Hildesheim. Gray line: simulated values; black line: measured data.
PO4-P [g m−3]
10.0 8.0 6.0 4.0 2.0 0.0 0.0
1.0
2.0
3.0 4.0 Time [d]
5.0
6.0
Figure 14 Simulation results compared with measured data of PO4–P (g m3) for a representative load at the WWTP at Hildesheim. Gray line: simulated values; black line: measured data.
In Table 2, the values of the kinetic and stoichiometric parameters that were adapted to simulate the measured data of the plants at Koblenz, Hildesheim, and Duderstadt are presented. The basis for these parameters are publications by Koch et al. (2000) for ASM 3 and Rieger et al. (2001) for the EAWAG BioP Module. It is noteworthy that for German municipal wastewater, the nitrification, denitrification, and phosphorus removal values in the model had to be increased slightly. The maximum autotrophic growth rate was mN ¼ 1.0–1.7 d1, that is, within typical values recommended by Koch et al. (2000) and Rieger et al. (2001). These authors assume that these values result from biofilm growth in activated sludge tanks. Koch et al. (2000) believe that higher CO2 stripping with increased pH is the reason for the calibrated maximum autotrophic growth rates. In few cases in literature, reduced half-saturation coefficients of oxygen KO,N are described to model better nitrification. Wentzel and Ekama (1995) recommend KO,N ¼ 0.02 g m3 for the simulation of activated sludge with
5.0
NH4−N [mg I−1]
4.0 3.0 2.0 1.0 0.0 0.0
1.0
2.0
3.0
4.0 5.0 Time (days)
6.0
7.0
8.0
9.0
Figure 15 Simulation results compared with measured data in the effluent of the AST for NH4–N (g m3) at the WWTP at Duderstadt. Gray line: simulated values; black line: measured data.
Modeling of Biological Systems
ASM 2 (Henze et al., 1995). In the STOWA report from the Netherlands, Hulsbeek et al. (2002) give KO,N ¼ 0.4 g m3, as does Seggelke (2002) for the simulation of the pilot plant at Gu¨mmerwald with ASM 2d (Henze et al., 1999). However, in
NH3−N (mg I−1)
5.0 4.0 3.0 2.0 1.0 0.0 0.0
1.0
2.0
3.0
4.0 5.0 6.0 Time (days)
7.0
8.0
9.0
PO4−P (mg I−1)
3.0 2.5 2.0
241
batch experiments the half-saturation coefficient for oxygen could be measured quite well (KO,N ¼ 0.5 g m3). Thus, following the publication of Koch et al. (2000), no change of this value can be recommended here. To model a better flux for autotrophic biomass in the activated sludge plant at Duderstadt (intermittent denitrification), the half-saturation coefficient for ammonia nitrogen was reduced to KNH,N ¼ 0.5 gN m3. Only by lower ammonia nitrogen concentrations could be described in simulation. Discrepancies in literature for KNH,N are significant (Horn and Hempel (1997): KNH,N ¼ 0.5 g m3; Seggelke (2002): KNH,N ¼ 0.1 g m3 with ASM 2d; Makinia et al., (2005): KNH,N ¼ 0.2 g m3 with ASM 3). The value calibrated here seems to be realistic. Furthermore, enhanced biological phosphorus removal was examined. The calibration was done via the storage rate of polyphosphate qPP and the maximum polyphosphate content of the biomass Kmax. The polyphosphate content of PAO biomass is between 0.1 and 0.4 gP g1 COD (e.g., Rieger et al. according to Wentzel and Ekama (1997), (2001), 0.38 gP g1 COD 0.4 gP g1 COD according to Johansson et al. (1996)). Here, calibration of Kmax showed Kmax ¼ 0.20–0.25 gP g1 COD.
1.5 1.0
4.08.4 Soil Filters
0.5
4.08.4.1 Introduction
0.0 0.0
1.0
2.0
3.0
4.0 5.0 6.0 Time (days)
7.0
8.0
9.0
Figure 16 Simulation results compared with measured data in the effluent of the SST for NO3–N (left) and PO4–P (right) (g m3) at the WWTP at Duderstadt. Gray line: simulated values; black line: measured data.
Planted soil filters and wetlands are used to treat municipal and industrial wastewater with low concentration of particulate material (Brix and Arias, 2005; Molle et al., 2005). Soil filters and wetlands are built as different kinds of systems with horizontal and vertical flow, with horizontal-flow filters being used rather for COD and nitrate nitrogen elimination, vertical-flow filters for nitrification. Through mathematical
Table 2 Calibrated biochemical parameters for simulation of the wastewater-treatment plants (Wichern, 2010) at Hildesheim, Duderstadt, Koblenz, and Gu¨mmerwald compared with values published by Gujer et al. (1999), Koch et al. (2000), and Rieger et al. (2001) Parameter
Unit
Gujer
Koch/Rieger
Hildesheim
Duderstadt
Koblenz
Gu¨mmerwald
Nitrification mN KO.N KNH.N
d1 gO2 m3 gN m3
1.0 0.5 1.0
0.9–1.8 0.5 1.0
1.1 0.5 1.0
1.4 0.5 0.5
1.5 0.5 1.0
1.7 0.5 1.0
Denitrification KO.H
gO2 m3
0.2 0.5
0.2 0.33
0.5 0.5
0.5 0.33
0.2 0.5
0.2 0.33
0.2 1.5 0.2
0.25 2.3 0.2
0.2 1.7 0.2
0.03 0.03 0.035 0.010 0.005 0.014
0.03 0.04 0.04 0.010 0.007 0.014
ZNO3,end,H P removal Kmax qPP KPO4.PP
gP gCOD 1 d1 gP m3
N and P content of chemical oxygen demand fractions 0.03 0.03 iNSS gN gCOD 1 0.03 0.03 iNXS gN gCOD 1 0.035 0.035 iNXI gN gCOD 1 0.010 iPXI gP gCOD 1 0.005 iPXS gP gCOD 1 0.014 iPBM gP gCOD 1
0.2 1.2 0.2 0.04 0.04 0.04
0.03 0.03 0.035 0.010 0.005 0.014
242
Modeling of Biological Systems
modeling of many different processes involving wastewater treatment on planted soil filters, the achievement of a more indepth understanding is expected, which might help further improvement of the utilization of these systems. It is possible to reproduce a simplified reality in a sensible way and extend the often insufficient 1D consideration of soil or sand filters. A calibrated model is an efficient tool to optimize the filter operation – taking into account the impact of the sand/gravel mixture, the wastewater feeding, the temperature, and the oxygen input. Another important phenomenon is the filter clogging, which frequently restricts the purification capacity of the filter and also may impair the utilization of the model. A crucial task for sand filter modeling is to describe substrate conversion that occurs through the biofilm which settles predominantly in the upper areas of the soil filters and on the roots of the plants (Langergraber, 2005). An important step to compare biofilm models with different levels of accuracy and level was achieved in Wanner et al. (2006). A wide range of models from 1D to 3D has been tested to describe the highly complex phenomena that occur in biofilms. Yet, researchers have still not agreed on a standard biofilm model. While the identification of biological and biofilm-specific processes such as diffusion, or the attachment or detachment of particles have been done to a large extent during recent years (Wanner and Gujer, 1986; Gujer and Wanner, 1990; Horn and Hempel, 1997; Wanner and Reichert, 1996), the quantification of single processes in natural systems such as soil or sand filters is still very difficult due to the inhomogeneity of the substratum and substrate. Both identification and quantification can be aided by mathematical modeling. Moreover, even though there has been important progress in the modeling of soil filters, for instance, through the analyses of McBride and Tanner (2000), Langergraber (2003), Dittmer et al. (2005), or Henrichs et al. (2007), complex multi-dimensional models are only rarely being used in practice. Despite the complexity of substrate conversion and hydrology, 1D models are still in use for dimensioning and process optimization (Rousseau et al., 2004; Kadlec, 2000). One reason for this is still the poor availability of measuring data, especially for processes occurring in the biofilm.
4.08.4.2 Material and Methods 4.08.4.2.1 Pilot-scale sand filter Experimental data from a pilot-scale plant consisting of a preliminary SBR, sand filter, and a storage tank (Figure 17) were used to verify the model. The first layer of the filter consisted of sand and gravel (U ¼ 2.93, d60 ¼ 0.63 mm, d10 ¼ 0.21 mm, diameter 0.06–3 mm) with a porosity of 35%. The bottom with a height of 10 cm was composed of gravel. The sand filter had a volume of 0.55 m3 and a depth of 70 cm. In the SBR, nitrate was reduced with the influent COD load. If necessary, methanol was dosed additionally to improve denitrification. Nitrification of ammonium occurred in the vertical-flow sand filter that was fed discontinuously with different kinds of wastewaters (Lindenblatt et al., 2007). During the feeding with landfill leachate, the SBR was operated in four cycles, and in five cycles during the feeding with municipal wastewater (see Table 3). Landfill leachate was taken from a dumping ground in Bavaria. Domestic waste was collected in a covered storage ground together with industrial and commercial waste, rubble, polluted soil, and residues from WWTPs. For landfill leachate, a COD reduction of 30% and an ammonium removal of 80% is required. The planted sand filter was capable of treating ammonium peaks of 10 gNH4N m2 d1 and hydraulic loads of up to 200 l m2 d1. To avoid filter clogging, particular material was mainly removed by sedimentation and substrate conversion in the SBR. Furthermore, analyses of the treatment of municipal wastewater stemming from a municipal WWTP were run at the pilot-scale plant. Table 3 summarizes the load cases used for the model calibration.
4.08.4.2.2 Analytical methods The analytical methods employed for TSSs, volatile suspended solids (VSSs), COD, AOX BOD were based on German Standard Methods (DEV, 1981) for the examination of water, wastewater, and sludge. Ammonium, nitrate, nitrite, and COD were measured spectro-photometrically (Dr. Lange ISIS 6000).
4.08.4.2.3 Mathematical model Here, an integrated model simulating biological processes and geometric distribution of the biofilm on the sand grains was used (Wichern et al., 2008a). The model was capable of
Recirculation
Inflow
Sand filter Storage tank Sequence Batch reactor
Figure 17 Schematic view of the pilot plant.
Outflow
Modeling of Biological Systems Table 3
243
Simulated load cases for the sand filter
Load case
qa (mm d1)
NH4–N ðgm 2Filter d 1 Þ (g d1)
COD ðgm 2Filter d 1 Þ (g d1)
NO3–N ðgm 2Filter d 1 Þ (g d1)
1
161
2
127
3
127
4
161
5
161
6
184
7
294
7.4 5.9 5.1 4.0 6.1 4.8 6.4 5.0 5.6 4.4 8.2 6.46 7.9 6.24
11.4 9.0 7.9 6.3 9.4 7.4 10.6 8.3 12.6 10.0 41.1 32.5 73.2 57.8
3.2 2.5 3.2 2.5 5.1 4.0 1.1 0.9 1.6 1.2 0.1 0.07 0.2 0.1
a
Hydraulic surface loading rate. Influent data for cases 1–5 with landfill leachate and cases 6 and 7 with municipal wastewater.
describing substrate degradation, clogging phenomena in the filter and nitrification capacity of the autotrophic metabolism. Biological processes include two different biomass groups, namely heterotrophic and autotrophic nitrogen consumers. A (quasi/discrete) 2D approach was adopted considering horizontal layers of filter in line, each composed of one completely mixed bulk water volume and biofilm on the substratum surface. Thus, it was possible to calculate substrate gradients through both the depth of the biofilm and the depth of the sand filter, an idea that was also used by Horn and Telgmann (2000) when simulating an upflow biofilter. The model was implemented on the AQUASIM software (Reichert, 1998), utilizing one biofilm reactor compartment for each filter layer. For all the simulations, a total of four layers were utilized. A schematic view of the filter configuration in the model is displayed in Figure 18.
used, who calculated biofilters for secondary denitrification. The remaining growth area on the sand grains and the maximum biofilm thickness is mathematically connected to the biological biofilm model. The sand grains in filter are represented by spheres of uniform size. The grains can touch each other at up to eight points. Where the spheres touch, biomass growth on the surface of the spheres is not possible. Furthermore, the available surface for new biomass growth depends on the biofilm thickness of existing biomass. Figure 19 illustrates this relation. Figure 19 describes the background of the developed equations. It is possible to determine the loss of biofilm surface area, ALoss, between two grains of sand (considered as spheres) in relation to the radius r of the single spheres and the thickness of the biofilm (LF):
ALoss ¼ BpLFð2r þ 2LFÞ ðm 2 Þ
4.08.4.2.4 Biological processes In the model, the growth and decay of heterotrophic and autotrophic biomass under aerobic and anoxic conditions were considered. The decay processes are described with endogenous respiration for both organism groups. From the measured data of the sand filter for municipal wastewater, it became apparent that COD effluent values were very low (B20 g m3 CODhom). Hardly any substances were discharged in particulate form. To describe the complete degradation of particular matter, two degradable fractions were defined. XXS means COD which is degraded very slowly resulting from biomass inactivation, SS means the mixed fraction for COD that is degraded via heterotrophic growth (Gujer et al., 1999). The implementation of XXS allows for modeling very slowly degradable compounds at high sludge ages in the filter. Tables 4 and 5 describe the entire biochemical model.
4.08.4.2.5 Biofilm modeling The biofilm properties are a determinant step in the sand filter model development. Here, the basic idea of Horn (1999) was
ð19Þ
with B representing the number of contact points per sphere. The equation is based on the calculation of the surface of spherical segments. In order to be able to estimate the number of grains of sand in the sand filter, the porosity of the sand filter must be known. In the case of hexagonal packing with the tightest possible structure (HCP grid) and a cubic areacentered structure (FCC grid), the porosity is 0.26. In nonidealized pouring cases, however, the porosity is considerably higher. For sand/gravel mixtures, values up to 0.50 have been documented. Thus, the number of grains of sand in a test reactor depending on the porosity e amounts to
N¼
VReactor VPores ð1 eÞ VReactor ¼ ðÞ p 3 VBall d 6 Ball
ð20Þ
with N being the number of sand grains (–), e the porosity (–), and VReactor, VPores and VBall the reactor, free pores, and sand grain volumes (m3), respectively. Considering the number of
244
Modeling of Biological Systems Inflow
17 cm
Inflow Biofilm reactor compartment1 Bulk liquid
Biofilm Biofilm reactor Compartment 2 70 cm
Diffusion
Biofilm growth Biofilm reactor compartment 3
Detachment
Biofilm reactor compartment 4 Outflow Outflow Figure 18 Flow sheet of the model setup for the sand filter.
Table 4
Stoichiometric matrix of the sand filter biofilm model (Wichern et al., 2008a) Process
XH
XN
1
Aerobic heterotrophic growth
1
2
Anoxic heterotrophic growth
1
3 4
Aerobic heterotrophic decay Anoxic heterotrophic decay
1 1
5 6 7 8
Aerobic heterotrophic inactivation Anoxic heterotrophic inactivation Aerobic maintenance Anoxic maintenance
1 1
9
Aerobic autotrophic growth
10 11
Aerobic autotrophic decay Anoxic autotrophic decay
12
Hydrolysis
XXS
SNH
1 YH 1 YH
1 YH 1 IB þ IBS YH IB IB
1 1 1 1 1 1 1
SO
IB þ IBS
IB IBS IB–IBS IBS IBS
1 IB YN
IB IB 1
contact points and the diameter of the spheres allows for calculating the remaining biofilm surface ARemaining:
ARemaining ¼ Npd2Ball NALoss ðm 2 Þ
SS
ð21Þ
Figure 20 makes clear that for the given values and depending on the number of contact points, a complete reduction of the remaining biofilm surface appears already at an
IB þ IBS
SNO
SI
1 YH
1 1 YH 2:9 Y H
1 1 2:9
1 1 2:9 4:6 Y N YN 1
1 YN 1 2:9
1
existing biofilm thickness between 60 and 110 mm. Thus, the possible biofilm thickness available for the growth of the biofilm in the sand filter stays within a relatively narrow range. Hence, it can be concluded that the biological processes which lead to the production of particulate components must be contained in the case of surface limitation. For this restriction, the following Monod-type term was introduced:
MLim ¼
ðLFmax LFÞ ðÞ ðLFmax LFÞ þ KLF
ð22Þ
Modeling of Biological Systems Table 5
Kinetic matrix of the sand filter biofilm model Process
Process rate (T0 ¼ 20 1C)
1
Aerobic heterotrophic growth
mH
2
Anoxic heterotrophic growth
mH ZH
3
Aerobic heterotrophic decay
bH
4
Anoxic heterotrophic decay
bH ZD
5
Aerobic heterotrophic inactivation
bH;Inakt
6
Anoxic heterotrophic inactivation
bH;Inakt ZD
7
Aerobic maintenance
mH
8
245
Anoxic maintenance
SS SO LFmax LF XH K S þ SS K O þ SO ðLFmax LFÞ þ KLF SS K OH SNO LFmax LF XH K S þ SS K OH þ SO SNO þ K NO ðLFmax LFÞ þ KLF
SO XH SO þ K O K OH SNO XH SO þ K OH SNO þ K NO SO LFmax LF XH SO þ K O ðLFmax LFÞ þ KLF K OH SNO LFmax LF XH SO þ K OH SNO þ K NO ðLFmax LFÞ þ KLF SS Main
mH
Main
KS ZH
Main
þ SS K O SS
KS
Main
SO XH þ SO
Main
K OH SNO þ SS K OH þ SO SNO þ K NO
XH Main
SNH SO2 LFmax LF XN K N þ SNH K OA þ SO ðLFmax LFÞ þ KLF
9
Aerobic autotrophic growth
mN
10
Aerobic autotrophic decay
bN
11
Anoxic autotrophic decay
bN ZD
12
Hydrolysis
SO XN SO þ K O
K OH SNO XN SO þ K OH SNO þ K NO X XS kH K X þ X XS =X H
Sand grain Biofilm Missing area percontact point
LF
r
Bulk liquid Figure 19 Idealized presentation of the sand body using spheres with the same diameter (left), missing growth area per contact points between two spheres (right).
where MLim is the surface limitation term to the growth rate (–), LFmax the maximal biofilm thickness (m), and KLF the half-saturation coefficient limiting the thickness of the biofilm growth (m). To avoid filter clogging, substrate conversion and particle detachment need to be included into the model. Biofilm obstruction – mainly in the two upper layers of the
model – is expected particularly at higher substrate concentrations. At the bottom of the filter, no detachment takes place, as the biofilm thickness is below the base thickness of 50 mm. Biofilm detachment has been a research focus for several years now. Different models – among others from Trulear and Characklis (1982), Wanner and Gujer (1986), and
246
Modeling of Biological Systems
Remaining area (m2)
4000
B=8 B=7 B=6 B=5 B=4
3000 2000 1000 0 0
25
50
75
100
125
150
Biofilm thickness, LF (µm) Figure 20 Dependence of the remaining surface for biofilm growth in relation to the existing biofilm thickness for e ¼ 0.35, dsphere ¼ 0.6 mm, and vreactor ¼ 0.55 m3.
Kreikenbohm and Stephan (1985) – have translated microbial findings into mathematical equations. Here, a detachment rate based on the findings of Wanner and Gujer (1986), Morgenroth and Wilderer (2000), and Horn et al. (2003) was developed further:
dLF ¼ kD rF ðLF LFBase Þ3 dt
ðm d1 Þ
ð23Þ
where kD is the detachment rate of the biofilm (m g1 d1), rF the biofilm density (g m3), and LFBase the minimal biofilm thickness (m). This equation allowed for a decent reproduction of the detachment processes in the sand filter. As a result from model calibration, a basic biofilm thickness of LFBase ¼ 50 mm was assumed. Thinner biofilms are not subject of biomass detachment. Oxygen input into the sand filter with vertical flow was calculated with the equations provided by Platzer (1999), who assumed a value of 1.0 gO2 h1 m2 for the oxygen input through diffusion. In the present simulations, this value was reduced slightly to 0.9 gO2 h1 m2. For air saturation, an oxygen flux of 0.4–0.5 gO2 h1 m2 was reported by Casey et al. (1999) for membrane-aerated biofilms, which indeed are less porous.
4.08.4.3 Results and Discussion
oxygen inhibition constant (KO,H) was increased from 0:5to2:0 gO2 m3 , which results in a better denitrification capacity, especially for low COD inflow concentrations. The diffusion coefficients used for nitrogen components are typical of biofilms (see Table 6) and have been used, among others, by Horn (2003) and Rauch et al. (1999). Wanner and Reichert (1996) assumed a value of DSs ¼ 1.0 104 m2 d1. Polprasert et al. (1998) documented a diffusion coefficient for the COD in planted soil filters of DSs ¼ 2.2 104 m2 d1. For oxygen, Horn (2003) documented a value of DO2 ¼ 2:2 10 4 m 2 d 1 ; Wanner and Reichert (1996) gave a value of DO2 ¼ 1:0 104 m2 d1 . Table 7 compiles the influent data and measurement data for the investigated system. Deriving from scenario calculations, further results are depicted to predict the best sand filter operation. The model was well capable of reproducing the different load cases. Conversion of ammonium nitrogen and COD is simulated for landfill leachate, landfill leachate with additional dosing of methanol into the SBR (5–20 ml d1 pure methanol) and municipal wastewater. Ammonia nitrogen effluent concentrations are consistent with results published by Cooper (2005) for wetlands with vertical flow. The simulated degradation of substrates over the depth of the planted sand filter is presented in Figure 21. The concentration of nitrate nitrogen rises necessarily with the degradation of ammonium nitrogen. Moreover, model calculations show that in municipal wastewater the nitrification takes place more slowly and in deeper layers of the filter. Furthermore, nitrification started after COD degradation had already progressed to a high degree. For the treatment of landfill leachate, the ratio between degradable COD and hardly degradable COD is much lower and the latter must be hydrolyzed first; thus, nitrification takes place much earlier there.
4.08.4.3.2 Sensitivity analysis A sensitivity analysis (SA) was conducted to verify the importance and influence of biochemical parameters. Furthermore, the influence of the number of contact points of the single sand grains was investigated. For the biochemical parameters, the SA procedure is based on the publication of Kim et al. (2006), who used the single parameter method (SVM) slope technique to investigate activated sludge models. The effluent quality index (EQ) is defined as
4.08.4.3.1 Model calibration and simulation results Table 6 summarizes the biochemical parameters used in the model. The calibrated biochemical parameters are based on the standard parameter set developed for municipal wastewater for the ASM 3 (Koch et al., 2000; Wichern et al., 2002). The hydrolysis rate kH describes the supply of very slowly degradable COD XXS for biological conversion at high sludge ages that usually do not occur in flocculent activated sludge systems. Thus, the hydrolysis rate kH was calibrated significantly lower (kH ¼ 2.0 d1). To describe the flux and degradation of ammonium, the autotrophic growth rate was calibrated to 1.8 d1 – which is at the upper end of typical ASM 3 values. As the soil filter in reality is not as homogeneous as the geometric model assumes, there are still nitrate-reducing zones in the sand filter that are not penetrated by oxygen at all. To consider this on the model scale, the
EQ ¼ bCOD CODe þ bNH4 NH4 Ne þ bNO3 NO3 Ne ðÞ ð24Þ The values 1, 10, and 1 for the weighting factors b of COD, NH4–N and NO3–N were chosen to reflect the importance of ammonium nitrogen removal in the planted sand filter. The sensitivity index DEQ is calculated as follows (Ref: referring to the calibrated reference simulation, Var: referring to the varied parameter).
DEQ ¼ bCOD CODe;Ref CODe;Var þ bNH4 NH4 Ne;Ref NH4 Ne;Var þ bNO3 NO3 Ne;Ref NO3 Ne;Var ðÞ
ð25Þ
For the SA, 10 model parameters were changed from 50% to 200% of their original values in 10% steps, which resulted in
Modeling of Biological Systems
247
Table 6 Stoichiometric and kinetic parameters after calibration (T ¼ 20 1C) compared to the values of the ASM 3: Koch et al. (2000) and Wichern et al. (2002) Parameter
Unit
Heterotrophic biomass mH d1
Sand filter
3 0.07 35 0.07 0.50a 0.07 0.75 0.07 1b 1c 0.5 0.5c 2.0
Koch/Wichern
3 0.07
mH_Main
d1
bH
d1
bH,Inact
d1
KS KS_Main KO KO_Main KOH
g g g g g
KNO KNO_Main
g m3 g m3
0.5 0.5
0.5
ZD ZH YH kH kX
g g1 d1 g m3
0.5 0.5 0.67 2 1
0.5/0.33 0.5 0.64d
Autotrophic biomass mN
d1
bN
d1
YN KN KOA
g g1 g m3 g m3
1.8 0.105 0.15 0.105 0.24 0.5e 0.4f
0.9–1.7 0.105 0.15 0.105 0.24 1.0 0.5/0.13
Other parameters IBS IB kD KLF DNH DNO DO2 DSs d
g g1 g g1 m2 d1 m m2 d1 m2 d1 m2 d1 m2 d1 g m3
m3 m3 m3 m3 m3
0.30 0.07
10 0.5 0.5
0.03 0.03 0.07 0.07 466 1.0 106 1.8 104 1.8 104 1.7 104 2.2 104 60 000
Description
Maximum heterotrophic growth rate Temperature factor Maximum heterotrophic rate of the sustenance metabolism Temperature factor Maximum heterotrophic endogenous decay rate Temperature factor Maximum heterotrophic inactivation rate Temperature factor Half-saturation concentration for COD Half-saturation concentration for COD (maintenance metabolism) Half-saturation concentration for oxygen Half-saturation concentration for oxygen (maintenance metabolism) Half-saturation concentration for oxygen (inhibition of the denitrification) Half-saturation concentration for nitrate nitrogen Half-saturation concentration for nitrate nitrogen (maintenance metabolism) Reduction factor for the anoxic heterotrophic decay process Reduction factor for the anoxic heterotrophic growth process Heterotrophic yield Maximum hydrolysis rate of the very slowly degradable COD Half-saturation concentration during the hydrolysis process Maximum autotrophic growth rate Temperature factor Maximum endogenous autotrophic decay rate Temperature factor Autotrophic yield Half-saturation concentration for ammonium nitrogen Half-saturation concentration for oxygen Nitrogen incorporated into all COD fractions except for the biomass Nitrogen incorporated into the biomass Detachment rate of the biofilm (detachment) Half-saturation coefficient limiting the thickness of the biofilm growth Diffusion coefficient for ammonium nitrogen Diffusion coefficient for nitrate nitrogen Diffusion coefficient for oxygen Diffusion coefficient for dissolved COD Biofilm density
a
Hulsbeek et al. (2002): bH ¼ 0.05–1.6 d1 for the ASM 1. Wanner and Reichert (1996) as well as Morgenroth and Wilderer (2000): KS ¼ 5 g m3, KS ¼ 2.5 g m3. c Horn et al. (2003): KS,Main ¼ 1 g m3, KO2,Main ¼ 0.2 g m3. d Results from the multiplication of the yield rates for substrate storage and heterotrophic growth. e Horn and Hempel (1997), KN ¼ 0.5 g m3; Makinia et al. (2005), KN ¼ 0.2 g m3, ASM 3. f Morgenroth and Wilderer (2000): KOA ¼ 0.1 g m3. COD, chemical oxygen demand. b
150 simulation runs. The sensitivity of the maximum heterotrophic growth rate can be seen in Figure 22. The nitrification capacity is influenced by oxygen concentrations that were affected by heterotrophic COD conversion. An increased diffusion coefficient for oxygen resulted in better COD removal, especially in the first 10 cm of the sand filter. Nitrification capacity was partly lost because of elevated COD removal.
Furthermore, especially with a lower autotrophic growth rate there is also a decrease in nitrification. A higher autotrophic yield leads to a significantly increased autotrophic biomass, and, due to the competition for oxygen, to a decreased COD removal. In general, it can be observed that the diagram does not change significantly when the COD weighting factor bCOD is increased to 10. The maintenance
248
Modeling of Biological Systems
Table 7
Selected results and corresponding measuring data for seven load cases with and without methanol dosage into the SBR
Case/scenario
Case 1 Case 2 Case 3 Case 4 (methanol) Case 5 (methanol) Case 6 Case 7 Scenario 1 Scenario 2 Scenario 3 Scenario 4 Scenario 5
Influent sand filter
Effluent sand filter
Simulation
COD/NH4–N measured (g m3)
NH4–N- surface load 1 (g m2 reactor d )
CODhom Meas/ Sim (g m3)
NH4–N Meas/ Sim (g m3)
NO3–N Meas/ Sim (g m3)
NH4–N degradation 1 (g m2 reactor d )
71.0/46.4 62.6/40.2 74.2/48.2 66.0/39.8 79.0/35.0 225.1/44.7 250.2/27.0 164.3/32.7 193.6/38.5 247.6/49.2 274.6/54.6 330.9/65.8
7.4 5.1 6.1 6.4 5.6 8.2 7.9 6.0 7.0 9.0 10.0 12.0
60.6/57.8 50.6/50.4 60.0/60.1 53.9/52.8 61.7/61.3 17.3/18.6 19.7/19.1 –/12.5 –/15.1 –/19.6 –/19.8 –/20.3
0.1/0.1 0.0/0.1 0.3/0.1 2.4/0.1 1.4/0.1 o1.0/0.25 0.8/0.31 –/0 –/0.1 –/6.7 –/17.3 –/41.0
62.0/66.5 59.5/65.9 81.6/88.3 44.0/46.7 38.3/45.1 27.1/31.2 21.9/21.1 –/36.5 –/39.6 –/24.6 –/18.1 –/5.3
7.4 5.1 6.1 6.4 5.6 8.1 7.8 6.0 7.0 7.8 6.8 4.5
Results for COD, NH4–N, and NO3–N (left: measuring; right: simulation); load cases 1–5: landfill leachate; load cases 6 and 7: municipal wastewater. Moreover, influent loads and degradation performance of the filter in different simulation scenarios with municipal wastewater.
COD (mg l−1)
300 250 Case 6 (municipal wastewater)
200 150
Case 1-5 (landfill leakage)
100
Case 7 (muncipal wastewater)
50 0 0
10
20
30 40 50 Depth (cm)
60
70
80
NH4−N (mg l−1)
60 50
Case 6 (municipal wastewater)
40 30
Case 1-5 (landfill leakage)
20 Case 7 (muncipal wastewater)
10 0 0
10
20
30 40 50 Depth (cm)
60
70
80
Figure 21 COD (left) and NH4–N (right) concentrations (mg l1) for all seven load cases related to the depth of the sand filter (load cases 1–5: landfill leachate; load cases 6–7: municipal wastewater).
process of heterotrophic biomass and the autotrophic yield are getting more sensitive with higher bCOD. Afterward, the effect of the number of contact points of the single grains was analyzed. The contact points were changed from a minimum of 4 to a maximum of 7. This variation directly affected the remaining area for the biofilm growth. Figure 23 shows the effects of the number of contact points on the substrate removal for ammonium and COD.
As expected, the COD and ammonium concentrations in the bulk liquid decrease over the filter depth. With the number of contact points higher than 6, substrate conversion and removal efficiency decreased extensively. The available area for biofilm growth limits the biomass growth. If the pore volume is filled by existing biomass, new bacteria can only grow when biomass detachment has occurred or existing biomass has been inactivated and hydrolyzed. Substrate conversion is considerably higher if new biomass grows. In Figure 24 one can see that according to model results due to limited growth area, an increased number of contact points leads to thicker biofilms in the examined depths (8.75, 26.25, 43.75, and 61.25 cm). Biofilm thickness increases for 4 to 6 contact points because for nearly constant substrate degradation the same quantity of biomass is necessary. In the case of 7 contact points, the pore volume is not sufficient to maintain substrate degradation. Furthermore, Figure 25 reveals that biofilm thickness in the upper layers of the sand filter is higher than in the lower ones. Apart from the impact on biofilm thickness, the biomass composition also changes when contact points are varied. Figure 25 shows the effect of different numbers of contact points on the biomass composition of the sand filter in depths of 8.75 and 26.25 cm. Depicted are the ratios of autotrophic biomass XA per total biomass and autotrophic XA plus heterotrophic biomass XH per total biomass. When contact points are varied between B ¼ 4–6, there has hardly any effect on biomass composition, not until there are 7 contact points. Mathematical modeling and microbial analyses showed higher quantities of autotrophic biomass than one would expect for municipal wastewater treatment. This happens because COD was mainly removed in the upstream SBR that worked as sedimentation reactor and denitrification tank. In the investigated system, the sand filter serves mainly for nitrification. Compared to municipal wastewater, the COD/NH4– N ratio (15:1) in the filter influent is considerably lower (5:1).
Modeling of Biological Systems
H H_main
125
Sensitivity index dEQ (−)
249
kOH kH N
100
75
50
25
0 60
80
100
120
140
160
180
200
Parameter variation (%) Figure 22 Results from sensitivity analysis for exemplary model parameters based on 150 simulation runs.
COD [g m−3] for B = 4 COD [g m−3] for cal. B = 5 COD [g m−3] for B = 6 COD [g m−3] for B = 7 COD [g m−3] for B = 6,5
COD (mg l−1)
250 200 150 100 50 0 0
10
20
60
70
80
NH4-N [g m−3] for B = 4 NH4-N [g m−3] for cal. B = 5 NH4-N [g m−3] for B = 6 NH4-N [g m−3] for B = 7 NH4-N [g m−3] for B = 6,5
250 NH4-N (mg l−1)
30 40 50 Depth (cm)
200 150 100 50 0 0
10
20
30 40 50 Depth (cm)
60
70
80
Figure 23 Effect of the number of contact points of the single grains of sand on the effluent concentrations of COD (left) and NH4–N (right). COD, chemical oxygen demand.
4.08.5 Waste Stabilization Ponds 4.08.5.1 Introduction Waste stabilization ponds (WSPs) are a very appropriate method of wastewater treatment in developing countries, where the climate is most favorable for this application. Their lower implementation costs and operational simplicity are commonly regarded as their main advantages. However, the processes that occur in wastewater-treatment ponds still are not completely understood. Environmental factors such as sun
radiation, wind, biological processes, and hydrodynamics have as yet not been fully analyzed or are difficult to validate with experimental data. The design of WSPs is mostly based on empirical equations (Pano and Middlebrooks, 1982; von Sperling and Chernicharo, 2005), and there are only a few published mathematical models to simulate the dynamics of such complex systems (e.g., Buhr and Miller, 1983). Juspin et al. (2003) developed an adaptation from the River Water Quality Model No.1 (Reichert et al., 2001) to simulate highrate algae ponds, including the influence of daily light variations. Here, ASM 3 (Gujer et al., 1999) is presented for the modeling of facultative and maturation ponds, with extensions to consider important processes relevant for WSPs: algae photosynthesis, growth and endogenous respiration, gas exchange for oxygen, carbon dioxide, and ammonia, ionic equilibrium processes, and pH calculation processes (Gehring et al., 2010). Solar radiation and wind velocity were implemented as model parameters. Solar radiation is the principal factor influencing algae growth and pathogen removal (which is not evaluated in the model). Although algae photosynthesis is identified as one of the main sources of oxygen input in facultative WSPs, superficial oxygen transfer has been identified as a relevant factor under conditions with high wind velocities (Ro et al., 2006; Pelletier and Chapra, 2008). Up to now, however, wind impact is rarely quantified and included into mathematical models for ponds. Wind can also play an important role in nitrogen removal. Smith and Arab (1988) identified turbulence as a main factor in the liquid–atmosphere interface for free ammonia desorption. Ni (1999) also verified the high importance of air velocity to determine the mass transfer coefficient. The authors analyzed 30 mechanistic models regarding ammonia release in cases of manure treatment. Significant ammonia desorption rates from WSPs were also related in Shilton (1996) and Rumburg et al. (2008). Biological ammonia nitrogen removal through nitrification followed by denitrification is another possible path, as published by Hurse and Connor (1999) and Zimmo et al. (2003).
250
Modeling of Biological Systems 80
B = 4−6 depth 8.75 cm B=7
Biofilm thickness (µm)
60
B = 4−6 depth 26.25 cm B=6 B = 4−6 depth 43.75 cm
40
B=5 B=4
20 B=6
B = 4−6 depth 61.25 cm
B=5 B=4
0 0
5000
10 000
15 000
20 000
Time (min) Figure 24 Effect of the number of contact points of the single grains on biofilm thickness.
1.0 0.8
Ratio of XA+XH, depth 8.75 cm
0.6 0.4
Ratio of XA, depth 8.75 cm
0.2 0.0 0.00 0.01
1.2 Biomass fractions (−)
0.03
0.04
0.05
1.0 Ratio of XA+XH depth 8.75 cm
0.8
Ratio of XA, depth 8.75 cm
0.6 0.4 0.2 0.0 0.00
0.01
0.02
0.03
0.04
0.05
0.06
0.6 Ratio of XA, depth 26.25 cm
0.4 0.2 0.0 0.00 0.01
0.02
0.03
0.04
0.05
0.06 0.07
Depth of the biofilm (mm) 1.2
Ratio of XA+XH + XXS, depth 26.25 cm
1.0 Ratio of XA+XH, depth 26.25 cm
0.8
Ratio of XA, depth 26.25 cm
0.6 0.4 0.2 0.0 0.00 0.01
0.07
Depth of the biofilm (mm)
Ratio of XA+XH, depth 26.25 cm
0.8
(b)
Ratio of XA+XH + XH, depth 8.75 cm
Ratio of XA+XH + XXS, depth 26.25 cm
1.0
0.06 0.07
Depth of the biofilm (mm)
(a)
(c)
0.02
1.2 Biomass fractions (−)
Ratio of XA+XH+XXS, depth 8.75 cm
Biomass fractions (−)
Biomass fractions (−)
1.2
(d)
0.02
0.03
0.04
0.05
0.06 0.07
Depth of the biofilm (mm)
Figure 25 Effect of the number of contact points b of the single grains of sand for: (a) B ¼ 4, depth 8.75 cm; (b) B ¼ 4, depth 26.25 cm; (c) B ¼ 7, depth 8.75 cm; and (d) B ¼ 7, depth 26.25 cm. From Wichern M, Lindenblatt C, Lu¨bken M, and Horn H (2008a) Experimental results and mathematical modelling of an autotrophic and heterotrophic biofilm in a sand filter treating landfill leachate and municipal wastewater. Water Research 42: 3899– 3909.
4.08.5.2 Material and Methods 4.08.5.2.1 Description of the pilot pond For the model calculation, experimental data from a pilot pond system gathered in the city of Floriano´polis in southern Brazil (Santa Catarina) were used. The system consisted of
three ponds in line (anaerobic–facultative–maturation), fed with leachate from an approximately 15-year-old municipal waste landfill (Silva, 2007). The results of the dynamic simulation presented here refer to the facultative (Fac) and maturation (Mat) ponds. Each pond had a volume of 1.1 m3, a
Modeling of Biological Systems surface area of 1.2 m2, and a depth of 1.0 m. The main characteristics of the influent and effluent of both ponds are summarized in Table 8. Wind was measured at an automatic hydrological station. Sun radiation measurements were collected in a solar station of the BSRN/ (271380 S, 481300 O; BSRN – Baseline Surface Radiation Network/WMO – World Meteorological Organization project).
Table 8
Mean effluent parameters of the three ponds
Parameter
Unit
Anaerobic Facultative Maturation
COD COD filtered COD/BOD5 Total suspended solids Total ammonia nitrogen Nitrate nitrogen Chlorophyll a Dissolved oxygen pH Temperature
g m3 1456 g m3 n.m. 6.8 g m3 301 g m3 505 g m3 8 mg l1 n.m. g m3 n.m. 8.7 1C 25.9
1233 1065 7.3 239 208 6 65.3 3.7 8.8 25.6
743 504 7.3 146 53 4 93.4 3.9 8.9 25.2
n.m., parameter not measured; BOD, biochemical oxygen demand; COD, chemical oxygen demand. Adapted from Gehring T, Silva J, Kehl O, et al. (2010) Modelling waste stabilization ponds with an extended version of ASM 3. Water Science and Technology 61(3): 713–720.
Table 9
251
For better characterization of the ponds, some parameters such as chlorophyll a, oxygen, pH, and temperature were measured at three different depths: 0.2, 0.5, and 0.8 m. They are referred here to as top, middle, and bottom layers, respectively. Average values in the different layers are depicted in Table 9.
4.08.5.2.2 Mathematical model Additional processes extending ASM 3 for simulations of facultative and maturations ponds were taken as proposed in Gehring et al. (2010), in order to describe the influences of sun and wind, algae biomass, gas exchange, and ionic equilibrium. The model implementation was done in Simba 4.2 (2005).
4.08.5.2.3 Hydraulic concept In order to understand the hydraulic behavior in a pond, it is crucial to describe the substrate flux in detail. Each pond was assumed as a combination of three completely stirred tank reactors (CSTRs). These CSTRs represented the depth of the system, with stratified layers: bottom, middle, and top (schematic view in Figure 26). The flow was constant with 60 l d1, resulting in a hydraulic retention time of 18 days. Sedimentation was neglected in the model, as it was considered irrelevant in the experimental observations.
4.08.5.2.4 Algal processes According to experiments, each gCOD m3 of algae had a concentration of 12 mg l1 chlorophyll a. Two different growth processes, based on ammonia and nitrate, and an endogenous
Mean measurements in three depths of the ponds
Parameters
Unit
Facultative pond
1
Chlorophyll a Dissolved oxygen pH Temperature
mg l g m3 1C
Maturation pond
Top
Middle
Bottom
Top
Middle
Bottom
79 3.9 8.8 26.9
61 3.7 8.8 25.1
57 3.4 8.8 24.8
163 4.5 9.1 25.7
110 4.2 8.9 23.4
101 3.9 8.7 23.0
Adapted from Gehring T, Silva J, Kehl O, et al. (2010) Modelling waste stabilization ponds with an extended version of ASM 3. Water Science and Technology 61(3): 713–720.
Gas exchange
Inflow
Gas exchange
Top layer 0.4 m3
Middle layer 0.35m3
Middle layer 0.35 m3
Bottom layer 0.35 m3
Bottom layer 0.35 m3
Facultative pond
Maturation pond
1m
Top layer 0.4m3
Figure 26 Pond configuration in the model.
Outflow
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Modeling of Biological Systems
respiration process were included in ASM3 in analogy to RWQM No. 1 (Reichert et al., 2001). Process rates are displayed in Table 10. Light attenuation through depth is usually described by Beer’s law, where the attenuation coefficient is considered according to the absorption properties of water. In WSPs, light absorption is mainly determined by the concentrations of gilvin (dissolved yellow matter), algae, and tripton (inanimate particulate matter) (Curtis et al., 1994). Heaven et al. (2005) observed that only a few light attenuation coefficients for WSPs have been published so far and highlighted their importance to mathematical description of algae
Table 10
1
processes. According to Pelletier and Chapra (2008), the photosynthetic available radiation (PAR) was assumed to be 47% of total light radiation. Attenuation across the water depth was calculated from the Beer–Lambert equation to determine the available light radiation IAV. The light attenuation parameter Kd was defined as function of the mixed liquor suspended solids in the tank:
Iav ¼ 0:47 I eKd H Kd ¼ a1 þ a2 XTSS
ðW m2 Þ
ð26Þ
ðm1 Þ
ð27Þ
Kinetic matrix of the waste stabilization ponds model Process
Process rate
Hydrolysis
kH
X XS K X þ X XS =X H
Heterotrophic processes 2
Aerobic storage
3
Anoxic storage
4
Aerobic growth
5
Anoxic growth
6
Aerobic endogenous respiration
7
Anoxic endogenous respiration
8
Aerobic respiration of XSTO
9
Anaerobic respiration of XSTO
SO SS K O þ SO K SS þ SS KO SS SNO k STO ZNO X H K O þ SO K SS þ SS K NO þ SNO SCO2 SO SNH X STO =X H mH X H K O þ So K NH þ SNH K CO2 þ SCO2 K STO þ X STO =X H SCO2 KO SNH X STO =X H SNO mH ZNO X H K O þ SO K NH þ SNH K CO2 þ SCO2 K STO þ X STO =X H K NO þ SNO SO XH bH SO þ K O KO SNO XH bN ZD SO þ K O SNO þ K NO SO bH X SO þ K O STO KO SNO bN ZD X SO þ K O SNO þ K NO STO k STO X H
Autotrophic processes 10
Aerobic growth
11
Aerobic endogenous respiration
12
Anoxic endogenous respiration
SNH SO2 SCO2 XN K N þ SNH K OA þ SO K SCO2 A SCO2 SO XN bN SO þ K O KO SNO XN bN ZD SO þ K O SNO þ K NO
mN
Algae processes 13
Growth with SNH
14
Growth with SNO
15
Aerobic endogenous respiration Physicochemical processes
16
Equilibrium1: HCO3 =CO2
17
Equilibrium 2: NH3 =NH4 þ
18
Gas exchange 1: O2
19
Gas exchange 2: CO2
20
Gas exchange 3: NH3
SNO þ SNH SNH K N;ALG þ SNO þ SNH K NH;ALG þ SNH K NH;ALG SNO þ SNH mALG K N;ALG þ SNO þ SNH K NH;ALG þ SNH SO bALG X ALG SO þ K O;ALG mALG
SH þ SHCO3 kACO2 SCO2 K a CO2 SH þ SNH3 kAIN SNH4 K a IN A klaO2 ðSsat O2 SO2 Þ V A klaCO2 ðSsat CO2 SCO2 Þ V A klaNH3 ðSsat NH3 SNH3 Þ V
IAV exp ð1 ð1=K I ÞÞX ALG KI IAV expð1 ð1=K I ÞÞ X ALG KI
Modeling of Biological Systems with I being the light radiation (W m2), H the depth (m), a1 the light attenuation constant from water color and turbidity (m1), a2 the light attenuation constant factor from suspended solids (m3 g1 m1), and XTSS the total mixed liquor solid concentration (g m3).
4.08.5.2.5 Physico-chemical processes Wett and Rauch (2003) described the importance of the inorganic carbon balance for nitrification in activated sludge systems treating highly concentrated ammonia wastewaters. They found it necessary to include the ionic equilibrium and also the stripping of carbon dioxide in the ASM models to better represent these processes. The carbon dioxide release as a function of pH was found to be of major importance for a correct evaluation of free ammonia concentrations (Ni, 1999). Here, the free ammonia concentration SNH3 ðgm3 Þ and carbon dioxide SCO2 (molC m3) were added into the model as new state variables. Thus, the ionized ammonium concentration SNH4 ðgm3 Þ could be dynamically determined as the difference of SNH (total ammonia) and SNH3 . The concentrations of free ammonia and carbon dioxide in the liquid phase were defined through two equilibrium processes, considering the acid/base pairs: SCO2 =SHCO3 and SNH4 =SNH3 . Both equilibrium equations and the equilibrium constants were set according to Reichert et al. (2001) and are depicted in Tables 10 and 11. Determination of pH was also necessary. It was realized through a charge balance, considering the influence of ionized ammonia, nitrate, and alkalinity. All equivalent charges together with the dissociation products of water, Hþ and OH, amounted to zero. Three different gas transfer processes between top layer and atmosphere were considered to determine the mass transfer of oxygen, free ammonia, and carbon dioxide, as shown in Table 10. The gas transfer rate was determined according to a convective mass transfer, which depends on the concentration gradient between atmosphere and liquid. If
Table 11
253
dissolved concentrations exceed atmospheric concentrations, the rates are negative and gas is released from the liquid. If rates are positive, there occurs gas absorption from the atmosphere. The oxygen transfer coefficient, klaO2, was calculated by the following equation according to Ro and Hunt (2006), who developed an empirical equation based on 297 data points from transfer coefficients published in the last 50 years. This equation was recommended by the authors to be applied to WSP:
klaO2 ¼ 0:24 170:6Sc1=2 U1:81 10
ra rw
1=2
ðm d1 Þ
ð28Þ
where Sc is the dimensionless Schmidt number, U10 is the wind velocity 10 m over the ground (m s1), and patm and pw are the atmosphere and water densities, respectively (kg m3). Saturation of oxygen was defined through an empirical equation, and carbon dioxide and free ammonia saturations as functions of Henry’s constant and the atmosphere pressure of the gas:
Ssat
O2
¼ 13:89 0:3825T þ 0:007311T 2 ðg m3 Þ
ð29Þ
ðmol m3 Þ
ð30Þ
0:00006588T 3 Ssat
gas i
¼
patm gas KH
i
with T being the temperature (1C), KH the Henry constant (atm m3 mol1), and patm_gas_i the partial pressure in the atmosphere (atm). Wind measurements were corrected to a height of 10 m with the seventh-root profile, and the mass transfer coefficients from carbon dioxide and ammonia were normalized to the oxygen transfer coefficient considering the surface renewal theory (Ro and Hunt, 2006). Figure 27 displays the schematic view of the processes in the model, with the three biomass groups, algae, heterotrophic
Kinetic parameters after calibration (T ¼ 20 1C) compared to the values of the River Water Quality Model No. 1 (Reichert et al., 2001)
Parameter
Unit
WSP model
Reichert
Description
Algae biomass mALG bALG KN,ALG KO,ALG KNH4,ALG KI
d1 d1 g m3 g m3 g m3 W m2
2 0.1 0.1 0.2 0.1 1200
2 0.1 0.1 0.2 0.1 500
Maximum algae growth rate Algae decay coefficient Half-saturation constant for nitrogen Half-saturation concentration for oxygen Ammonia inhibition constant for growth with nitrate Light limitation and saturation coefficient
4.15 104 7.75 108 1 1014 1 105 1 105 0.03 44 4 106 2 109
4.15 104 3.87 107 -
CO2/HCO 3 equilibrium coefficient Inorganic nitrogen equilibrium coefficient Water dissociation equilibrium coefficient Equilibrium rate for carbon dioxide/bicarbonate Equilibrium rate for inorganic nitrogen Henry’s constant to carbon dioxide Henry’s constant to free ammonia Atmospheric pressure from carbon dioxide Atmospheric pressure from free ammonia
Physicochemical parameters gH m3 KaCO2 gH m3 KaIN kw gH m3 d1 kACO2 d1 KANH3 atm m3 mol1 KHCO2 atm m3 mol1 KHNH3 atm patmCO2 atm patmNH3
254
Modeling of Biological Systems
Sun radiation
Wind
Heterotrophs Algae
HCO3−
CO2
O2
Inflow
NH3
Autotrophs
Outflow
NH4+
pH calculation
Light attenuation Figure 27 Extended version of the ASM 3 for the simulation of WSP processes.
and autotrophic biomass, the ionic equilibriums, the gaseous transfers, and the presence of sun and wind.
Meas. COD (g m−3)
2500
−3
Meas. CODs (g m )
Sim. COD (g m−3) Sim. CODs (g m−3)
4.08.5.3.1 Model calibration and simulation results Table 11 shows the adopted parameter values for the new processes for the simulation of both ponds. As observed from experimental data, mean COD degradations in the facultative and maturation ponds were approximately 50%. Determined COD fractions in the inflow to the facultative pond are as follows: inert fraction SI ¼ 45% of total COD, and readily and slowly biodegradable substrate fractions SS ¼ 35% and XS ¼ 20%, respectively. Measured environmental data, sun radiation and wind velocity, were applied in the simulations with their mean values per hour and temperature (25 1C). The results of COD removal are depicted in Figure 28. The heterotrophic growth rate determined by model calibration (0.52 d1) is almost 4 times smaller than found for activated sludge (Gujer et al., 1999). The reduced rate could be explained by the high concentration of ammonia in the leachate (Li and Zhao, 2001). This was also reported by other authors (Yang et al., 2004; Lee et al., 2000), who have reported the effect of free ammonia in the liquid phase on heterotrophic growth and oxygen consumption rates. The mean concentrations of free ammonia obtained from model calculations in the facultative and maturation ponds were 11.4 and 1.5 g m3, respectively; both concentrations could explain the inhibiton of the heterotrophic biomass growth. In the model, oxygen input was considered from two different sources: algae growth and wind aeration. The low values of chlorophyll a suggest that wind is most important here. Model calculations showed a mean oxygen input from wind in the facultative pond of 14.4 and in the maturation pond of 11.1 gO2 m2 d1. The TSS effluent concentrations for both ponds are presented in Figure 29. These results are directly correlated to algae growth, due to their influence on light inhibition. However, more experimental data are necessary to establish the relationship of chlorophyll a, TSS, and available light to photosynthesis. Here, light attenuation coefficients were used
1500 1000 500 0 0
50
100 150 Time (days)
Meas. COD (g m−3)
2500
−3
Meas. CODs (g m )
200
250
Sim. COD (g m−3) Sim. CODs (g m−3)
2000 COD (g m−3)
4.08.5.3 Results and Discussion
COD (g m−3)
2000
1500 1000 500 0 0
50
100 150 Time (days)
200
250
Figure 28 COD effluent measured and simulated results in the facultative pond (left) and maturation pond (right). COD, chemical oxygen demand. From Gehring T, Silva J, Kehl O, et al. (2010) Modelling waste stabilization ponds with an extended version of ASM 3. Water Science and Technology 61(3): 713–720.
in accordance with Jupsin et al. (2003): a1 ¼0.3 and a2 ¼ 0.032. Algae concentrations in both ponds treating leachate wastewater were very low, with 93 and 63 mg l1 chlorophyll a concentrations in the maturation and facultative pond, respectively. Commonly reported data for facultative ponds are in the range of 500–2000 mg l1 (Mara et al., 1992). One
Modeling of Biological Systems
reason for low chlorophyll a could also be free ammonia concentrations. Concentrations in the order of 0.02 g m3 may inhibit the growth of various algae species (Azov and Goldman, 1982). In our simulations, no inhibition function was considered and only the light inhibition/saturation constant KI was calibrated to fit the algae concentration. The final calibrated value was 1200 W m2. Other algae parameters from processes 13, 14, and 15 were maintained, as suggested by Reichert et al. (2001).
500
Meas. COD (g m−3)
Sim. COD (g m−3)
Meas. CODs (g m−3)
Sim. CODs (g m−3)
Simulated algae concentrations and pH for the facultative pond are shown in Figure 30. Algae biomass achieved in simulations was in the range of experimental data and also followed the observed seasonal variations. Despite the strong influence of depth on light availability to photosynthesis, good mixing of the three pond layers equalized the algae concentrations. The model was not able to explain peaks of chlorophyll a at the end of the experiments. The same goes for pH values. In contrast, the pH calculations met the experimental data for the first 150 days (cf. Figure 30). The ammonia dissociation constant showed a strong influence on the pH results, and thus on many other parameters. As no data for the ammonia dissociation constant in leachate were available, this value was adjusted to fit experimental data. The value found corresponds to one-fifth of pure water (Reichert et al., 2001). For animal manure, Ni (1999) found that the dissociation constant ranged from one-fifth to one-sixth of that of ammonia in pure water. Gas release that occurs exclusively in the top layer explains why pH variations are bigger in the first layer than in the other two. In both ponds, the nitrate concentrations were low. Variations were small, with average effluent values of 6 gNO3 2N m3 in the outflow of the facultative pond and 4 gNO3 2N m3 after the maturation pond. According to measurement data, the oxygen concentrations never fell below 1 mg l1 in either pond, which
TSS (g m−3)
400 300 200 100 0 125
150
175 200 Time (days)
225
250
Figure 29 TSS effluent measured and simulated results. TSS, Total suspended solids.
600
255
12
Top
500 10
400 300
8
200 100
Chrolophyll a (mg m−3)
0 500
6 12
Middle
400 10
300 200
8
100 6 12
0 500 Bottom 400
10
300 200
8
100 0 60
80
100
120
140
160
180
Time (days) Figure 30 Chlorophyll a and pH measured and simulated results in the facultative pond in three layers.
200
220
6 240
Modeling of Biological Systems
Meas. Fac. ammonia (g m−3) Meas. Mat. ammonia (g m−3)
1000
Sim. Fac. ammonia (g m−3) Sim. Mat. ammonia (g m−3)
800 600 400 200 0 0
100 150 Time (days)
50
200
250
60 55 50 45
1200
800
400
6
0
8.8 800
4
700 2
600
40 35
9.0
900
65 Chrolophyl a (mg m−3)
Total ammonina nitrogen (g m−3)
70
Sun radiation (W m−2)
Figure 31 Total ammonia nitrogen in effluent, measured, and simulated results.
investigated (575–705 gN m3) here, pH was adjusted above 11. This makes the comparison of these investigations with present results more complicated. The same applies to other published stripping rates from different sources with different pH and nitrogen concentrations. Rumburg et al. (2008) reported fluxes from a plant-scale anaerobic dairy waste lagoon of 2.6–13.0 gN m2 d1 determined through tracer experiments. Experimental stripping rates measured for domestic WSPs (Zimmo et al., 2003; Camargo Valero and Mara, 2007) are very much below these values. Figure 32 illustrates the interaction of some model variables and their variation throughout the day. Presently apparent is direct correlation of sun radiation and chlorophyll a concentrations. Algae growth directly follows sun radiation. Daily peaks of both parameters show a gap of 4–5 h. On day 204, with sun radiation below 300 W m2, algae biomass decreased very quickly. It is still possible to visualize the influence of algae nitrogen uptake that was suggested as the main path for nitrogen removal in WSPs treating domestic wastewater (Camargo Valero and Mara, 2007). Daily peaks of chlorophyll a concentrations corresponded to daily minima of ammonia concentrations. Apart from the oxygen production due to photosynthesis, it was not possible to establish a direct relation between chlorophyll a and the dissolved oxygen concentrations, which are also influenced by variations of the organic load rates and wind velocities. The developed model may be one step toward more detailed WSP modeling which considers complex interactions between microbial and physical/chemical processes. More
8.4 8.2 8.0 7.8
500
30 200
8.6
pH
Tital amount nitrogen (g m−3)
indicates that no denitrification occurred there. More information about the nitrification calibration and nitrate concentration simulation results can be found in Gehring et al. (2010). The simulated concentrations of the total ammonia nitrogen effluent values showed good results for both ponds (Figure 31). The average ammonia stripping rates were 18.2 and 4.5 gN m2 d1 in the facultative and maturation pond, respectively. Unfortunately, there are not many published data available for comparison. Smith and Arab (1988) and Cheung et al. (1997) measured high ammonia stripping rates in free tanks (without aeration) treating landfill leachate. Calculated values ranged between 45 and 142 gN m2 d1. Although initial ammonia concentrations in both studies were very similar to the ponds
Dissolved oxygen (g m−3)
256
202
204
206
208
0 210
7.6
Time (days) Figure 32 Example of the dynamic behavior of the WSP model (top layer of the maturation pond between days 200 and 210). The presented variables include: total ammonia nitrogen (black line), chlorophyll a (gray line), dissolved oxygen (dashed gray line), pH (dashed black line), and the measurements from sun radiation (at the top in dark gray line).
Modeling of Biological Systems
detailed experimental data would help to better understand the findings and to include additional processes such as phosphorus removal, anaerobic digestion in bottom layers, and more sophisticated hydraulic concepts.
4.08.6 Anaerobic Treatment 4.08.6.1 Introduction In the following, results from the anaerobic treatment of manure are presented. Anaerobic processes are widely used, especially for industrial and agricultural wastewater that have much higher COD concentrations than typically found in municipal wastewater. For a few years, energy crops have been appreciated for their potential of producing energy. To be more independent from traditional sources such as oil and gas, agriculturists are being subsidized in order to produce not only food, but to cultivate energy-rich substrates such as maize, rye, and grass as well. To stabilize and increase the reactor operation, these substrates are ensilaged, and cattle, pig, and chicken manure is used as co-substrate. With manure, sufficient trace elements are added to avoid process inhibitions which have been reported to take place sometimes in anaerobic processes run with pure energy crops. To assure hygienic quality of the treated co-substrates and manure, detailed microbial analyses have been executed (e.g., Lebuhn et al., 2005). Manure as one major source of environmental pollution from livestock farming should be treated efficiently in order to avoid its hazardous impact on soil and groundwater. As yet, only little research has been done on the modeling of agricultural biogas plants. Angelidaki et al. (1993), who described the inhibition of the anaerobic processes by ammonia, and Angelidaki et al. (1999), who investigated the cofermentation of agricultural substrate and fats, dealt with the fermentation of manure. Both papers highlight the importance of a detailed fractioning of the input substrate, due to the fact that proteins, carbohydrates, and lipids have different degradation paths in the anaerobic process. Myint et al. (2007) focused on the hydrolysis and acidogenesis in the dry digestion of cattle manure with process TS concentrations higher than 20%. Complementing modeling results are summarized by Batstone et al. (2006), who described the adaptation of ADM 1 (Batstone et al., 2002) to sludge treatment and various industrial wastewaters; recent publications deal with the calibration of ADM1 to pure energy crops (e.g., Amon et al., 2007; Gerin et al., 2008; Wichern et al., 2009; Koch et al., 2009). Apart from the ADM1 calibration for the digestion of cattle manure and co-substrates (Wichern et al., 2008b), this chapter presents further information on parameter sensitivity and energy balances for the optimization of reactor operations. The latter findings were based on research by Lu¨bken et al. (2007). The presented results showed the importance of mathematical modeling to improve reactor operation, increase methane yield, and decrease impact on the environment.
4.08.6.2 Material and Methods
257
State Research Centre for Agriculture. The 3500-l fermenter was discontinuously fed with liquid manure from cattle farming and total mixed ratio (TMR – fodder for cows). The reactor was operated at 38 1C under mesophilic conditions. The contact time of the substrate in the fully mixed reactor amounted to 21 days. Measurements were based on German standard methods. In order to characterize the substrate in terms of carbohydrates, proteins, and fats, a method according to Van Soest and Wine (1967) and Weender (described in Naumann and Bassler (1993)) was applied. This application resulted in a fractionation of the organic matter between crude protein, crude fat, crude fiber, and nitrogen-free extract (Weender analysis). The so-called van Soest analysis allows for the determination of neutral detergent fiber (NDF), acid detergent fiber (ADF), and acid detergent lignin (ADL). The carbohydrates were further divided into hemi-cellulose (NDF– ADF), cellulose (ADF–ADL), and lignin (ADL). The total biogas production was measured by the drum chamber gas meter TG5/5 (Ritter, Germany). Values for biogas production were normalized. Methane and carbon dioxide were quantified by means of the infrared two-beam compensation method with pressure compensation (measuring error as specified:72%). Oxygen and hydrogen were measured by electro-chemical sensors (measuring error as specified:73%).
4.08.6.2.2 Reactor operation The investigated anaerobic reactor was operated with a substrate mixture of liquid manure of cattle and TMR. The co-substrate TMR was composed of 43% corn silage, 18% gramineous silage, 12% crop groats, 9% water, 7% soy pellets, 7% cow grain, and 4% hay. The mean influent volume flow amounted to 175 l d1. This resulted in a volumetric load rate of 3.6 kgVS m3 d1 and a COD load of 15.3 kgCOD d1. COD in the effluent was reduced by 30–35% compared to the influent COD. The COD/TS ratio of the inflow substrate was iCOD/TS ¼ 1.2 kgCOD kg1 TS . The dry gas production was measured to be 3.65 m3Gas d1 (287 lBiogas kg1 VS ) at operation temperature. The pH value in the reactor was relatively constant at 7.6. Table 12 shows the characteristics of the inflow substrate in detail.
4.08.6.2.3 Mathematical model and sensitivity functions In 2002, the IWA Task Group on Mathematical Modelling of Anaerobic Digestion Processes presented the Anaerobic Digestion Model No. 1 (Batstone et al., 2002). ADM 1 is a highly complex model, characterized by 19 biochemical conversion processes and 24 substances. To investigate the application of ADM 1 to agricultural substrates, a sensitivity analysis of both inflow COD fractioning and biochemical parameters was run. The applied technique is called SVM slope technique, published by Kim et al. (2006), and was adapted here for use with ADM 1. The calibrated ADM 1 model (see Table 13) was used as reference parameter set. An effluent base quality EQ was defined:
EQ ¼ bCOD CODe þ b CH4 CH4 % þ b CO2 CO2 % þ bgasflow qgasflow ðÞ
ð31Þ
4.08.6.2.1 Analytical methods Analyses of the agricultural substrates were run by the Institute of Agricultural Engineering and Animal Husbandry, Bavarian
with CODe being the COD in the effluent (g m3), CH4% the percentage of methane in the gas, CO2% the percentage of
258
Modeling of Biological Systems
carbon dioxide in the gas, qgasflow the dry gas flow in m3 d1, and b the weighting factors (10, 20, 20, 20). The sensitivity index DEQ results from a procedure where each model parameter was changed incrementally by 10% (Ref: referring to reference simulation, Var: varied parameter):
been conducted to analyze the substrate (see Table 12). The following equations are implemented to define particular model fractions:
XPr ¼ FM TS iCOD=TS RP
DEQ ¼ bCOD CODRef ;e CODVar;e þ bCH4 jCH4 %Ref CH4 %Var j þ bCO2 jCO2 %Ref CO2 %Var j þ bgasflow qgasflow;Ref qgasflow;Var ðÞ
ð32Þ
4.08.6.3 Results and Discussion 4.08.6.3.1 Calibration of the ADM 1 The inflow fractioning of the total COD is of highest importance for the calibration of the ADM 1 and strongly affects the gas composition. For this, detailed measurements have Table 12
Characteristics of the influent substrate
Parameter
Unit
Manure
TMR
TS VS COD iCOD/VS VFAtotal Alkalinity pH NH4–N Raw protein Raw fiber Raw lipid NfE NDF ADF ADL
(%) (%TS) (kg m3) (kgO2 kg1 VS ) (g m3) (mmol l1) (–) (g m3) (% TS) (% TS) (% TS) (% TS) (% TS) (% TS) (% TS)
6.1 81.4 76 1.53 6657 241.9 7.4 2289 12.2 17.8 4.3 47.1 47.1 33.9 20.1
50.3 93.7 609 1.29 3765 48.2 4.9 1345 19.2 17.2 2.6 54.7 50.0 23.4 19.6
ADF, acid detergent fiber; ADL, acid detergent lignin; COD, chemical oxygen demand; NDF, neutral detergent fiber; NfE, nitrogen-free extract; TMR, total mixed ratio; TS, total solids, VS, volatile solids; and VFAtotal, total volatile fatty acids. From Wichern M, Lu¨bken M, Schlattmann M, Gronauer A, and Horn H (2008b) Investigations and mathematical simulation on decentralized anaerobic treatment of agricultural substrate from livestock farming. Water Science and Technology 58(1): 67–72.
Table 13
ðkgCOD d1 Þ
ð33Þ
XLi ¼ FM TS iCOD=TS RL ðkgCOD d1 Þ
ð34Þ
The two parameters proteins XPr and lipids XLi are defined by the fresh mass FM (kgFM d1), the TS (%), the COD content of manure iCOD/TS, as well as raw protein RP (%TS) and raw lipids RL (%TS), respectively. The calculation of carbohydrates XCH and inert material is more complicated and is based on additional information from the Van Soest analysis (Van Soest and Wine, 1967).
XI ¼ FM iTS=FM iCOD=TS ADL þ ðADF ADLÞnon
deg
ðkgCOD d1 Þ
ð35Þ
To quantify the inert material, the load of lignin ADL (%TS) is needed. ADF (%TS) comprises lignin and cellulose. Measurements showed that cellulose is degraded by 28%, which is reasonable according to Fuchigami et al. (1989). The low degradation of COD implies that manure had been degraded in the animal intestines before:
XCH ¼ FM iTS=FM iCOD=TS h ðRF þ NfeÞ ADL þ ðADF ADLÞnon
i deg
ðkgcod d1 Þ
ð36Þ
Raw fiber RF (%TS) and nitrogen-free extract Nfe (%TS) represent the total content of carbohydrates, whereas the latter part of the equation describes the inert part consisting of lignin and nondegradable cellulose. If Equations (33)–(36) are applied, the particulate COD, which is 80% of the total COD, can be divided as follows: raw protein Xpr ¼ 13.6%, raw fat Xli ¼ 4.0%, carbohydrates Xch ¼ 54%, and inert material XI ¼ 28.4%. All particular material was split during the disintegration process into the aforementioned fractions. To fulfill the nitrogen mass balance after the disintegration step (Blumensaat and Keller, 2004), the nitrogen content of the composite and inert material was fitted to NXC,I ¼ 0.0014 moln g1 COD. Acetate, propionate, butyrate, and valerate were measured as 5.2%, 3.5%, 2.0%, and
Calibrated biochemical ADM 1 parameters for the treatment of cattle manure (see also Wichern et al., 2008b)
Parameter
Description
Unit
ADM 1 value
Calibrated
Notes
kDis km,ac pHUL,acid pHLL,acid km,pro KS,pro KS,H2 NXC,I
Disintegration constant Acetate uptake rate Upper pH limit for acidogens Lower pH limit for acidogens Propionate uptake rate Half-saturation coefficient for propionate uptake Half-saturation coefficient for hydrogen uptake Nitrogen content of composite and inert material
d1 g g1 d1
0.5 8 5.5 4 13 0.1 7 106 0.002
0.05 4.2 8 6 4.5 0.34 1.65 105 0.001 4
1 1 1 1 2 2 1 1
g g1 d1 kg m3 kg m3 molN m3
(1) Values determined by best fit between dynamic simulation and measurement. (2) Values in accordance with Angelidaki et al. (1999).
Modeling of Biological Systems
0.6% of CODtot, respectively. Calibration was mainly done with the disintegration constant changed to kDis ¼ 0.05 d1 and the acetate uptake rate to km,ac ¼ 4.2 g g1 d1. The complete list of calibrated biochemical parameters can be found below in Table 13.
4.08.6.3.2 Modeling reactor performance Exemplary modeling results are depicted in Figure 33. Results for gas flow (Figure 33 (left)) and gas composition (Figure 33 (right)) showed only smaller deviations between measuring and simulation results. Besides gas flow and gas composition, propionate and acetate concentrations were also calibrated (not shown).
4.08.6.3.3 Sensitivity analysis for the biochemical parameters and the inflow fractioning Identifying sensitive biochemical parameters is not only necessary to better understand the applied model, it is also most useful to identify sensitive inflow parameters that should be measured in great detail to avoid false results or misleading conclusions. In the following, results from sensitivity analyses are presented. It becomes apparent that for cattle manure and co-substrates, the hydrolysis rate kHyd, the uptake rates for amino acids km,aa and sugars km,su, the hydrogen inhibition constant KI,H2,pro, and the half-saturation coefficients for propionate KS,pro and hydrogen KS,H2 are less sensitive. In contrast, the disintegration rate kDis, the acetate uptake rate
400 300 200 100 0 0
100
10
20
30 40 Time (days)
Meas. CH4 (%) Sim. CH4 (%)
50
60
Sensitivity index dEQ (−)
250
kdis khyd km,ac
200
km,ac High sensitivity
300 100 50
Low sensitivity 0
Meas. CO2 (%) Sim. CO2 (%)
50
75
80 60
100 125 150 Parameter variation (%)
175
250
40 20 0 0
10
20
30 40 Time (days)
50
60
Figure 33 Gas flow in m3 d1 (left) and dry gas composition in % (right). From Wichern M, Lu¨bken M, Schlattmann M, Gronauer A, and Horn H (2008b) Investigations and mathematical simulation on decentralized anaerobic treatment of agricultural substrate from livestock farming. Water Science and Technology 58(1): 67–72.
Sensitivity index dEQ (−)
Gas flow (m3 d−1)
km,ac, the ammonia inhibition constant KI,NH3, and the biomass decay rates kdec are more sensitive. When applying the sensitivity analysis of Kim et al. (2006), the results very much depend on the weighting factors b. Here, gas volume and gas composition (CH4 and CO2) were considered to have higher importance than COD effluent values. If the acetate uptake rate km,ac and the inhibition constant for ammonia KI,NH3 were further reduced compared with the calibrated parameter set, this would result in complete process inhibition. If, for instance, measured acetate concentrations are available, in ADM 1 the parameters km,ac, KI,NH3, and KS,ac notably affect the acetate simulation results. In many cases, these three parameters cannot be identified exactly. From the mathematical point of view, various typical parameter sets can be identified, which leads to an even simulation quality. However, the modeling engineer has to decide which parameter set is reasonable from the microbial and engineering points of view. Figure 34 presents the results of the sensitivity analysis for the inflow fractioning of COD after the disintegration step. The total sum of composite material stayed the same; only the fractions of proteins, carbohydrates, lipids, and inert material were changed. The diagram outlines the importance of inert material and carbohydrates fraction for the gas flow and gas composition. For the investigated reactor and substrates, proteins are less sensitive. The quantity of ammonia nitrogen in cattle manure is much higher than the incorporated nitrogen in the COD fractions (Figure 35).
Meas. gas ammonia (g3 d−1) Sim. gas flow (m3 d−1)
500
Gas composition (%)
259
200
kdis khyd km,ac km,su km,aa
200 300 100 50 0 50
75
100 125 150 Parameter variation (%)
175
200
Figure 34 Sensitivity index DEQ for exemplary parameters of ADM 1.
260
Modeling of Biological Systems
fpr,xc fli,xc fch,xc fxi,xc
Sensitivity index dEQ (−)
200 150 High sensitivity 100 50
Low sensitivity
0 50
75
100 125 150 Parameter variation (%)
175
200
with Qin being the inflow rate (kg s1), H the conveyor height (m), r the density of the pumped media (kg m3), g the acceleration of gravity (m s2), tp the time for pumping (h d1), and Zecc_worm the efficiency degree (–). The efficiency degree strongly depends on the type of the pump used and can vary considerably. For the eccentric worm pump used for the pilotscale digester, Z values range between 0.3 and 0.7. Here, we considered that reactor feeding was 15 min once a day. The required energy for the stirrer ensuring good mixing and sufficient contact between substrate and microorganisms depends on reactor volume, reactor geometry, and viscosity of the medium:
Ploss stir ¼ V liq S ts
Figure 35 Sensitivity index DEQ for the COD inflow fractioning of cattle manure.
4.08.6.3.4 Simulation of the energy balance As a next step, a dynamic energy balance model was derived, which considers energy production and consumption. The model was derived by Lu¨bken et al. (2007). Energy consumption in a reactor results from pumping, stirring, substrate heating, and the compensation of radiation loss, whereas energy is mainly produced by utilizing the energy contained in biogas. The basic equation for the dynamic energy balance model is
dPnet prod loss ¼ Pelect Ploss P pump stir dt
prod loss loss þ Pprod therm Prad Psub heat þ Pmic heat
ðkW h d1 Þ
ð37Þ
with Pnet being the net energy production of the digester (kW h prod d1), Pelect the electrical energy production (kW h d1), Ploss pump the mechanical power of the pump (kW h d1), Ploss stir the prod mechanical power of the stirrer (kW h d1), Ptherm the thermal energy production (kW h d1), Ploss rad the radiation loss (kW h the heat requirement for substrate heating (kW h d1), Ploss sub heat 1 d1), and Pprod mic heat the microbial heat production (kW h d ). The single terms were calculated as follows:
Pprod elect
¼ QG PCH4 HC Z elect
1
ðkW h d Þ
Pprod therm ¼ QG PCH4 HC Ztherm
ðkW h d1 Þ
ð38Þ ð39Þ
with QG being the biogas production (m3 d1), PCH4 the methane content (%), HC the calorific value of methane (kW h N m3), Zelect the electrical degree of efficiency () and Ztherm the thermal degree of efficiency (). Using Equations (38) and (39) assumes a co-generation plant for heat and power. When biogas is used in a combined heat and power plant (CHP), it is of utmost importance for cost-effective plant operation to use both mechanical/electrical energy and thermal energy. Electrical energy is produced in a CHP unit with a mechanical/electrical efficiency degree of approximately 35%, whereas the thermal efficiency degree is approximately 50%:
Ploss pump ¼ Qin H r g tp
1 Zecc
worm
ðkW h d1 Þ
ð40Þ
ðkW h d1 Þ
ð41Þ
with Vliq being the liquid volume (m3), S the specific power of the stirrer (kW m3), and ts the time for stirring (h d1). Here it was supposed that the stirrer is used every half hour for 10 min with a consumption of 0.005 kW m3:
Ploss rad ¼ Kheat trans Tliq Tambient Vliq þ Tgas Tambient Vtot V liq 2 24 ðkW h d1 Þ r 1000
ð42Þ
with Kheat_trans being the heat transfer coefficient (Wh m2 h1 K1), Tliq the temperature of the substrate within the digester (K), Tambient the ambient temperature, Vliq the liquid volume (m3), Tgas the gas temperature (K), Vtot the total digester volume (m3), and r the radius of the digester (m). Ambient temperature varied between 25 and 10 1C:
1 Ploss sub heat ¼ Qin c Tdigester Tsubstrate 3:6
ðkW h d1 Þ
ð43Þ
with Qin being the reactor inflow (m3 d1), c the heat capacity of the substrate (kJ kg1 K1), Tdigester the temperature of the digester (K), and Tsubstrate the temperature of the stored substrate (K:
Pprod mic heat ¼
X j¼512
DEj f j rj Vliq
1 3:6
ðkW h d1 Þ
ð44Þ
with DEj being the energy released to the environment due to microbial activity of process j (kJ mol1), fj the molar mass per gCOD of the educt of process jðmolgCOD 1 Þ, rj the kinetic rate equation of process j of ADM 1 (kgCOD m3 d1), and Vliq the liquid volume (m3). The thermodynamic parameters of selected anaerobic biochemical reactions are presented in Table 14. Long-chain fatty acids (LCFAs) are represented by palmitate, and glucose stands for monosaccharide. The thermodynamics of amino acids were calculated for glycine and alanine degradation in the Stickland reaction. Figure 36 suggests that energy production was nearly constant for the investigated period. As energy consumption strongly depends on the respective season, the lowest energy consumption was calculated in summer time (July) and the maximum in winter time (December). For the investigated pilot-scale digester, heating of substrate needed the most
Modeling of Biological Systems Table 14
261
Thermodynamics of biochemical reactions implemented in ADM 1 DH0f (kJ DG0f (kJ DG0’ (kJ DG0T’ (kJ ATP M1) M1) M1) M1) (M1)
Reaction
DE (kJ M1)
Glucose: 1. C6H12O6 þ 2H2O-2CH3COO þ 2Hþ þ 2CO2 þ 4H2 89.5 136.0 215.7 183.4 231.5 -311.3 2. 3C6H12O6-4CH3CH2COO þ 2CH3COO þ 6Hþ þ 2CO2 þ 2H2O 3. C6H12O6-CH3CH2CH2COO þ Hþ þ 2CO2 þ 2H2 47.5 224.1 264.0 þ þ CH3CH(NHþ 4.0 51.1 51.1 3 )COO þ 2CH2(NH3 )COO þ 2H2O-3CH3COO þ CO2 þ 3NH4
Amino acids: Palmitate: CH3(CH2)14COO þ 14H2O-8CH3COO þ 7Hþ þ 14H2 Valerate: CH3CH2CH2CH2COO þ 2H2O-CH3CH2COO þ CH3COO þ Hþ þ 2H2 Butyrate: CH3CH2CH2COO þ 2H2O-2CH3COO þ Hþ þ 2H2 Propionate:CH3CH2COO þ 2H2O-CH3COO þ CO2 þ 3H2 Acetate: CH3COO þ Hþ-CH4 þ CO2 Hydrogen: 4H2 þ CO2-CH4 þ 2H2O
966.2 136.2 137.0 204.7 16.2 63.2
670.2 88.2 88.2 71.7 75.7 –32.7
391.1 48.3 48.3 71.7 35.8 32.7
225.53 313.35 271.70 53.13
4 4/3 3 1/3
25.53 246.69 121.70 36.46
378.22 14/6 494.88 46.24 0.875 89.99 46.17 0.75 83.67 65.87 0.50 90.87 39.84 0.25 27.34 –31.36 1/4 –18.86
Values for the thermodynamic parameters refer to the molar mass of the selected educts. The fraction of glucose which degrades via the first, second, and third reactions is: (1) ¼ 50%, (2) ¼ 35%, and (3) ¼ 15% (Lu¨bken et al., 2007).
20.0 (e) 18.0
(e)
(e)
(e)
(e)
(e) 16.0
P (kWh d−1)
14.0 12.0 10.0 8.0 6.0 (c) 4.0 2.0 0.0
(c) (b) (d) (a) Jul. 04
(c) (b) (a)
(d)
Aug. 04
(c) (b)
(c) (b) (a)
(d)
Sep. 04
(a)
(d)
Oct. 04 Months
(b)
(a)
(d)
Nov. 04
(c) (b)
(a)
(d)
Dec. 04
Figure 36 Comparison of the simulated energy production and the simulated energy consumption: (a) energy consumption of stirrer, (b) energy consumption due to radiation loss, (c) energy consumption due to substrate heating, (d) energy production due to microbial activity, and (e) sum of thermal and electric energy production of CHP.
energy, followed by radiation loss. Other factors were negligible. Heat produced by microbial degradation of organics made about 10% of the energy necessary for substrate heating.
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Hulsbeek JJW, Kruit J, Roeleveld PJ, and Van Loosdrecht MCM (2002) A practical protocol for dynamic modelling of activated sludge plants. Water Science and Technology 45(6): 127--136. Hurse JT and Connor MA (1999) Nitrogen removal from wastewater treatment lagoons. Water Science and Technology 39(6): 191--198. Johansson P, Carlsson H, and Jo¨nsson K (1996) Modelling of an anaerobic reactor in a biological phosphate removal process. Water Science and Technology 34(1–2): 49--55. Jupsin H, Praet E, and Vasel J-L (2003) Dynamic mathematical model of high rate algal ponds (HRAP). Water Science and Technology 48(2): 197--204. Kadlec RH (2000) The inadequacy of first-order treatment wetland models. Ecological Engineering 15: 105--119. Kim JR, Ko JH, Lee JJ, et al. (2006) Parameter sensitivity analysis for Activated Sludge Models No. 1 and 3 combined with one-dimensional settling model. Water Science and Technology 53(1): 129--138. Koch G, Ku¨hni M, Gujer W, and Siegrist H (2000) Calibration and validation of Activated Sludge Model No. 3 for Swiss municipal wastewater. Water Research 34(14): 3580--3590. Koch K, Wichern M, Lu¨bken M, and Horn H (2009) Mono fermentation of pure grass silage by means of loop reactors. Bioresource Technology 100(23): 5934--5940. Kreikenbohm R and Stephan W (1985) Application of a two-compartment model to the wall growth of Pelobacter acidigallici under continuous culture conditions. Biotechnology and Bioengineering 27: 296--301. Langergraber G (2003) Simulation of subsurface flow constructed wetlands – results and further research needs. Water Science and Technology 48(5): 157--166. Langergraber G (2005) The role of plant uptake on the removal of organic matter and nutrients in subsurface flow constructed wet-lands: A simulation study. Water Science and Technology 51(9): 213--224. Lebuhn M, Effenberger M, Garces G, Gronauer A, and Wilderer PA (2005) Hygienisation by anaerobic digestion: Comparison between evaluation by cultivation and quantitative real-time PCR. Water Science and Technology 52(1–2): 93--100. Lee S-M, Jung J-Y, and Chung Y-C (2000) Measurement of ammonia inhibition of microbial activity in biological wastewater treatment process using dehydrogenase assay. Biotechnology Letters 22: 991--994. Li XZ and Zhao QL (2001) Efficiency of biological treatment affected by high strength of ammonium-nitrogen in leachate and chemical precipitation of ammoniumnitrogen as pretreatment. Chemosphere 44: 37--43. Lindenblatt C, Wichern M, and Horn H (2007) Wastewater treatment with activated pre-clarifier and planted soil filters. Water Science and Technology 55(7): 195--202. Lu¨bken M, Wichern M, Schlattmann M, Gronauer A, and Horn H (2007) Modelling the energy balance of an anaerobic digester fed with cattle manure and renewable energy crops. Water Research 41(18): 4085--4096. Makinia J, Rosenwinkel K-H, and Spering V (2005) Long-term simulation of the activated sludge process at the Hanover-Gu¨mmerwald pilot WWTP. Water Research 39(8): 1489--1502. Mara DD, Alabaster GP, Pearson HW, and Mills SW (1992) Waste Stabilization Ponds: A Design Manual for Eastern Africa, 121pp. Leeds: Laggon Technology International. McBride GB and Tanner CC (2000) Modelling of biofilm nitrogen transformations in constructed wetland mesocosms with fluctuating water levels. Ecological Engineering 14: 93--106. Myint M, Nirmalakhandan N, and Speece RE (2007) Anaerobic fermentation of cattle manure: Modelling of hydrolysis and acidogenesis. Water Research 41(2): 323--332. Molle P, Lienhard A, Botin C, Merlin G, and Iwema A (2005) How to treat raw sewage with constructed wetlands: An overview of the French systems. Water Science and Technology 51(9): 11--22. Morgenroth E and Wilderer PA (2000) Influence of detachment mechanisms on competition in biofilms. Water Research 34(2): 417--426. Murnleitner E, Kuba T, Van Loosdrecht MCM, and Heijnen JJ (1997) An integrated metabolic model for the aerobic and denitrifying biological phosphorus removal. Biotechnology and Bioengineering 54(5): 434--450. Naumann C and Bassler R (1993) Die chemische Untersuchung von Futtermitteln, 3rd edn. Darmstadt: VDLUFA-Verlag. Ni JQ (1999) Mechanistic models of ammonia release from liquid manure: A review. Journal of Agricultural Engineering Research 72: 1--17. Nowak O (1996) Nitrifikation im Belebungsverfahren bei maXgebendem IndustrieabwassereinfluX. Dissertation, Wiener Mitteilungen, 135. Pano A and Middlebrooks EJ (1982) Ammonia nitrogen removal in facultative ponds. Journal of the Water Pollution Control Federation 4(54): 344--351.
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4.09 Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens H Furumai, F Nakajima, and H Katayama, The University of Tokyo, Tokyo, Japan & 2011 Elsevier B.V. All rights reserved.
4.09.1 4.09.2 4.09.2.1 4.09.2.2 4.09.2.3 4.09.2.4 4.09.2.5 4.09.2.6 4.09.3 4.09.4 References
Introduction Physicochemical Characterization of Road Dust and Soakaway Sediment Contamination by Polycyclic Aromatic Hydrocarbons and Their Source Size and Density Distributions of PAHs in Road Dust Evaluation on Heavy Metal Retention in Road Dust and Soakaway Sediments Identification of Elements in Individual Particles by Electron Probe Microanalysis Leaching Potential of Heavy Metal in Road Dust Runoff Behavior of Particle-Associated PAH Pathogenic Pollution in a Seaside Park after CSO Summary
4.09.1 Introduction Urbanization increases the variety and amount of pollutants carried into receiving waters. Therefore, urban environmental water is very susceptible to pollutants from urban activities, which have an adverse effect on water quality and aquatic ecosystems. The sources of the discharge of such pollutants are categorized into two types: point and nonpoint sources. Point source refers to pollutant sources which are easily identified facilities discharging pollutants to the environment. Factories and sewage treatment plants are included as representatives in this category. In many countries, the quality of the discharged water from these point sources is regulated by water pollution control laws and can be managed by a wide range of treatment technologies. Nonpoint source pollution is generally considered to be a diffuse source of pollution not associated with a specific point of entry into water bodies. The urban sources of pollutants, such as vehicles and urban surface materials, are called nonpoint sources. Nonpoint source pollution occurs with rainfall or snowmelt. The water from rain or snow dissolves the
265 266 266 267 267 270 270 271 272 274 275
atmospheric pollutants, washes off the pollutants on the impervious surfaces, and finally flows into rivers, lakes, and coastal waters. Impervious surfaces, such as building roofs, traffic roads, and parking lots, are constructed during urban development. During rainfall and other precipitation events, these surfaces carry polluted runoff to drainage system, instead of allowing the water to percolate through soil. The urban runoff behavior shows the difference between combined and separate sewer systems as shown in Figure 1. In Japan, separate sewer systems have been installed in many cities since 1970 to improve water pollution control in public water bodies. However, combined sewer systems are in place for historical reasons in old and large cities, including Tokyo or Osaka. The combined sewer systems are sewers that are designed to collect rainwater runoff, domestic sewage, and industrial wastewater in the same pipe. Most of the time, combined sewer systems transport all of their wastewater to sewage treatment plants, where it is treated and then discharged to water bodies. During periods of heavy rainfall or snowmelt, however, the wastewater volume in a combined sewer system can exceed the
Combined sewer overflow
River
Combined sewer
Storm sewer
River
Primary treatment
Sanitary sewer Treatment Secondary treatment plant
Combined sewer system
Treatment plant
Secondary treatment
Separate sewer system
Figure 1 Two types of sewer systems and combined sewer overflow.
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capacity of the sewer system or treatment plant. For this reason, combined sewer systems are designed to overflow occasionally and discharge excess wastewater directly to nearby streams, rivers, or other water bodies. This phenomenon is called combined sewer overflow (CSO) and has been recognized as a serious source of environmental water pollution. Naturally, the CSO and the nonpoint source pollution are difficult to control, since the water is irregularly discharged. In order to control water pollution and manage water quality in urban environmental waters, we have to investigate the irregularly occurring wet weather pollution phenomena and their pollutant sources. The monitoring or sampling of such irregular water discharge requires special devices and/or incurs high labor costs. The source responsible for the pollution is often unclear, not least because the water runoff itself is a natural phenomenon and the pollutant sources are diverse; responsibility is thus difficult to assign. As such, the significance of nonpoint source pollution in the water environment tends only to be recognized after the controlling system of the point sources has been spread well in the society. This chapter includes the following: 1. characterization of urban nonpoint pollutants and urban runoff behavior and 2. pathogenic pollution in coastal area after CSO. The first part introduces research on physicochemical characterization of road dust and sediments in infiltration facilities as pollutant source in urban area. Infiltration facilities have been constructed mainly aiming at inundation control in rapidly urbanized areas by reduced storm-water peak flows. These infiltration facilities are likely to contribute to reduction of nonpoint pollutant loads from urban surfaces. Therefore, accumulated sediment in infiltration facilities should be regarded as secondary pollutants in urban runoff pollution and a possible source to groundwater contamination. The second part explains microbial contamination in Tokyo Bay after CSO events, in which enteric virus behavior as well as bacterial indicators, such as total coliforms and Escherichia coli, were extensively investigated.
environments and aquatic ecosystems. Hence, control strategies for PAHs in urban runoff are required to ensure human and ecosystem safety. The effective control can be achieved by investigation of contamination levels of PAHs and understanding their sources. However, there have been limited attempts carried out for quantitative assessment of comparative contribution of various PAH sources to road dust. Road dust has been recognized as bringing a large volume of PAHs into the water environment via road runoff (Brown et al., 1985; Maltby et al., 1995a, 1995b; Boxall and Maltby, 1995). Possible PAH sources in road dust include diesel vehicle exhaust, gasoline vehicle exhaust, tire, pavement (asphalt or bitumen), and oil spill. Based on the enrichment factor (EF), Takada et al. (1990) identified vehicle exhaust as the primary PAH contributor to road dust collected from roads with heavy traffic, while atmospheric fallout was more significant in residential areas in Tokyo. Pengchai et al. (2004, 2005) estimated the comparative contribution from potential PAH sources in road dust samples, using a statistical approach based on a large number of reported PAH profiles. Seven types of PAH sources were defined: diesel vehicle exhaust, gasoline vehicle exhaust, tires, asphalt–pavement, asphalt or bitumen, petroleum products excluding tires and asphalt, and combustion products except those in vehicle engines. As many as 189 PAH data of possible sources were obtained from literature and by additional sampling and measurement. The obtained source data were categorized into seven possible groups as shown in Table 1: diesel vehicle exhaust [D], gasoline vehicle exhaust [G], tire [T], asphalt– pavement [P], asphalt or bitumen [A], petroleum products, excluding tire and asphalt [O], and combustion products, except for those in vehicle engines [E]. Using cluster analysis combined with principal component analysis, the 189 source data were classified into 11 source groups based on the content percentage of 12 individual PAHs (12-PAH profiles), as shown in Table 2. It could be interpreted that the 12-PAH profiles of samples in S1, which have pyrene, benzo(ghi)perylene, and fluoranthene as the predominant PAH species (43%, 19%, and 13%), indicated [T] because all the [T] data were included in S1 and, in reverse, most of the
4.09.2 Physicochemical Characterization of Road Dust and Soakaway Sediment 4.09.2.1 Contamination by Polycyclic Aromatic Hydrocarbons and Their Source Micropollutants, such as polycyclic aromatic hydrocarbons (PAHs) and heavy metals, are widely distributed in dust, soils, and sediments, and are found in roof and road runoff (Hoffman et al., 1984; Takada et al., 1990; Sansalone and Buchberger, 1997; Roger et al., 1998; Heaney et al., 1999; Fo¨rster, 1999; Krein and Schorer, 2000; Chebbo et al., 2001; Furumai et al., 2002; Brenner et al., 2002; Murakami et al., 2003). PAHs are known to be acutely toxic, genotoxic, and carcinogenic compounds (Phillips, 1983; Hagris et al., 1984; Baumann, 1998). Hoffman et al. (1984) estimated that 36% of environmental PAH input was due to urban runoff; for the highermolecular-weight PAHs, the figure was 71%. Urban runoff has been recognized as an important PAH pathway to water
Table 1
Number of collected data as possible PAH sources
Source category
Number of PAH data Data reported in Pengchai et al. (2004)
Diesel vehicle exhaust [D] Gasoline vehicle exhaust [G] Tire [T] Asphalt-pavement [P] Asphalt or bitumen [A] Petroleum products excluding tire and asphalt [O] Combustion products except for those in vehicle engines [E] Total
2 4 8 8 8
Literature cited in Pengchai et al. (2004) 77 49
3 10 20
30
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Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens Table 2
267
Classification of PAH profiles in 189 source data S1
Average PAH profiles Phenanthrene (Ph) Anthracene (An) Fluoranthene (Fr) Pyrene (Py) Benzo(a)anthracene (Ba) Chrysene (Ch) Benzo(k)fluoranthene þ benzo(b)fluoranthene (Bf) Benzo(a)pyrene (Bpy) Indeno(1,2,3-cd)pyrene (In) Dibenz(a,h)anthracene (Db) Benzo(ghi)perylene (Bpe) Total % Sample number belonging to each group Diesel vehicle exhaust [D] Gasoline vehicle exhaust [G] Tire [T] Asphalt-pavement [P] Asphalt or bitumen [A] Petroleum products, excluding tire and asphalt [O] Combustion products, except for those in vehicle engines [E] Total number of samples
S2
S3
S4
S5
S6
S7
S8
S9
S10
S11
6 1 13 43 0 3 3 4 6 1 19
26 4 22 31 4 4 2 2 2 1 2
7 43 9 8 3 3 8 4 7 4 4
3 27 15 13 26 1 4 3 2 4 2
4 7 3 14 11 10 28 0 1 21 1
31 8 1 2 7 10 27 0 2 7 5
16 3 6 11 5 12 15 8 6 3 15
17 2 6 19 2 2 10 31 1 8 1
11 3 51 6 4 4 7 5 3 0 5
1 2 0 0 1 1 6 11 23 46 9
48 15 8 10 5 2 3 3 1 2 2
100
100
100
100
100
100
100
100
100
100
100
1
49 8
5 2
10
8
2 1
2 2
1
20 21
8 2
9
6 9
1 1
1
61
8
10
8
3
19
7
5
7
5
2
2 7 6
3
56
From Pengchai P, Nakajima F, and Furumai H (2005) Estimation of origins of polycyclic aromatic hydrocarbons in size-fractionated road dust in Tokyo with multivariate analysis. Water Science and Technology 51(3–4): 169–175.
data in S1 were from [T]. Likewise, S2 and S3 implied [D] and S7 represented [P] and [A]. S6 and S10 were minor groups having only three data each. S11 included a large number of source data in various categories and was difficult to be translated. Thirty-seven dust samples on nine streets in Tokyo were collected and subjected to PAH analysis both with and without particle size fractionation. Multiple regression analysis was applied to estimate the sources of the PAHs in the dust samples. The result demonstrated that the abrasion of tires and asphalt–pavement contributed a certain amount of PAHs to road dust, in addition to diesel vehicle exhaust, which has been recognized as the main source of PAHs in road dust.
4.09.2.2 Size and Density Distributions of PAHs in Road Dust Particle size and density are important parameters in wash-off processes for urban runoff. PAH distribution in harbor sediment fractions has been reported in both size and density (Ghosh et al., 2000; Rockne et al., 2002; Ghosh et al., 2003). Rockne et al. (2002) revealed that 85% of the total PAHs in Piles Creek sediment was found in the light fractions (o1.7 g cm3), despite the fact that light density components comprised only 4% of the total sediment mass. In addition, they suggested that the preferential sequestration in the Piles Creek sediment was likely due to the presence of detrital plant debris. Ghosh et al. (2000) showed that the coal-/wood-derived particles (specific gravity o1.8) constituted only 5% of Milwaukee Harbor sediment by weight, but contained 62% of the total PAHs.
Murakami et al. (2005) reported PAH concentrations in size- and density-fractionated road dust collected in Japan, as shown in Figure 2. The percentage contribution by weight of light density particles to the total deposition mass was less, but the light fractions accounted for a significantly higher ratio of PAH mass in road dust. It is suggested that light fractions in road dust contribute significantly to storm-water contamination, despite their minor contribution to the total deposition mass, due to their high PAH contents as well as their physical property of high mobility. The cluster analysis revealed that there was a significant difference in the PAH profiles between locations rather than between size fractions, density fractions, and period of sampling. Apart from the locations, the PAH sources might differ due to sampling time or size fractions. Multiple regression analysis indicated that asphalt/pavement was the major source of road dust in residential areas, and that tires and diesel vehicle exhaust were the major source of road dust in heavily trafficked area.
4.09.2.3 Evaluation on Heavy Metal Retention in Road Dust and Soakaway Sediments Infiltration facilities have the potential to serve as both sinks and sources of urban nonpoint pollutants during the process of groundwater recharge by storm water. Surface sediments found in infiltration facilities are known to have high heavy metal content (Mikkelsen et al., 1996; Datry et al., 2003; Dechesne et al., 2004). This implies that promoting urban storm-water infiltration may result in a pollutant transport from urban surface to groundwater. To minimize the
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens
12-PAHs (µg g−1)
40 20 106−250 µm (H)
63−106 µm (H)
0.6−63 µm (H)
106−1250 µm (L)
63−106 µm (L)
0.6−63 µm (L) 80 60 40
250−2000 µm (H)
106−250 µm (H)
0.6−63 µm (H)
63−106 µm (H)
250−2000 µm (L)
106−250 µm (L)
20 0
(c)
Hongo Street
63−106 µm (L)
12-PAHs (µg g−1) 250−2000 µm (H)
(b)
106−250 µm (H)
10 0
60
106−250 µm (H)
63−106 µm (H)
0.6−63 µm (H)
20
63−106 µm (H)
250−2000 µm (H)
63−106 µm (H)
106−250 µm (H)
0.6−63 µm (H)
250−2000 µm (L)
0
106−250 µm (L)
10
30
Light particles: 44 ± 8%
0.6−63 µm (H)
20
Hongo Street
250−2000 µm (L)
Light particles: 3.4 ± 1.0%
12-PAHs (%)
30
40
17-Nov-03 10-Feb-04
100
50
Hongo Street
Shakujii
80
0.6−63 µm (L)
40
100
0 0.6−63 µm (L)
106−250 µm (H)
63−106 µm (H)
0 0.6−63 µm (H)
0 106−1250 µm (L)
10 63−106 µm (L)
10
0.6−63 µm (L)
Deposition mass (%) (a)
20
106−250 µm (L)
20
30
106−1250 µm (L)
17-Nov-03 10-Feb-04
17-Nov-03 10-Feb-04
Light particles: Nov: 28 ± 10% Feb: 33 ± 3%
0.6−63 µm (L)
30
40
63−106 µm (L)
Light particles: Nov: 4.0 ± 1.4% Feb: 0.69 ± 0.03%
Shakujii
63−106 µm (L)
40
50
Shakujii
12-PAHs (%)
50
0.6−63 µm (L)
Deposition mass (%)
60
63−106 µm (L)
268
Figure 2 Mass distribution (a), total 12-PAH distribution (b), and total 12-PAH content (c) in road dust by size and density fractions (mean7SE). L, light particles; H, heavy particles. From Murakami M, Nakajima F, and Furumai H (2005) Size- and density-distributions and sources of polycyclic aromatic hydrocarbons in urban road dust. Chemosphere 61: 783–791.
groundwater pollution, we have to understand the physicochemical characteristics of urban road dust and sediments in infiltration facilities. Heavy metal has higher concern of contamination because of higher exchangeability and leaching potential than PAHs. Boller (1997, 2004) described that heavy metals accumulate in various environmental components such as the sediments in the receiving via sewer system waters, whereas infiltration facilities functioned to control the accumulation of heavy metals as a short-term measure. Accumulation and potential release of heavy metals were investigated in infiltration facilities installed in Tokyo (Aryal et al., 2006, 2007). It is necessary to evaluate the multifunctions of infiltration facilities such as inundation control, groundwater recharge, and pollutant retention. Sixteen soakaway sediments, whose depths ranged widely, were then collected from bottom to top with a plastic pipe from the four sublocations in April 2004. In soakaway, with a 410 cm depth, 10 cm of surface of sediment was collected. Five road dust and two soils in pervious areas were collected in the same area in September 2004. Five road dust from heavily used roads (hereinafter referred to as heavy traffic road dust), where traffic volume ranged from 17 030 to 36 666 vehicles per day, was collected in Bunkyo Ward, Tokyo, Japan, in November 2004. Road dust was collected using a Hitachi CV-100S6 vacuum cleaner from a road gutter. Dry weather periods were more than 1.5 days for preliminary investigation and
sampling of soakaway sediments and more than 4 days for sampling of road dust and soils. Table 3 shows heavy metal contents in soakaway sediments, road dust from heavy traffic road dust and the residential area, and soils in pervious area. Heavy metal contents in thick and thin soakaway sediments are separately shown in the table. Aluminum, Mn, Fe, and As contents in heavy traffic road dust were lower than those in soils in pervious area. On the contrary, Cr, Ni, Cu, Zn, Cd, and Pb contents in heavy traffic road dust were approximately 2–8 times as high as those in soils in pervious areas. These heavy metals, except Cu in road dust from the residential area, were higher than soils in pervious area and lower than heavy traffic road dust. It was revealed that these heavy metals were derived from urban traffic activities. In particular, Pb has remained the key pollutant in heavy traffic road dust even since the early 1980s, when leaded gasoline was not a major source of Pb in aerosol in Japan (Mukai et al., 1993). Nevertheless, the sources of Pb in aerosol and road dust are not still clearly known (Mukai et al., 1993; Shinya et al., 2006). Further investigations are needed to apportion the sources of Pb. Chromium, Ni, Cu, Zn, Cd, and Pb contents in sediments, dust, and soils are plotted in Figure 3. It is interesting to note that thick soakaway sediments contained Cr, Cd, and Pb at significantly higher levels than thin soakaway sediments. In particular, thick soakaway sediments contained Cd and Pb at
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens Table 3
269
Heavy metal contents in soakaway sediments, road dust and soils in pervious area (mean7SE)
Al (g kg1) Cr (mg kg1) Mn (mg kg1) Fe (g kg1) Ni (mg kg1) Cu (mg kg1) Zn (mg kg1) As (mg kg1) Cd (mg kg1) Pb (mg kg1)
Thick soakaway sedimenta n¼5
Thin soakaway sedimentb n ¼ 11
Heavy traffic road dust n¼5
Road dust from the residential area n¼5
n¼2
5174 9776c 780760 4372 5474 4007180 17007100 1171 2.370.2c 230720c
4874 6675c 910780 4673 4973 210740 12007200 1071 1.570.1c 140720c
2674 180710 780780 5375 9073 730780 16007100 7.971.1 1.570.2 180720
3272 7077 700780 4075 5275 120720 10007300 7.470.6 0.9470.21 63718
8071 5270 120070 6873 4472 16070 240730 9.870.7 0.4570.10 2473
Soil in pervious area
a
Soakaway sediment whose depth was Z8 cm. Soakaway sediment whose depth waso8 cm. c Significant difference between thick and thin soakaway sediments (p o 0.05). n, number of samples. b
100
0
0
4000
0
4
3000 2000 1000 0
400 Cd Content (mg kg−1)
Zn
Content (mg kg−1)
3 2 1
Pb 300 200 100 0
Heavy traffic road dust Road dust from the residential area Soil in pervious area
Heavy traffic road dust Road dust from the residential area Soil in pervious area
Thin soakaway sediment
Thick soakaway sediment
0 Thin soakaway sediment
Content (mg kg−1)
20
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50
40
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100
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Thick soakaway sediment
150
80
Content (mg kg−1)
Cr
1500 Ni
Thick soakaway sediment
200
Content (mg kg−1)
Content (mg kg−1)
250
Figure 3 Chromium, Ni, Cu, Zn, Cd, and Pb contents in soakaway sediments, road dust, and soils in pervious area. (Thick soakaway sediment: sediment depth was Z8 cm; thin soakaway sediment: sediment depth was o8 cm.)
higher levels than road dust from the residential area and soils in pervious areas. Cadmium and Pb contents in thick soakaway sediments were higher than or equal to those in heavy traffic road dust. This indicated that soakaway sediments work
as adsorbents to retain heavy metals. Dissolved metals in runoff water possibly have been adsorbed on the soakaway sediments for the long-term operation. The clogging with deposited sediments in infiltration facilities causes the low
270
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens Cr (count) 1
10
100
10 000
1000 1
100
1
10
( ( (
1 Only Cr
1
10
n = 14) n = 2) n = 2)
100 10
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Pb (count)
1000
Heavy traffic road dust
Both Cr and Pb
(n = 40) (n = 9) (n = 4)
Only Pb
1000 10 000 10 000 Road dust from the residential area Soakaway sediment Yellow road line marking (raw material) Yellow road line marking (curb) Yellow road line marking (road surface) 1
(n = 9) (n = 7) (n = 6)
100 1000 10 000 Cr (count)
Figure 4 The X-ray intensities of Cr and Pb during WDS measurement of the identified particles in heavy traffic road dust, road dust from the residential area, soakaway sediment, and yellow road line markings. Only Cr: particles containing Cr but not Pb; only Pb: particles containing Pb but not Cr; both Cr and Pb: particle containing both Cr and Pb. From Murakami M, Nakajima F, Furumai H, Tomiyasu B, and Owari M (2007) Identification of particles containing chromium and lead in road dust and soakaway sediment by electron probe microanalyser. Chemosphere 67(10): 2000–2010.
infiltration rate. The longer contact time in runoff infiltration through thick soakaway might cause the accumulation of heavy metals at higher levels than those through thin soakaway sediments. At the same time, it is of concern that limited adsorption of heavy metals on soakaway sediment finally causes groundwater contamination. It will be necessary to find out the breakthrough characteristics of heavy metals accumulated on sediments to prevent groundwater contamination through infiltration facilities. It was found that thick soakaway sediments (Z8 cm) contained traffic-related heavy metals, such as Cr, Cd, and Pb, at significantly higher levels than in road dust and thin soakaway sediments (o8 cm), possibly due to adsorption phenomena. Adsorption and desorption of heavy metals are related to their speciation, which differs among source materials as well as storm-water characteristics such as pH. It is important to gain an understanding of both the source and the fate of heavy metals in infiltration facilities.
4.09.2.4 Identification of Elements in Individual Particles by Electron Probe Microanalysis Electron probe microanalysis (EPMA) has been used to identify individual particles and to find the sources and their carrier particles. EPMA measurement has advantages of being able to target an individual specific particle as well as to determine multiple elements. The use of EPMA is practicable for source apportionment and investigation of carrier particles for Cr and Pb in road dust and soakaway sediment, considering that these metals were widely distributed in urban areas and were adsorbed by soakaway sediment. In most previous studies, EPMA has been applied to identify one-by-one particles comprising aerosol or road dust. To apportion the sources of a specific heavy metal, such as Cr and Pb, it is necessary to distinguish very minor individual particles containing the heavy metal at significantly high levels from other
particles in aerosol or dust by wavelength dispersive spectrometry (WDS) map analysis. We collected road dust and soakaway sediment in Tokyo and applied them to EPMA analysis to distinguish individual particles containing Cr and Pb at significantly high levels. Figure 4 shows the X-ray intensities during WDS measurement of road dust and soakaway sediment as well as yellow road line marking, which has been known to contain lead and chromium at high levels. Analysis of variance (ANOVA) testing showed that the X-ray intensities of Pb in the identified particles were significantly higher in heavy traffic road dust than in road dust from the residential area and soakaway sediment, whereas no significant difference was observed for the X-ray intensities of Cr in the identified particles among heavy traffic road dust, road dust from the residential area, and soakaway sediment. This indicates that particles containing Pb at high levels were more common in heavy traffic road dust than in soakaway sediment at the individual particle level, whereas the Pb content in soakaway sediment (340710 mgPb kg1) was approximately twice as high as that in heavy traffic road dust (17070 mgPb kg1) at the conglomerate level. In addition, Welch’s test in heavy traffic road dust showed that the X-ray intensities of Pb in the identified particles containing both Cr and Pb were significantly higher than those in the identified particles containing high levels of Pb only, whereas the X-ray intensities of Cr in the identified particles containing both Cr and Pb were lower than those in the identified particles containing only Cr. These results reveal that heavy traffic road dust contains Pb source materials that also contain Cr, and that different source materials contain only Cr at high levels. Figure 4 also shows the results of three yellow road line marking samples. The plots of X-ray intensities of Cr versus Pb were almost linear (1:1) for the identified particles containing both Cr and Pb in heavy traffic road dust. The line was a close fit with the plots of three yellow road line markings. One identified particle
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens
containing both Cr and Pb, which deviated from other points, may have originated in a different source material.
pollution level. It is reported that the EF more than 10 implies the remarkable anthropogenic pollution (Han et al., 2005):
4.09.2.5 Leaching Potential of Heavy Metal in Road Dust
Leached concentration (µg l−1)
1000
100
10
EF ¼
ðX=AlÞsample ðX=AlÞref
where X denotes metal content, Al is aluminum content, and ref is reference natural soil. Figure 6 shows the relationship between EF and leached concentration of these metals. It is clear that anthropogenic pollution is significant for Pb and Cr. Regarding Cr, all the leached concentration from heavy traffic road dust was higher than the 10% value of cleanup criteria, while that of road dust from residential area was lower than the 10% value. Although there was high anthropogenic pollution in Pb, their leached concentration was not so high compared to the cleanup criteria, except for two samples. It is also shown that leaching of Cd is not so significant. On the contrary, attention is needed for leaching of As, while their anthropogenic pollution level was low.
4.09.2.6 Runoff Behavior of Particle-Associated PAH The monitoring of urban runoff often requires highly sophisticated equipment and human organizations to catch rainfall events whenever they happen. Runoff water quality varies with rainfall patterns and antecedent pollutant deposition conditions. Due to the difficulties inherent in regular monitoring of urban runoff, mathematical models are utilized to simulate rainwater runoff and pollutant transportation. Such models are useful in evaluating the effectiveness of pollution-control measures in protecting the water environment from nonpoint source pollution. Runoff models for suspended solids (SSs) have been developed by many researchers (Sartor and Boyd, 1972; Tomonvic and Makishimovic, 1996; Furumai et al., 2001; Hijioka et al., 2001; Uchimura et al., 1997). These models can 1000 Leached concentration (µg l−1)
Surface sediments found in infiltration facilities are known to have high heavy metal content. While sediments in infiltration facilities function to accumulate heavy metals as ‘sinks’, sediments that have high heavy metal contents are possible ‘sources’ to aquatic environments along with desorption processes. Infiltration facilities have the potential to serve as both ‘sinks’ and ‘sources’ of urban nonpoint pollutants during the process of groundwater replenishment by storm water. Leaching tests on road dust are necessary to clarify the transport of heavy metals in water cycle in urban areas. Figure 5 shows the result of leaching tests (liquid/solid ratio was 10 l/ kg-dry; pH ¼ 5.8–6.3; 6 h under room temperature) conducted to evaluate the leaching characteristics of heavy metals from road dust in a heavy traffic area and in a residential area in Tokyo. The results show that the leached concentrations of Cr, Fe, Ni, and Cu from road dust in the heavy traffic area were significantly higher than those from road dust in the residential area. Additionally, the leached fractions of Cr from road dust in the heavy traffic area were also significantly larger than those from road dust in the residential area. The leached Cr from road dust seems to be derived from traffic sign markers such as yellow paint. There were different tendencies in leaching characteristics from fine to coarse fractions among the heavy metals. Leaching tests on size-fractionated road dust revealed that the leached concentrations of Al, Cr, Cu, As, and Cd were higher in their fine fractions (o106 mm), whereas the leached concentrations of Mn, Zn, and Pb were higher in their coarse fractions (106–2000 mm). Among the tested heavy metals, Cr, As, Cd, and Pb are listed as toxic parameters of drinking water quality standard and Soil Contamination Countermeasures Act in Japan. Since heavy metals are contained in natural soil, it is not enough to discuss the anthropogenic pollution by their contents. Therefore, the following equation of EF was proposed by Zoller et al. (1974), which can be used to evaluate the anthropogenic
271
Cr(VI ) 100 As, Cd, Pb 10 Cr(VI)
1
As, Cd, Pb N.D. 10 Enrichment factor
1 1 Road dust in residential area Heavy traffic road dust
Cr
Cr* Mn Fe* Ni* Cu* Zn
Cd
Pb
Road dust in residential area
N.D. Al
As
As
Cd
Pb
Figure 5 Leached concentration of heavy metals from road dust (Murakami et al., 2006). Asterisks refer to significantly higher from heavy traffic road dust.
Cr
As
100
Cd
Pb
Heavy traffic road dust
Figure 6 Relationship between enrichment factor (EF) and leached concentration of Cr, As, Cd, and Pb (Murakami et al., 2006). Solid and dashed lines represent the cleanup criteria in the Soil Contamination Countermeasures Act and those 10% values, respectively.
272
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens
be utilized in the simulation of particle-associated pollutants such as PAHs. The runoff of particle-associated micropollutants is thought to depend on particle size distribution. Several field surveys have shown that micropollutants are attached to fine sediments and particles (Sansalone and Buchberger, 1997; Roger et al., 1998; Murakami et al., 2005; Sartor and Boyd, 1972). It has been reported that the runoff and sedimentation characteristics of fine particles are different from those in the coarse fraction (Furumai et al., 2002; Brenner et al., 2002; Roger et al., 1998; Tomonvic and Makishimovic, 1996; Andral et al., 1999). Therefore, the SS runoff models need a modification of particle categorization to extend their application to PAH runoff models. Urban surface category is also an important factor in runoff modeling. Hijioka et al. (2001) proposed a SS runoff model with two particle size categories and two urban impervious surface types: fine (smaller than 45 mm) and coarse (larger than 45 mm) particles on roads and roofs. In the model, roof runoff was characterized as a faster process than road runoff because roofs have steeper slopes and smoother surfaces. As shown in the previous section, the PAH composition varies according to emission source. Murakami et al. (2003) revealed differences in PAH compositions between roof dust and road dust. The result of cluster analysis on PAH profiles in
the size-fractionated dust showed that the roof dust formed a separate cluster to the road dust, irrespective of either particle size or roof structure. Factor analysis revealed that phenanthrene, indeno(1,2,3-cd) pyrene, and benzo(ghi)perylene were important PAHs for distinguishing the road dust and the roof dust. The result of the factor analysis also suggested that the contribution of tires, pavements, or asphalts to PAHs was greater in road dust than in roof dust, and that the contribution of vehicle exhaust emission to PAHs was greater in roof dust than in road dust. A nonparametric test indicated that the PAH content was higher in the fine dust (smaller than 106 mm) than in the coarse dust (larger than 106 mm). A model was developed by Murakami et al. (2004), explaining the dynamic runoff behavior of particle-associated PAHs. In the model, roads and roofs were considered separately as impervious surfaces, and particle sizes were classified into fine and coarse fractions (Figure 7). A field survey for model development was conducted in a densely populated area in Japan. Consideration of two types of road dust with different mobility is conceptually useful to explain the PAH profiles in runoff particles. Such a model scheme achieves good agreement with observations of SS and PAH runoff behavior for fine particles, except during heavy rainfall. To improve the disagreement, it may be necessary to take account of
400
0 ‘Road’ (PAHs) ‘Roof’ (PAHs) Observed value (PAHs)
200
100
10
3rd phase
2nd phase
Fine particles
1st phase
PAHs (µg s−1)
300
20
30
Rainfall intensity (mm h−1)
Rainfall intensity
40 5:50
2:50
23:50
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Time 0
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20
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80
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PAHs (µg s−1)
Coarse particles
‘Road’ (PAHs) ‘Roof’ (PAHs) Observed value (PAHs)
120
Time Figure 7 Simulated and observed runoff of particle-associated PAH in residential area (Murakami et al., 2004).
Rainfall intensity (mm h−1)
Rainfall intensity
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens
additional sources of SSs and PAHs washed off by heavy rainfall.
4.09.3 Pathogenic Pollution in a Seaside Park after CSO CSOs have been recognized as a serious source of environmental water pollution. However, very limited field surveys have been conducted to evaluate the magnitude of the problem and the duration of its impact on receiving waters in Japan, where rather conventional and less informative water quality parameters have been monitored to assess risk of infection. Therefore, long-term monitoring data and high-quality information are required to estimate the impact of CSO events on human health. Studies in the United States investigated the relationship between waterborne infectious diseases and rain events using public records from 1971 to 1994 (Rose et al., 2000, 2001). They concluded that 20–40% of disease instances were due to contamination during heavy rain events, suggesting that public health could be improved by upgrading rainwater management systems in urban areas. The conventional bacterial parameters, such as total and fecal coliforms, are good indicators of fecal contamination; however, they are less reliable as an index of viral or protozoan contamination because these pathogens behave differently in the water environment. To assess the risk of infection caused by CSO events, both traditional bacterial indicators and viruses should be monitored in the receiving water bodies. Katayama et al. (2004) conducted spatial and temporal monitoring to investigate the fate of pathogens and indicator bacteria in the coastal area in Tokyo. They evaluated the magnitude and duration of the impact of CSO events spatially and temporally after rain events. In particular, the fate of enteric viruses that cause gastroenteritis (noroviruses G1 and G2 and enteroviruses) was investigated. The fates of the viruses as well as conventional indicators were evaluated by serially
monitoring the receiving water body after CSO events. However, their monitoring data are not sufficient to discuss difference behaviors of bacteria and virus after CSO event as well as under normally fine weather conditions. Therefore, we conducted a 2-month survey to evaluate the effects of rainfall on the fate of human adenoviruses, total coliforms and E. coli in coastal water in the Odaiba area in Tokyo Bay (Haramoto et al., 2006). The Odaiba area is suspected to be contaminated with the effluents from several domestic wastewater treatment plants. The Odaiba Seaside Park is located near the sampling site, and more than 1 million people visit the park for recreational purposes annually. Although playing on the beaches is allowed, swimming in the sea is prohibited. To compare the behavior of viruses and indicators, the duration and frequency of sampling were prioritized using only one sampling site and determining limited parameters. The sampling point at Odaiba Seaside Park is shown in Figure 8. Samples were collected from 4 August to 15 October 2004. During the survey period, a total of 774 mm of rainfall was observed, including some heavy rainfall events caused by typhoons. Samples were usually collected in the morning, delivered to the laboratory within a few hours on ice, and analyzed for human adenoviruses, total coliforms, and E. coli. Total coliforms and E. coli in 10 ml of coastal water were determined by an m-Coliblue broth membrane filtration procedure (Millipore, Tokyo). The acid rinse method (Katayama et al., 2002) was used for concentrating the virus from 1000 ml of coastal water samples as in the previously described study. Viral DNA was extracted using commercially available methods, followed by polymerase chain reaction to determine the concentration of viral genomes. At the same time, a decimal dilution series of DNA from human adenovirus serotype 40 were used to create a calibration curve. Total coliforms and E. coli were detected in all 47 tested samples with geometric mean concentrations of 68 CFU ml1 (range: 1.8–3700 CFU ml1) and 4.4 CFU ml1 (range: 0.15– 280 CFU ml1), respectively. On the other hand, human
Sampling point
Odaiba Seaside Park
Tokyo Bay
Figure 8 Sampling points in Tokyo Bay.
250 m
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Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens Rainfall
Human adenoviruses
Total coliforms
E. coli 250
Human adenoviruses, total coliforms, or E. coli (PDU ml−1 or CFU ml−1)
10000 1000
200 100 10
150
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Rainfall (mm d−1)
274
50 0.001 0.0001 10/09
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adenoviruses were detected in 38 (81%) of 47 samples at a maximum concentration of 5.5 PDU ml1. The daily change in the concentrations of human adenoviruses, total coliforms, and E. coli in coastal water is shown in Figure 9 (from 23 August to 10 September). The concentrations of these microorganisms increased after rainfall events. For instance, following a heavy rainfall from 4 to 5 September (84.5 mm), the concentration of human adenoviruses, total coliforms, and E. coli increased from 0.14 to 5.5 PDU ml1, from 13 to 240 CFU ml1, and from 2.0 to 55 CFU ml1, respectively. These increased concentrations decreased gradually to the level before the rainfall event within a few days to 0.24 PDU ml1, 21 CFU ml1, and 1.9 CFU ml1, respectively, on 8 September. The relationship between the concentration of human adenoviruses and that of total coliforms or E. coli was determined. As shown in Figure 10, a moderate positive correlation (r ¼ 0.536) was observed between the logarithms of the concentration of human adenoviruses and that of E. coli among the adenovirus-positive samples. Human adenoviruses were chosen as a target virus, and total coliforms and E. coli were used as indicator bacteria to compare the fates of viruses and bacteria in CSO-contaminated coastal water. After a rainfall event, the concentration of the tested microorganisms in coastal water usually increased by 10–100 times, followed by a gradual decrease to the level before the rainfall event within a few days. It suggests that E. coli could be used as an indicator of human adenovirus contamination in coastal water susceptible to CSO. Total coliforms and E. coli are present in the feces of both humans and animals, while human adenoviruses are present only in human feces. The results of this study indicate that untreated sewage, or CSO, could be a major source of these microorganisms in Tokyo Bay. Interestingly, this monitoring study indicated that the high contamination of human adenoviruses following heavy rainfall persisted for at least a few days after the event. Accordingly, recreational activities in the
Human adenoviruses (PDU ml−1)
Figure 9 Occurrence of human adenoviruses, total coliforms and E. coli in coastal water (from 23 August to 10 September 2004) (Haramoto et al., 2006).
1 r = 0.536 0.1 0.01 0.001 0.0001 0.1
10
10
100
1000
10 000
E. coli (CFU ml−1) Figure 10 Relationships between concentrations of human adenoviruses and E. coli in coastal water samples from the same location on several days (Haramoto et al., 2006).
contaminated area after rain will pose a higher risk of infection by human adenoviruses.
4.09.4 Summary Urban road dust, as a representative pollutant on urban surfaces, was well characterized using several analytical and statistical methods. Source and distribution of PAHs and heavy metals in road dust were discussed referring to research papers on characterization of urban road dust. The desorption process of heavy metals was also reported, looking at their accumulation process in infiltration facilities for urban inundation control. The heavy metal desorption/adsorption phenomena in the soil collected near the infiltration facilities were further assessed. EPMA is a powerful method to distinguish individual particles containing heavy metals at significantly high levels in
Urban Nonpoint Source Pollution Focusing on Micropollutants and Pathogens
road dust and soakaway sediment. This type of element composition analysis could be applied to more environmental samples in order to enable source apportionment and investigation of the carrier particles. Those research results greatly enhance our understandings of urban nonpoint pollutant behavior in infiltration process, although the interaction between pollutants and soils seems dependent on soil characteristics. In addition to urban nonpoint source pollution, CSOs have been recognized as one of the serious sources of pollution to the water environment during rain events. This chapter focuses on pathogenic pollution after CSO events, which is closely related to human health risk at bathing and water amenity activities. The intensive monitoring work indicated that concentration of tested microorganisms increased after a rainfall event and then gradually decreased to the level before the rainfall event within a few days. There was no significant difference in the concentration of the tested microorganisms between the increasing and decreasing tide groups, suggesting that the impact of rainfall on the fate of microorganisms is stronger than that of tidal movement. The behaviors of a pathogenic virus as well as bacterial indicators were gradually elucidated by the long-term sampling. In addition, techniques of molecular biology and virus concentration enabled us to study the fate of human pathogenic viruses in a CSO-impacted water body. This chapter summarizes the research results in the research work related to wet weather pollution phenomena in urban area, consisting of urban nonpoint source pollution and CSO. We should pay more attention to irregularly occurring events as well as to stable discharge of pollutant load so that we could protect water environment from urban pollution sources. It means that sustainable urban water environment can be developed by sufficient understanding and scientific diagnosis of current situation. Then it is necessary to propose and prepare effective countermeasures and their implementation toward the sustainability of urban water environment, coexisting with sound urban activities.
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4.10 Constructed Wetlands and Waste Stabilization Ponds C Polprasert, Thammasat University, Pathumthani, Thailand S Kittipongvises, Asian Institute of Technology, Pathumthani, Thailand & 2011 Elsevier B.V. All rights reserved.
4.10.1 4.10.1.1 4.10.1.2 4.10.1.3 4.10.2 4.10.2.1 4.10.2.2 4.10.2.2.1 4.10.2.2.2 4.10.2.3 4.10.2.3.1 4.10.2.3.2 4.10.2.3.3 4.10.2.3.4 4.10.2.3.5 4.10.3 4.10.3.1 4.10.3.1.1 4.10.3.1.2 4.10.3.1.3 4.10.3.2 4.10.3.2.1 4.10.3.2.2 4.10.3.2.3 4.10.4 4.10.4.1 4.10.4.1.1 4.10.4.1.2 4.10.4.1.3 4.10.4.2 4.10.4.2.1 4.10.4.2.2 4.10.5 4.10.5.1 4.10.5.1.1 4.10.5.1.2 4.10.5.2 4.10.5.3 4.10.6 References
Introduction Millenium Development Goals, Water Supply and Sanitation Problems Potentials of Natural Systems Objectives and Scope Principles of CW and WSP Systems for Wastewater Treatment and Reuse Symbiotic Reactions among Aquatic Plants, Algae, and Bacteria Types and Functions of CW and WSP Systems Constructed wetlands types Wastewater stabilization pond types Major Mechanisms for Wastewater Treatment and Recycling in CW and WSP Systems BOD5 removal SS removal N and P removal Heavy metals Pathogen removal Design Criteria and Operation of CW and WSP Systems CW Design Criteria and Operation Empirical approach Rationale approach Operation, maintenance, and troubleshooting of CW WSP Design Criteria and Operation Empirical approach Rationale approach Operation, maintenance, and troubleshooting for WSP Case Studies of CW and WSP CW Case Studies CW treatment of an industrial wastewater CW treatment of a municipal wastewater CW treatment of fecal sludge or septage WSP Case Studies WSP treatment of industrial wastewater and fish production WSP treatment of wastewater containing endocrine disrupting chemicals Emerging Environmental Issues versus Potentials of CW and WSP Food Production Biomass production in CW Biomass production in WSP Energy and Water Conversation and Climate Change Mitigation CW and HRAP as Future Life-Support Technology Summary
Nomenclature Ac d Io Is ks kT
Cross-sectional area Dispersion number Amount of visible solar energy penetrating a smooth water surface Saturation light intensity Hydraulic conductivity Reaction rate at temperature T
k20 Pa Qa Qe Qi td Vv /
277 277 277 278 278 279 280 280 280 282 282 282 283 283 283 284 284 284 285 286 287 287 287 288 289 289 289 290 290 291 291 292 292 292 292 293 295 296 296 298
Reaction rate at temperature of 20 1C Algal productivity Average flow rate Effluent flow rate Influent flow rate Area doubling time Void volume Media porosity
277
278
Constructed Wetlands and Waste Stabilization Ponds
4.10.1 Introduction
The United Nations (UN, 2005, 2008) member states and international organizations have committed their actions to achieve the millenium development goals (MDGs), which include the issues of poverty alleviation, sustainable development, and environmental protection. One of the MDGs is to provide water supply and sanitation for all by the year 2025, with an interim target of halving the proportion of people living in extreme poverty or those without adequate water supply and sanitation by the year 2015. To meet the 2025 target, with continued population growth, around 2.9 billion people will need to be provided with improved water supply and about 4.2 billion people will need to be serviced with improved sanitation.
4.10.1.1 Millenium Development Goals, Water Supply and Sanitation Problems Presently, the world population is projected to increase from the current number of 6 billions to 9 billions in 2050 (Figure 1). Urbanization and rapid population growth are the main driving forces of adverse environmental impacts worldwide. Increase in the number of the global population will result in increased energy and food demands and environmental problems, especially with regard to the aspects of water supply and sanitation. Supporting evidence for the above statement is the result of a World Health Organization report (WHO, 2000) which showed that in the year 2000 around 1.1 billion people (about 18% of the world population) did not have access to improved water supply (Figure 2), and 2.4 billions (about 40% of the world population) were without access to any sort of improved sanitation facility (Figure 3). The vast majority of those without access to water supply and sanitation services are in Asia.
4.10.1.2 Potentials of Natural Systems The current situations of water supply and sanitation, as stated in Section 4.10.1.1, suggest that employing conventional technologies alone will not be adequate to manage the increasing quantity of human wastes. This is mainly due to the
World population: 1950−2050 10 9 9 billion
Population (billions)
8 8 billion
7 7 billion
6 6 billion
5 5 billion
4 4 billion
3 3 billion
2 1
2050
2040
2030
2020
2010
2000
1990
1980
1970
1960
1950
0
Year Figure 1 World population increases with human advances in science and technology (US Census Bureau, 2007).
Total unserved: 2.4 billion Total unserved: 1.1 billion
5% 2%
7% 2% Asia
13% Asia
Africa
Africa
Latin America and the Caribbean Europe
28%
Latin America and the Caribbean Europe
63% 80% Figure 2 Distribution of the global population not served with improved water supply, by region (WHO, 2000).
Figure 3 Distribution of the global population not served with improved sanitation, by region (WHO, 2000).
Constructed Wetlands and Waste Stabilization Ponds
high investment and operation costs of conventional technologies such as activated sludge, tricking filter, and sewerage systems, which require highly skilled manpower for maintaining. Natural systems such as constructed wetlands (CWs) and waste stabilization ponds (WSPs) are being considered as a potential alternative technology for wastewater treatment. Because they utilize the photosynthetic reactions of aquatic plants and algae to produce gaseous O2 for bacteria to biodegrade organic matter, the CW and WSP systems are usually less expensive to construct and operate than the conventional treatment systems (Table 1; Stowell et al., 1980). The CW and WSP systems offer several advantages as a wastewater treatment process as they could significantly reduce 5-day biochemical oxygen demand (BOD5), suspended solids (SSs), nitrogen (N), phosphorus (P), as well as metals, trace organics, and pathogens present in wastewaters or sludge (Polprasert, 2007). There is considerable yield of biomass through the production of plants, algae, and fish, which could be utilized as animal feeds or human foods (Figure 4). Some of these biomass materials could be converted to become compost fertilizer and alternative energy sources such as biogas, ethanol, and biodiesel oils (Haag, 2007). By integrating biotechnology and nanotechnology with these natural systems, their potentials in treating wastewaters containing toxic compounds and in resource recovery are expected to be improved considerably (Figure 5; Stern, 2006).
279
Figure 4 Food production from natural systems.
4.10.1.3 Objectives and Scope This chapter describes the potentials of some natural systems such as CW and WSP for wastewater treatment and resource recovery. The principal reactions occurring in the CW and WSP systems and the treatment mechanisms occurring in these natural systems such as photosynthesis, microbial reactions, symbiotic reactions of algae and bacteria and aquatic plants will be explained, including their potentials for resources’ recovery through the production of human and animal foods and alternative energy sources. Performance of full-scale CW
Table 1 Costs and energy requirements of conventional and aquatic treatment systems Item
Treatment plant size 378.5 m3d1
3785 m3d1
Conva Aquaticb Conv 0.71 0.37 Capital cost (US$ 106) 21 O & M cost (US$/year 103) 35 Energy, (kJ/year 109) 0.93 0.53
Aquatic
1.60 0.90 117 74 5.06 2.19
a
Activated sludge þ chlorination. Primary clarification þ artificial wetlands þ chlorination. O & M ¼ operation and maintenance. kJ ¼ kilojoules. Modified from Stowell R, Ludwig R, Colt J, and Tchobanoglous G (1980) Towards the Rational Design of Aquatic Treatment Systems. Paper presented at the ASCE Convention, Portland, OR, USA, 14–18 April. Department of Civil Engineering, University of California, Davis, CA, USA.
b
Figure 5 Integrated natural systems for wastewater treatment: (a) attached-growth WSP and (b) integrated nano-phytotechnology CW.
280
Constructed Wetlands and Waste Stabilization Ponds
and WSP systems in the treatment of municipal and industrial wastewaters under different climatological conditions and additional benefits derived from the production of algae, fish, agricultural products, and energy will be presented. Emerging issues and research topics of CW and WSP are discussed.
matter decomposition by bacteria, respectively:
NH 3 + 7.62O 2 + 2.53H 2 O
CW is a wastewater treatment system consisting of shallow ponds or channels cultured with emerging aquatic plants. The processes by which wastewater is treated include a wide range of interacting biological, physical, and chemical mechanisms. WSPs are shallow man-made basins into which wastewater continuously flows and from which, after a retention time of several days (rather than several hours in conventional treatment processes), a well-treated effluent is discharged. WSP systems comprise a series of anaerobic, facultative, and maturation ponds, or two or more such series in parallel. Anaerobic and facultative ponds are designed primarily for BOD5 removal, while maturation ponds are for SS and pathogen removal; although some BOD5 removal occurs in maturation ponds and some SS and pathogen removal occurs in anaerobic and facultative ponds.
C7.62 H8.06 O 2.53 N
Facultative bacteria
5CO 2 + NH 3 + 2H 2 O
Wastewater Algae CO2
O2
Bacteria
Aerobic zone Fish
Effluent
Facultative zone Anaerobic zone
Figure 6 Schematic diagram of treatment mechanisms in WSP.
Effluent outlet for FWS CW
Slope 1% CW media (sand and gravel) Clay liner or membrane Figure 7 Schematic diagram of horizontal-flow CW.
ð2Þ
Sunlight
Emergent plants Influent wastewater
ð1Þ
Note: In Equation (1) C7.62 H8.06 O2.53 N represents algal cells, and in Equation (2) C5H7O2N represents organic matter. Equations (1) and (2) explain the symbiotic relationships between algae and bacteria in a facultative WSP in which these natural reactions could be utilized in wastewater treatment without the addition of mechanical energy. The algal cells and zooplankton, which graze on them, can be used as food for herbivorous fish (Figure 6). A schematic diagram of the CW process illustrating the symbiotic reactions between aquatic plants and bacteria is shown in Figure 7. Similar to the algal photosynthesis, the aquatic plants perform photosynthesis under sunlight and the produced O2 is transferred from leaves to the root zones where bacteria will utilize it to decompose the organic matter. Other reactions occurring in a CW system such as adsorption and plant uptake will, respectively, result in better removal of SS and nutrients (such as N and P). Depending on climatic conditions, the harvested plant biomass can be used as animal
4.10.2.1 Symbiotic Reactions among Aquatic Plants, Algae, and Bacteria The relationship between the phototrophic micro-algae and bacteria is often classically illustrated as mutual reactions between aerobic bacteria and algae cells. In essence, with sunlight the algae produce oxygen that is used by aerobic and facultative bacteria in organic matter decomposition. The algae benefit by utilizing the CO2 produced by bacterial respiration along with the released nutrients to derive energy and fix carbon for growth via photosynthesis. If C5H7O2N represents organic matter, Equations (1) and (2) represent the photosynthetic oxygen production by algae and organic
Algae
+ 7.62O 2 C5 H7 O 2 N + 5O 2
4.10.2 Principles of CW and WSP Systems for Wastewater Treatment and Reuse
Sunlight
Effluent outlet for SF CW
Constructed Wetlands and Waste Stabilization Ponds
foods or converted to become compost fertilizer. Since there is no need for energy inputs, the operation and maintenance of CW units are cheaper than those of conventional treatment systems (see Table 1).
4.10.2.2 Types and Functions of CW and WSP Systems The CW and WSP systems have been designed and operated in various ways to treat municipal and industrial wastewaters. The variations include physical design, hydraulic flow patterns, and organic loading rates. The major types of these variations are given in the following.
4.10.2.2.1 Constructed wetlands types CWs are often classified into two basic types: free water surface (FWS) in which the water surface is maintained 10–50 cm above the CW bed and subsurface flow (SF) in which the water level is maintained below the CW bed. Examples of emergent aquatic plants growing in CW beds are shown in Table 2. Details of these two CW types and their functions are presented as follows. Free surface flow CW. According to US EPA (2000), FWS CW typically includes one or more shallow basins or channels with a barrier to prevent seepage to sensitive ground waters and a submerged soil layer to support the roots of selected emergent vegetation, inlet and outlet structures to distribute and collect wastewater, control water levels, and maintain hydraulic retention times (HRTs). In FWS CW treatment, wastewater at a relatively shallow depth of 10–50 cm flows over the vegetated soil surface (Figure 8) and the intended flow path through the system is horizontal. BOD5 matter is degraded by the bacteria attached to the surface of the CW media and plant roots using O2, photosynthetically produced by the plant leaves, that is transferred to the soil–water matrix. N and P are removed by plant uptake and sedimentation while, due to long HRT, most SS matters would settle down in the CW beds and some could be Table 2
removed by filtration and adsorption. Due to less exposure of the wastewater to sunlight, pathogen removal by CW systems is usually not adequate. Table 3 shows the functions of aquatic plants in CW treatment systems. Subsurface flow CW. SF CW consists of a set of basins or channels with barrier to prevent seepage, and a suitable depth of porous media, gravel, rock, and different soil, that supports the growth of emerging vegetation (Figure 7). Through proper design of outlet structure, the wastewater level is maintained to be below the CW bed. Wastewater flows horizontally (Figure 7) or vertically (Figure 9) through the medium and is purified during the contact with the surfaces of the medium and the root zone of the vegetation by the physical, chemical, and biological reactions, similar to those occurring in FWS CW. Some of the major differences between FWS and SF CWs are presented in Table 4.
4.10.2.2.2 Wastewater stabilization pond types WSP technology is one of the most important natural methods for wastewater treatment. There are four types of WSPs, namely, anaerobic ponds, facultative ponds, maturation ponds, and high-rate algal ponds. The major features and functions of these ponds are given in Table 5 and are described as follows. Anaerobic ponds (APs). APs are normally employed to treat wastewaters with high organic concentrations. With long HRTs of 10–50 days and depths of 3–5 m, anaerobic bacteria growing in AP could biodegrade the incoming BOD5 and accumulated SS resulting in the production of biogas such as methane (CH4), CO2, and hydrogen sulfide. The important
Emergent plants Wastewater distribution
Emergent aquatic plants for wastewater treatment
Common name, scientific name
Cattail, Typha spp. Common reed, Phragmites communis Rush, Juncus spp.
Bulrush, Scirpus spp.
a
281
Distribution
Common in Southeast Asia Common in America and Europe Common in America and Europe Common in America and Europe
ppt ¼ part per thousand.
Desirable temperature (1 C)
Maximum Optimum salinity pH tolerance (ppt a)
Effluent outlet Slope 1% Rhizome network
Soil, sand or gravel
10–30
30
4–10
10–30
45
2–8
15–25
20
5–7.5
Table 3 systems
4–9
Roots and stems in water column Surfaces for microbial growth Media for filtration and adsorption
Liner
Figure 8 Schematic diagram of horizontal-flow FWS CW.
15–25
20
Functions of emergent aquatic plants in CW treatment
Stems and leaves above water surface Light attenuation and preventing algae growth Minimizing wind effects Transfer of oxygen to root zone
282
Constructed Wetlands and Waste Stabilization Ponds Influent wastewater Emergent plants
Effluent outlet
Slope 1%
CW media (sand and gravel)
Clay liner or membrane
Figure 9 Schematic diagram of vertical-flow SF CW. Table 4 Differences between free water surface and subsurface flow constructed wetland systems Free water surface
Subsurface flow
Typically long, narrow channels with an impermeable liner to prevent seepage With emergent vegetation Wastewater flows horizontally at a shallow water depth and in CW media and is purified by microorganisms attached to plant stalks, litter, and on medium surface
Trench or bed with impermeable liner to prevent seepage
Media: usually soil, sand, and gravel Odor and mosquito problems likely
With emergent vegetation Wastewater flows laterally or vertically through the medium and is purified by microorganisms attached on the surfaces of the root zone of the vegetation and the medium surface (biofilm) Media: sand and gravel Odor and mosquito problems less likely
Modified from US EPA (2000) Free Water Surface Wetlands for Wastewater Treatment: A Technology Assessment Washington, DC: Office for Water Management, US Environmental Protection Agency.
stages of anaerobic degradation are hydrolysis, acid formation, and methane formation, all of which occur in appropriate steps and rates for effective BOD5 reduction. As 50–60% of the biogas is CH4, which has calorific value of 9000 kcal m3, equivalent to 13 kcal g1or 211 kcal g1 mol1 (Polprasert, 2007), the CH4 gas could be captured and utilized as an alternative fuel in heat or electricity generation. Facultative ponds (FPs). Among the different types of WSPs, FP is most commonly used for wastewater treatment. With a HRT of 5–20 days and depths of 1.5–3 m, FP utilizes the algal– bacterial symbiotic reactions in which gaseous O2 photosynthetically produced by the algae is used by the facultative bacteria in biodegrading the influent organic matter. Byproducts of bacterial decomposition, such as CO2 and NH3– N, serve as nutrients for the algae (Figure 6).
As shown in Table 5 the medium organic loading rates of 100–300 kg BOD5 ha1 d1 applied to FP usually result in 80– 90% BOD5 removal. The nutrients such as N and P are also removed by algal uptake and sedimentation to the pond bottom. The FP effluent normally contains algal cells, which could make its SS concentrations and color not suitable for discharge into receiving waters. Therefore, maturation ponds should be used in series to polish the FP effluent and further reduce pathogenic microorganisms. On the other hand, since algal cells are rich in protein and nutrients, FP effluent could be used as feed to fish ponds or irrigation water to grow crops. More details of these applications are provided in a later section. Maturation ponds (MPs). MP receives the effluent from an FP and its size and number depend on the required microbiological and SS quality of the final effluent. MP units are shallow (1–2 m in depth) with less vertical stratification and, due to the low organic loading rates (Table 5), their entire volume is well oxygenated throughout the day. Algal population in MP is much more diverse and less in concentrations because most of them settle down to the pond bottom. The main removal mechanisms, especially of pathogens and fecal coliforms, are ultraviolet (UV) light inactivation, high pH during day time because CO2 is uptaken by algal cells (Equation (1)), grazing by protozoa, and sedimentation with SS to the pond bottom. High-rate algal ponds (HRAPs). HRAP conventionally takes the form of a continuous channel equipped with an aeratormixer to recirculate the contents of the ponds and is used for both production of algal biomass and wastewater treatment. It is characterized by large area/volume ratios, and shallow depths in the range of 0.2–0.6 m to allow sunlight to penetrate the whole pond depth (Table 5). To minimize short-circuiting, baffles are normally installed in the pond to make length/ width ratio of the channel greater than 2/1. A diagram and photograph of HRAP are shown in Figures 10 and 11, respectively. Effluent overflow from HRAP, containing high algal suspension, normally goes into an algal separation unit. The
Constructed Wetlands and Waste Stabilization Ponds Table 5
283
Comparative features of waste stabilization pond
Feature
Anaerobic pond
Facultative pond
Maturation pond
High-rate algal pond
Depth (m) HRT (days) OLR, (kg BOD5 ha1 d1) Microorganisms responsible for BOD5 reduction Major functions
3–5 10–50 High, 4300 Anaerobic bacteria
1.5–3 5–20 Medium 100–300 Facultative bacteria and algae
1–2 5–10 Low, o100 Aerobic bacteria
0.2–0.6 1.5–8 Medium 100–200 Aerobic bacteria and algae
Pretreatment, BOD5 reduction Biogas, CH4, CO2
BOD5 reduction
Polishing, SS and pathogen reduction -
Algal biomass production and BOD5 removal Algal biomass
By-products
SS in forms of algal and bacterial cells
Influent Baffles Effluent to algal separation unit or fish pond
Aerator and mixer
Wastewater flow Figure 10 Schematic plan of high-rate algal pond.
temperature, light intensity, mixing or agitation, pond depth, and HRT. It is generally known that light intensity is the important factor for photosynthesis and therefore, algal production, while temperature influences the biodegradation rate of the organic matter. Under appropriate HRT and organic loading rate (Table 5), the average production of algal biomass in HRAP is 70 or 35 tonnes ha1 yr1 of algal protein (algal cells contain about 50% protein). These algal cells could be used as fish or animal foods and as biomass to produce biofuels (Polprasert, 2007).
4.10.2.3 Major Mechanisms for Wastewater Treatment and Recycling in CW and WSP Systems As stated in the previous section, there are several mechanisms occurring in CW and WSP systems responsible for wastewater treatment. These mechanisms are physical, chemical, and/or biological, which may take place individually or in combination in removing certain pollutants from the influent wastewater. Although there are emergent plants growing in CW and algae cells growing in WSP, the treatment mechanisms of these two natural systems are similar.
4.10.2.3.1 BOD5 removal
Figure 11 High-rate algal pond at Royal Chitrlada Palace, Bangkok, Thailand.
effluent obtained after the algae have been separated is expected to have BOD5 of 20 mg l1 and DO of 0.5 mg l1. Because of these advantages, HRAP has, in recent years, received increasing attention as a means of both treating wastewater and producing algal biomass for food or energy production. Factors affecting the performance of HRAP and algal production include available carbon and nutrient sources,
As stated in the previous section, BOD5, chemical oxygen demand (COD), or organic matter removal in CW and WSP units is due mainly to the biodegradation reactions of bacteria as shown in Equation (2) using O2 photosynthetically produced by the emergent plant and algal cells, respectively. Polprasert et al. (1998) found biofilm bacteria growing on surfaces of the media and root zones in the CW beds to be the major organisms responsible for BOD5 removal. For the WSP systems, both suspended bacteria growing in the pond water and biofilm bacteria growing on the pond’s side walls were found to be responsible for BOD5 removal (Polprasert and Agarwalla, 1994). By-products of the bacterial reactions such as CO2 and NH3 are used by the emerging plants and algae cells for photosynthesis under sunlight. Since photosynthetic and bacterial reactions are dependent on climatological conditions, tropical areas having both plenty of sunlight and high temperature would be ideal for the CW and WSP systems to function effectively.
4.10.2.3.2 SS removal SS removal is very effective in both types of CWs and is accomplished mainly through adsorption on the media’s surface
284
Constructed Wetlands and Waste Stabilization Ponds
and sedimentation. Controlled dispersion of the influent flow to CW units with proper diffuser pipe design can help to ensure low velocities for SS removal. Since WSP systems are operated at long HRT, most SSs would settle down to the pond bottom, while some may be adsorbed on the algal cells, forming larger aggregates and be removed by sedimentation. In general, high algal growth in FP results in high SS concentrations in the FP effluent, which can be polished further in MP where, under low organic loading and long HRT, there is less algal growth, algal grazing by protozoa, and sedimentation, all leading to low SS concentrations in the MP effluent.
Table 6 Treatment mechanisms of constructed wetland and waste stabilization pond systems Treatment mechanisms
Constructed wetland
Waste stabilization pond
BOD5 removal
Emerging plantsbacterial symbiotic reactions (effective)a Filtration, adsorption, some sedimentation (effective)a Plant uptake, adsorption, nitrification/ denitrification (effective)a
Algal-bacterial symbiotic reactions (effective)a
SS removal
N and P removal
4.10.2.3.3 N and P removal Being nutrients, N and P are uptaken by emergent plants and algal cells for their growth and, consequently, removed from the influent wastewater. The range of N and P removal in CW by plant uptake was reported to be 10–50% (Sawaittayothin and Polprasert, 2007; US EPA, 1988), whereas the percent N and P removal in WSP by algal biomass uptake was 80–90% and less than 50%, respectively (Pano and Middlebrooks, 1982; Silva et al., 1995; Mara et al., 1992). Nitrification and denitrification can be significant reactions responsible for N removal in CW and WSP systems if suitable environmental conditions conducive to the growth of nitrifying bacteria (oxic conditon) and denitrifying bacteria (anoxic condition) are maintained. N removal via nitrification and denitrification reactions in CW and WSP systems was reported to be 20–40% and 10–30%, respectively (Sawaittayothin and Polprasert, 2007; WEF, 2001). P removal in CW systems is due mainly to adsorption on the CW media and sedimentation and was found to be about 50%. Improved P removal efficiencies of more than 90% could be achieved when oyster shells were used as CW media to adsorb P (Park and Polprasert, 2008). P removal in WSPs is not effective and is in the low removal range of 20–30% (WEF, 2001).
4.10.2.3.4 Heavy metals The predominant removal mechanisms of heavy metals in CW are precipitation and adsorption. Precipitation is enhanced by increased pH in the CW system (Equation (1)), which occurs during photosynthesis. The large surface areas of CW media and root zones are also effective for heavy metal adsorption. Removals of Cu, Zn, and Cd at 99%, 97%, and 99%, respectively, for CW units operating at HRT of 5.5 days in Santee, California, have been reported (US EPA, 1988). Visesmanee et al. (2008) found that CW units operating at HRT of 5.8 days could remove more than 99% of Cd through adsorption. Heavy metal removal in WSP is caused mainly by precipitation and was reported to be about 90% (WEF, 2001).
4.10.2.3.5 Pathogen removal Pathogen removal in CW beds is due mainly to adsorption on CW media and natural decay caused by unfavorable conditions. The growth of emergent plants usually prevents UV light from reaching the CW beds and, consequently, from inactivating the pathogens present in the influent wastewater. FP and MP are exposed to sunlight and UV light, which is effective in inactivating pathogens present in the influent
Heavy metals
Plant uptake, adsorption (effective)
Pathogen removal
Adsorption, natural decay (not effective)a
Sedimentation (not effective)a Algal uptake, nitrification/ denitrification, sedimentation (effective)a Algal uptake, sedimentation ( not effective)a UV light, predator grazing, sedimentation, high pH, natural decay (effective)a
a
Denote relative efficiency and may vary from places to places due to climates and operating conditions.
wastewater. Other mechanisms responsible for pathogen dieoffs in WSP are the high pH during algal photosynthesis (Equation (1)), grazing by zooplanktons, sedimentation, and natural decay. In general, pathogen removal in WSP could be 99.99% which is more effective than those of 50–90% obtained in CW systems (WEF, 2001). A comparison of major treatment mechanisms occurring in CW and WSP systems is given in Table 6.
4.10.3 Design Criteria and Operation of CW and WSP Systems This section describes design criteria and environmental requirements for CW and WSP systems. Methods for the design and operation of these systems based on empirical and rationale approaches are given.
4.10.3.1 CW Design Criteria and Operation The presence of emergent plants and media in CW beds, which have effects on the flow characteristics, need to be taken into considerations in the design and operation of CW systems. Depending on climatic conditions, evapotranspiration caused by emergent plants, especially during day time, could result in water loss in the CW effluent more than 50% of the influent. The media size and shape and its porosity have effects on the specific surface area and HRT of the CW systems. In designing a CW unit, average flow rate (Qa) and actual HRT (HRTa) should be used in calculation. The values of Qa can be determined from
Qa ¼
Qi þ Qe 2
ð3Þ
Constructed Wetlands and Waste Stabilization Ponds
where Qi and Qe are the influent and effluent flow rates, respectively, of a CW unit, m3 d1. The values of Qi can be obtained from municipal or industrial sources, which generate the wastewaters. The value of Qe depends on evapotranspiration rates, which can be determined from the data of similar CWs or from agricultural stations; otherwise, actual measurements at some pilot-scale CW units should be done during a summer period. The value of HRTa can be determined from
HRTa ¼
Vf Qa
ð4Þ
where V is the volume of CW bed, m3, that is, length width depth of liquid (L W D) in the CW bed, Vv the void volume in the CW bed, m3, and f the media porosity or Vv/V, decimal. Table 7 shows the approximate porosity values of sand and gravel commonly used as media in CW beds. It should be noted that, due to accumulation of solids and biofilm growth, the media porosity or void fraction values will change with the operation time of CW. For practical purposes, the f value of SF CW could be taken as 0.30 (Table 7). For FWS CW whose liquid depth is maintained 10–60 cm above CW bed, the f values could be taken as 0.70 (WEF, 2001). Due to the complex interactions among the physical, chemical, and biological processes occurring in CW, the available design criteria do not include all of these aspects, but are based on either empirical approach or rationale approach (first-order reaction rate and plug-flow conditions), which are described in the following.
285
North America and Australia. Based on these and other data, a summary of CW design guidelines is shown in Table 9, which should be applicable for CW located in temperate and tropical areas. Depending on wastewater characteristics, the BOD5 and TN loading rates of CW could be higher than those given in Table 9 if they are operated under tropical conditions. Due to better contact between wastewater and media in SF CW units (Figure 9), the maximum hydraulic and BOD5 loading rates can be higher than those of the FWS CW units. As emergent plants can grow and decay in CW beds, vegetation harvesting should be done regularly and more frequently under tropical conditions to maintain satisfactory performance. To avoid media clogging, pretreatment of the influent wastewaters by at least sedimentation would reduce accumulation of settleable solids in the CW beds. SF CW units can be operated in horizontal-flow or verticalflow modes (Figure 12). The cross-sectional area (Ac) of an SF CW unit is calculated from (US EPA, 1988)
Ac ¼ Qa ðks SÞ1
ð5Þ
where Ac is the cross-sectional area of SW CW bed perpendicular to the flow direction, m2, ks the hydraulic conductivity, m3 m2 d1 (Table 7), and S the bed slope, as fraction or decimal, Table 9. US EPA (1988) recommended that the values of ksS or Qa Ac1should be less than 8.6 m d1 to ensure sufficient contact time between the wastewater and media, which is essential for effective wastewater treatment.
4.10.3.1.1 Empirical approach
4.10.3.1.2 Rationale approach
Both FWS and SF CW systems have been applied to treat wastewaters at several locations worldwide with satisfactory results. Table 8 shows performance data of CW located in
As previously mentioned, the complex reactions occurring in CW beds responsible for wastewater treatment have made it difficult to develop a comprehensive model to predict
Table 7
Media characteristics for CW systems
Media type
D10a effective size (mm)
Porosity (f)
Hydraulic conductivityb (ks) (m3 m2 d1)
Coarse sand Gravelly sand Medium gravel
2 8 32
0.28–0.32 0.30–0.35 0.36–0.40
100–1000 500–5000 10 000–50 000
a
D10 ¼ effective size of the media at 10% (cumulative weight of total, or the media size such as 10% by weight are smaller). Hydraulic conductivity ¼ flow rate divided by area perpendicular to flow direction. Modified from WEF 2001. Natural Systems for Wastewater Treatment, Manual of Practice FD–16, 2nd edn. Alexandria, VA: Water Environment Federation.
b
Table 8
Summary of BOD5 and SS removal from CW
Project
Listowel, Ontario Santee, CA Sydney, Australia Arcata, CA Emmitsburge, MD Gustine, CA
Flow (m3 d1)
17 240 11 350 132 3785
Type
BOD5 (mg l1)
SS (mg l1)
Reduction (%)
FWS SF SF FWS SF FWS
60 120 30 40 60 150
110 60 60 40 30 140
82 75 86 64 71 84
10 30 5 13 20 25
8 5 4 30 8 20
Modified from Polprasert (2007) Organic Waste Recycling, Technology and Management, 3rd edn. London: IWA Publishing.
93 90 92 28 73 86
Hydraulic surface loading rate (m3 m2 d1)
900 1540 410
286 Table 9
Constructed Wetlands and Waste Stabilization Ponds Summary of wetland design considerations
Design consideration
Maximum water depth (cm) Bed deptha (cm) Minimum aspect ratiob Bed slopec (%) Minimum hydraulic retention time (day) Maximum hydraulic loading rate (cm d1) Minimum pretreatment Configuration Maximum loading (kg ha1 d1) BOD5 TN Additional considerations
Constructed wetland Free water surface
Subsurface flow
10–60 Not applicable 2:1
Water level below ground surface 30–90 2:1
1–5 5–10
1–5 5–10
2.5–5
6–8
Screening and sedimentation Multiple units in parallel and series
Screening and sedimentation Multiple units in parallel and series
100–110 60 Mosquito control with mosquito fish; vegetation harvesting regularly
80–120 60 Vegetation harvesting regularly
Figure 12 Horizontal and vertical- flow modes of SF CW: (a) horizontalflow wetlands and (b) vertical-flow wetlands.
a
Bed depth is to support plant growth. Aspect ratio is length/width (L/W ), in this case for horizontal-flow CW. c Bed slope is for horizontal-flow mode. Modified from Polprasert (2007) Organic Waste Recycling, Technology and Management, 3rd edn. London: IWA Publishing and WEF (2001). Natural Systems for Wastewater Treatment, Manual of Practice FD–16, 2nd edn. Alexandria, VA: Water Environment Federation.
as shown in the following information:
b
treatment efficiency. Reed et al. (1988) suggested a first-order reaction and plug-flow model in a CW bed as follows:
Ce ¼ e kT t C0
ð6aÞ
where Ce, C0 are the effluent and influent concentrations (g m3), kT the reaction rate (¼ k20 (1.06)T20, T the liquid temperature (1C), and T (¼ HRTa) the actual hydraulic retention time, day. For BOD5 removal, k20 ¼1.104 d1 for SF CW and k20 ¼ 0.687 d1 for FWS CW. The higher k20 value of SF CW was due mainly to better contact between the wastewater and the media or bacteria responsible for BOD5 degradation. However, due to a lower f value of SF CW, the volume or area of SF CW required to achieve the same treatment efficiency is usually higher than that of FWS CW, as shown in Example 1. A study by Polprasert et al. (1998) found the activity of biofilm bacteria growing in CW beds to be much more significant than suspended bacteria in organic or COD removal.
Example 1. Compare the surface areas of an FWS CW unit and an SF CW unit required to achieve the treatment efficiency
Flow rate (Qa) ¼ 760 m3 d1 Influent BOD5, Co ¼ 130 mg l1 Effluent BOD5, Co ¼ 10 mg l1 Liquid temperature ¼ 25 1C Media porosity, f ¼ 0.75, for FWS CW f ¼ 0.3, for SF CW Depth, d ¼ 0.5 m, for SF CW (to support cattail roots) D ¼ 0.7 m, for FWS CW (cattails root depth of 0.5 m þ water depth of 0.2 m). For FWS CW,
k25 ¼ k20 ð1:06Þ2520 ¼ 0:678ð1:06Þ5 ¼ 0:907 For SF CW,
k25 ¼ k20 ð1:06Þ2520 ¼ 1:104ð1:06Þ5 ¼ 1:477 BOD5 removal
C0 ¼ e kT t Ce For FWS CW,
10 ¼ e 0:907t ; tFWS ¼ 2:8 days 130 For SF CW,
10 ¼ e 1:477t ; tSF ¼ 1:7 days 130 t¼
LWDf Vf tQ ¼ ; AFWS or ASF ¼ Qa Qa Df
ð6bÞ
Constructed Wetlands and Waste Stabilization Ponds
where AFWS or ASF are surface areas of FWS CW or SF CW, respectively:
AFWS ¼
2:828 760 ¼ 4093:87 m2 ¼ 0:410 ha 0:7 0:75
•
1:737 760 ¼ 8800:8 m2 ¼ 0:88 ha ASF ¼ 0:5 0:3 AFWS ; L:W ¼ 10:1-205:20 ASF ; L:W ¼ 10:1-300:30
4.10.3.1.3 Operation, maintenance, and troubleshooting of CW Two operational considerations associated with CW for wastewater treatment are mosquito control and plant harvesting, in addition to system perturbations, which can occur from time to time. Mosquito control. Mosquito problems may occur in FWS CW units located in the tropics especially when they are overloaded and under anaerobic conditions. Strategies used to control the mosquito population include effective pretreatment to reduce total organic loading; step feeding of the influent wastewater stream with effective influent distribution and effluent recycle; vegetation management; natural controls, principally by mosquitofish (Gambusia affinis), in conjunction with the above techniques; and application of man-made control agents (or pesticides). In general, natural controls are preferred because of a concern that man-made control agents might develop resistant strains of mosquito (Wieder et al., 1989). Plant harvesting. The frequency of plant harvesting in CW treatment systems depends on several factors, such as climate, plant species, and their growth/decay rates. Regular plant harvesting will minimize the accumulation of decayed biomass, maintaining system performance and reducing congestion at the water surface. Where a segmented wetland system is used, drying each segment separately allows harvesting with conventional equipment (Wieder et al., 1989). The frequency of plant harvesting should be in accordance with the plant’s growth rate. In the tropics and under favorable conditions, the area doubling time of most emergent aquatic plants is 2–3 months (Polprasert, 2007). System perturbations and operation modifications. Perturbations generally are of two types (Girts and Knight, 1989): (1) predictable perturbations, which can be predicted and occur periodically and (2) unpredictable perturbations, which are unforeseen in the design phase or which occur so infrequently that incorporation into the design would entail unnecessary expense. Some of the suggestions to minimize perturbations on CW operation given by the US EPA (2000) are listed below. FWS CW. Routine operation and maintenance (O&M) requirements for FWS CW include hydraulic and water depth control, inlet/outlet structure cleaning, wetland vegetation management, mosquito and vector control, and routine monitoring. The following points for specific treatment performance are suggested.
•
If N or P removal is required, the land area required for FWS CW could be large. The removal of N in biological
•
•
287
processes such as nitrification/denitrification and plant uptake requires longer HRT. The P, metals, and some persistent organics removed by the system are bound in the wetland sediments and accumulate over time. In cold climate areas, low temperatures reduce the rate of BOD5 removal and the biological reactions. An increased HRT can compensate for this, but the increased wetland size required in extremely cold climates may not be cost effective or technically feasible. The bulk water in most FWS CW systems is essentially anoxic, limiting the potential for rapid biological nitrification of ammonia. Increasing the wetland size and, therefore, the HRT, may compensate for this. Mosquitoes and other insect vectors can be a problem. FWS CW systems can remove fecal coliforms by at least one log from typical municipal wastewaters, but may not be sufficient to meet discharge limits in all locations and supplemental disinfection or polishing units may be required. The situation is further complicated because birds and other wildlife in the wetland produce fecal coliforms.
SF CW. Routine O&M requirements for SF CW are similar to those of FWS CW and include hydraulic, water depth control, inlet/outlet structure cleaning, wetland vegetation management, and routine monitoring. The water depth in the SF CW may need periodic adjustment on a seasonal basis or in response to increased resistance over long term from the accumulating detritus in the media void spaces. Mosquito control should not be required for SF CW systems as long as the water level is maintained below the media surface. Vegetation management and other specific treatment performance in SF CW should follow the same procedures indicated for the FWS CW. Due to its low porosity, an SF CW will require larger land area than an FWS CW (see Example 1) and a conventional treatment process.
4.10.3.2 WSP Design Criteria and Operation The mechanisms responsible for wastewater treatment in WSP are complex and similar to those of CW, but in this case it is the algal cells, not emergent plants, growing in the pond water that perform photosynthesis and produce O2 for the bacteria to biodegrade the influent wastewater (Figure 6). The available design criteria of WSP based on empirical and rationale approaches are presented.
4.10.3.2.1 Empirical approach Empirical approach for WSP design is based mainly on relationships between organic loading rates and HRT and the treatment performance of WSPs. In general, the design criteria of AP, FP, MP, and HRAP could follow those stated in Table 5 from which satisfactory results of the treatment performance could be expected.
4.10.3.2.2 Rationale approach Since FP is the major type of pond effective for organic (BOD5 or COD) removal, the formula for completely mixed and assuming first-order kinetics is commonly used in FP design for
288
Constructed Wetlands and Waste Stabilization Ponds
both organic and fecal coliform reductions (Marais, 1974):
Ce 1 ¼ C0 1 þ kT t
and Bhattarai (1985) proposed a dispersion model for WSP:
d¼
ð7Þ
ð8Þ
The value of kT for fecal coliform reduction is given by
kT ¼ 3:6ð1:19ÞT20
ð9Þ
The terms Ce, Co, T, and t are as defined previously. In general, the value of t depends on the magnitude of treatment required (Ce/Co) and the calculated t value is used to determine the pond volume or the pond surface area if a pond depth is chosen (see Table 5). To achieve maximum fecal coliform die-off, Marais (1974) recommended that ponds (FP or MP) should be of the same size and laid out in series. In this case, Equation (10) is applicable for designing FP and MP for fecal coliform die-off:
Ce 1 ¼ C0 ð1 þ kT tÞn
Example 2. Design a facultative pond (FP) to treat an industrial wastewater flow of 1000 m3 d1 with a soluble BOD5 of 500 mg l1. Assume the following conditions apply: 1. 2. 3. 4. 5. 6.
ð10Þ
where n is the number of ponds of identical size in series. Other terms are as defined previously. Since the hydraulic conditions in WSP are neither completely mixed nor plug flow, but partially mixed, Wehner and Wilhelm (1956) derived an equation for which a dispersion number (d) included:
1 4a exp Ce 2d ¼ C0 ð1 þ aÞ2 exp a ð1 aÞ2 exp a 2d 2d
Influent SS ¼ negligible Effluent BOD5 ¼ 30 mg l1 First-order BOD5 removal-rate constant ¼ 0.30 d1 at 20 1C Temperature coefficient ¼ 1.05 at 20 1C Average pond temperature ¼ 25 1C Pond depth ¼ 2 m
Solution 1. Determine the BOD5 removal-rate constant.
k25 ¼ k20 ð1:05ÞT20 ¼ 0:38 d1
2. Determine the detention time. From Equation (7),
ð11Þ
30 1 ¼ 500 1 þ 0:38t
pffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi where a ¼ 1 þ 4ktd, with k being the first-order reaction rate, 1 d , and d the dispersion number, dimensionless. The value of d of a reactor can be determined by tracer study (Levenspiel, 1972; Metcalf and Eddy, 2003). Polprasert
t¼
16:67 1 0:38
BOD5 or fecal coliform remaining
60 50 40 d=1 30 d = 0.25 d=0
20 10 0 0
0.5
1
1.5
2
2.5 k.t
Figure 13 Design formula chart for partially mixed reactors.
ð12Þ
where u is the kinetic viscosity, W the width of the pond, L the length of fluid travel path from influent to effluent, and D the liquid depth of pond. Equation (12) was validated with WSP in USA and Thailand with satisfactory results. A design flowchart of Equation (11) is given in Figure 13 to simplify the calculation for BOD5 or fecal coliform reduction. In general, depending on HRT, d, and inlet/outlet arrangements, the values of d for most FP and MP are in the range of 0.1–0.5. The k values for BOD5 and fecal coliform removal of Equation (11) are different from those given in Equations (8) and (9) and need to be determined separately.
The values of kT for BOD5 reduction is given by
kT ¼ 0:3ð1:05ÞT20
0:184½tuðW þ 2DÞ 0:489 ðWÞ 1:511 ðLDÞ1:489
3
3.5
4
4.5
Constructed Wetlands and Waste Stabilization Ponds
t¼
15:67 0:38
289
Table 10 Summary of light saturation intensities (Is) for different freshwater algae Is
T ¼ 41:24 days Species
Temperature (1 C)
Chlorella pyrenoidosa
25 25 26 25 25
3. Determine the required pond area.
Pond volume ¼ 42 1000 ¼ 42 000 m 3
Chlorella vulgaris Scenedesmus obliquus Chlamydomonas reinhardti
Choose a pond depth of 2 m.
42 000 m 3 2m ¼ 21 000 m 2
Surface area ¼
Illuminance (ft-candle)
Irradiance (gcal cm2 d1)
250 500
51.8 36 82.1 18 36
500
36
500
25
Modified from Goldman (1979) Outdoor algal mass cultures. II. Photosynthetic yield limitation. Water Research 13: 119–136.
Therefore, the required pond is about 2.1 ha. Use three ponds in parallel; each pond’s surface area ¼ 7000 m2 (length width ¼ 140 50 m). Depending on the effluent standards or reuse, MP could be built in series to the FP to further polish the FP effluent.
50 Algal yield (P )
HRAP. There are several rationale methods available for HRAP design which can give different results depending on wastewater characteristics, climates, and operational procedures. Since the main objective of HRAP is to produce algal biomass, Goldman (1979) derived the following model in which the effect of algae decay was excluded:
60
40 30 20 10
Pa ¼ 0:28 Is ½In 0:45 I0 =Is þ 1
ð13Þ 2
1
where Pa is the algal productivity (dry weight), g m d , Is the saturation light intensity, gcal cm2 d1, and I0 the amount of visible solar energy penetrating a smooth water surface, varying from 0 to 800 gcal cm2 d1. The values of Is depend on temperature and algal species as shown in Table 10 and for most HRAP are in the range of 30– 80 gcal cm2 d1. The I0 value depends on latitude and weather conditions, which can range from 0 to 800 gcal cm2 d1. Figure 14 shows effects of I0 and Is on algal productivity. The average production of algae is reported as 70 or 35 ton ha1 yr1 algal protein; comparing with the productivity of conventional crops, wheat 3.0 (360 kg protein), rice 5.0 (600 kg protein) and potato 40 (800 kg protein) ton ha1 yr1 (Becker, 1981). With the design criteria as given in Table 5, the efficiency of BOD5, N and P removal could be expected to be more than 80%. As shown in Figures 10 and 11, the algal concentration in the HRAP liquid could be maintained up to 500 mg l1, which need to be harvested for further reuse and making the treated effluent suitable for discharging into a receiving water. Details of algal harvesting and reuses can be found in Polprasert (2007).
4.10.3.2.3 Operation, maintenance, and troubleshooting for WSP As with any wastewater treatment systems, good performance of WSP depends on regular maintenance of the system, which could be categorized into the following aspects: Control of short circuiting. Short circuiting of flow occurs to varying degrees in most WSP and could cause the actual HRT
0 0
100
200
300
400
500
600
700
Total sunlight irradiance (I0) Figure 14 Algal yield (Pa) as a function of total solar irradiance (Io) according to Equation (13).
to vary from 25% to 90% of the theoretical design HRT (Middlebrooks et al., 1982). The use of multiple ponds operated in series and multiple port inlet structures is effective in reducing short circuiting. In-basin baffles similar to those used in HRAP can also be effective. Control of seepage. To avoid water seepage and inflow of groundwater, lining of the pond bottom and inner dike surfaces is strongly recommended for permeable soils. Control of sludge accumulation. Sludge will accumulate to varying degrees on the pond bottom, but most of the accumulation will occur at or near the inlet structures. Occasional removal of the accumulated sludge should be done for maintaining the desired HRT and minimizing short circuiting. Control of scum and aquatic plants. Scum accumulation on the pond surface may occur in AP and FP, which could reduce the effective pond volume and HRT. Although scum accumulation is useful for maintenance of anaerobic conditions in AP, too much of them will interfere with hydraulic flow and should be regularly removed. There may be growth of aquatic plants such as duck weeds, for example, Lemna sp. and water hyacinth (Eichhornia crassipes) on the surface of FP and MP, which will prohibit the availability of sunlight for algal photosynthesis and reduce the pathogen die-offs due to less
290
Constructed Wetlands and Waste Stabilization Ponds
UV light penetration. Excessive growth of these aquatic plants will result in accumulation of decayed plant biomass, leading to the impairment of effluent quality. The growth of these aquatic plants should be prevented in these ponds or regular harvesting has to be done.
4.10.4 Case Studies of CW and WSP Natural systems such as CW and WSP have been applied for the treatment of wastewaters originating from municipal, industrial, and agricultural resources in several regions of the world. Some of the successful case studies and their performance results are presented in this section.
4.10.4.1 CW Case Studies
During the period of September 2001–May 2002, these two SF CW units were operated at the following conditions: hydraulic loading rates of 60–160 l m2 d1; organic loading rates of 57–140 kg BOD5 ha1 d1; and HRT of 1.4–4.0 days. The treatment performance of these CW units was found satisfactory (Table 11) with the effluent BOD5 and SS concentrations being 4 and 10 mg l1, respectively. Because of the nitrification reactions occurring at the CW beds, there was an increase in NO3–N concentrations from 0.06 mg l1 in the influent to 4.75 mg l1 in the effluent. This treated water is being sold to some factories located in the ESIE for uses in factory air-cooling and other processes. Due to prolific growth of the emerging vegetation under tropical conditions, plant harvesting was done once in 4 months, with the annual yields of 130–150 ton ha1 yr1 (wet weight). These harvested plants are being used to make furniture and other decorations, becoming another income-generation avenue for the ESIE.
4.10.4.1.1 CW treatment of an industrial wastewater As several industrial wastewaters contain both heavy metals and organic compounds, the required treatment processes should be those that can effectively remove these pollutants. A case study of CW application for the treatment of an industrial wastewater is given as follows: Eastern Seaboard Industrial Estate (ESIE), Rayong province, Thailand. The ESIE is located in Rayong province, 200 km east of Bangkok, Thailand (Polprasert, 2006). The industries at ESIE are required to pre-treat their wastewaters to remove heavy metals and other toxic compounds in accordance with the Thailand effluent standards, prior to discharging into combined sewers and mixing with other domestic wastewaters. The combined wastewater flow rate was about 7000 m3 d1. Part of this combined wastewater is being treated by two pilot-scale SF CW in series, each with a dimension of 35 18 0.8 m (length width depth). The CW beds are lined with high-density polyethylene sheet and filled with 1-cm gravel (Figure 15). The wastewater is applied intermittently over the CW beds in a downward vertical flow, and the percolates are collected through underdrainage pipes. Cattails, bulrushes, and canna are the primary vegetation grown in these CW beds (Figure 16).
4.10.4.1.2 CW treatment of a municipal wastewater Rehabilitation of wastewater treatment system on Phi Phi Island. After the catastrophic Tsunami disaster of December 2004,
Figure 16 CW at Eastern Seaboard Industrial Estate (ESIE), Rayong province, Thailand.
Unit dimensions for each unit Area = 630 m2 Width = 18 m Length = 35 m Figure 15 Schematic diagram of pilot-scale CW units (Koottatep et al., 2001).
Media arrangements Top soil = 10 cm Sand layer = 15 cm Gravel layer = 55 cm
Constructed Wetlands and Waste Stabilization Ponds
several infrastructures and dwellings on Phi Phi Island, a worldfamous tourist attraction of Krabi province in Southern Thailand, were drastically damaged (Figure 17) besides hundreds of deaths (Koottatep et al., 2007). The existing wastewater treatment systems were heavily destroyed, resulting in the discharge of untreated wastewater into the sea. Even before the year 2004, the 4-m depth WSP units at a design capacity of 400 m3 d1 Table 11
Treatment performance of the CW units
Parameter Average concentration (mg l1) a
BOD5 COD SS TKN NH3–N NO3–N TP a
Overall removal b
Influent Effluent unit 1
Effluent unit 2
(%)
90 230 98 24.1 14.2 0.06 7.0
4 19 10 4.6 3.3 4.75 1.5
95.5 91.7 89.5 81.1 76.7 -c 78.5
20 50 16 14.5 10.8 0.53 4.7
Percolate of CW unit 1. Percolate of CW unit 2 in series. c Increased, likely due to nitrification reaction. From Polprasert C (2006) Design and operation of constructed wetlands for wastewater treatment and reuses. In: Ujang Z and Henze M (eds.) Municipal Wastewater Management in Developing Countries: Principles and Engineering. London: IWA Publishing. b
291
were not properly functioning because of poor maintenance. Local people and tourists perceived these ponds as stinky and smelly units locating in the center of the island. With the generous support of the Danish government, the wastewater treatment systems were rehabilitated using the integrated CW system. The challenges in the rehabilitation of the wastewater treatment systems on Phi Phi Island were due to the public resistance to the previous wastewater treatment systems, which were not well maintained, whereas conventional systems such as activated sludge processes were not preferable because of the high costs as well as the sophistication in operation and maintenance. Based on several public consultative meetings, the CW system was considered to be a promising treatment alternative for this island. Furthermore, because of the freshwater scarcity, it was suggested that the CW system should be able to produce the effluent suitable for plant irrigation. The conceptual design of the CW systems included the integration of different flow patterns in series: vertical flow; horizontal SF and FWS in order to ensure the effluent quality. This integrated CW system was designed at the capacity of 400 m3 d1 for treating effluent of septic tanks and gray water from households, restaurants, hotels, and other dwellings. To prevent odor problem at the manholes and to avoid rainwater infiltration, the wastewater was collected by a separated sewer system and pumped to the CW units. Figure 18 shows ‘the integrated flower and the butterfly’ CW design. The wastewater is fed to the flower vertical flow
Figure 17 Damages of WSP systems on Phi Phi Island (Koottatep et al., 2005).
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Constructed Wetlands and Waste Stabilization Ponds
and horizontal SF CW units before feeding to the butterfly FWS CW units (Koottatep et al., 2007). The total surface area of the CW units is 8200 m2, each having 60 cm of gravel media. The treated effluent is collected at a 100-m3 underground tank where the local people can use it for irrigating their gardens; otherwise it is discharged to the sea. The CW units are planted with Canna, Heliconia, and Scirpus, most of which are colorful flowers, in vertical flow; horizontal SF and FWS CW units, respectively. In addition, some other plants and flowers are planted with the landscaping design as a community park (Figure 18). According to the design criteria of CW units, the integrated CW system can treat the wastewater with BOD5 and TKN concentrations of 100 and 20 mg l1, respectively. The performance of the integrated CW system has been found to be satisfactory and the effluent quality could meet the effluent standards for discharge into receiving waters (Table 12) or for agricultural reuse (Table 17).
4.10.4.1.3 CW treatment of fecal sludge or septage In tropical regions where most of the developing countries are located, septic tanks and other onsite sanitation systems are the predominant form of the storage and pre-treatment of excreta and wastewater, thus generating fecal sludge or septage (Koottatep et al., 2005). Uncontrolled septage management could endanger public health, hence the environmental protection for effective management strategies. Four demonstration CW units have been constructed at the Asian Institute of Technology, Pathumthani province, Thailand (Figure 19), to treat septage collected from Bangkok city. Each CW unit has a surface of 5 5 m2 with a substrata depth of 65 cm and a freeboard of 1 m. Cattail plants are grown in these CWs units and the solid loading rate of 250 kg TS m2 yr1 has been applied at the 6-day percolate impoundment. The COD, TS, and TKN removal efficiencies of these CW units were in the range of 80–96%. The biosolid accumulations in each of the CW units during the 7 years of operation were about 80 cm. There was no removal of these biosolids and the CW performance was not impaired probably due to extensive
growth of the cattail roots, which enhance the CW bed permeability and transfer O2 from the leaves to the root zone for bacterial degradation. The biosolids were found to contain viable helminth eggs below the standard for agricultural use (Table 17).
4.10.4.2 WSP Case Studies 4.10.4.2.1 WSP treatment of industrial wastewater and fish production A WSP unit, located in Yen So commune, Hanoi city, Vietnam, is being fed with water from the Kim Nguu River, which is heavily polluted with domestic and industrial wastewaters (Figure 20). The area of this WSP is 17 ha with depth of 1.2–1.5 m and is operated at an HRT of 7 days. Herbivorous fish are being cultured in this WSP. The characteristics of the Kim Nguu River water are: BOD5 ¼ 70–80 mg l1, TSS ¼ 80–120 mg l1, TN ¼ 50– COD ¼ 120–150 mg l1, 1 60 mg l , As ¼ 0.03–0.04 mg l1, Hg ¼ 0.0005–0.0016 mg l1, and fecal coliforms ¼ (2 –9) 108 MPN (100 ml)1. According to Toan (2008), the removal efficiencies of BOD5, TSS, and TN by this WSP were found to be 60%, 40%, and 75%, respectively. However, there were only about 98% removal of fecal coliform and the WSP water still contained fecal coliform concentrations above the safe limit for fish production (Table 17). The heavy metal (including As) concentrations were found to be within the EU threshold values or the Codex
Table 12
Effluent Standards from municipal sources
Parameters
Concentrations
pH BOD5 SS TKN Fat and oil
5–9 20–50 mg l1 30–50 mg l1 35–40 mg l1 20 mg l1
From MOI (2005). Wastewater Treatment Standards. Bangkok, Thailand: Ministry of Industry.
Figure 18 Bird-eye view of The Flower and The Butterfly CW system (Koottatep et al., 2007).
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4.10.4.2.2 WSP treatment of wastewater containing endocrine disrupting chemicals Endocrine disrupting chemicals (EDCs) are synthetic organic chemicals introduced to the environment by anthropogenic inputs. Although they may be present in the water environment at low concentrations, their ability to bioaccumulate and biomagnify in the food chains can pose health risks to humans and animals. This study reported the performance of three wastewater treatment plants in the removal of estrogenicity from wastewaters. The WSP units with HRT of 7–10 days demonstrated 90–95% removal of estrogenicity, higher than those of the tricking filters, which could remove about 40% estrogenicity. The mechanisms responsible for EDC removal in these WSP units were hypothesized to be biodegradation, adsorption on the settleable solids, and sedimentation. The MP with HRT of 26–47 days could enhance further estrogenicity removal, hence demonstrating the effectiveness of the WSP system in EDC removal (Gomez et al., 2007).
4.10.5 Emerging Environmental Issues versus Potentials of CW and WSP
Figure 19 Demonstration CW units have been installed at the Asian Institute of Technology, Pathumthani province, Thailand.
The world’s population has grown significantly during the past 100 years and is expected to increase further from the present number of about 6 billions in 2008 to about 9 billions (a 50% increase) in 2050 (Figure 1). The high population growth, together with industrialization, will obviously result in more demands for energy as shown in Figure 21. In addition, there will be increases in quantity of municipal and industrial wastewaters that need to be managed in a satisfactory manner. As indicated in the previous sections, some natural systems such as CW and WSP, if properly designed and operated, can treat these wastewaters to the degree comparable to secondary or tertiary treatment. Other advantages of CW and WSP include their potentials for waste recycling through the production of food for humans or animals, energy and water conversation, and climate change mitigation. Examples of these potentials are given in the following.
4.10.5.1 Food Production The operation of CW and WSP systems could lead to the production of biomass, which could be converted to or utilized as foods for humans or animals or other useful products. The principles of waste recycling and biomass production in these systems are described below.
4.10.5.1.1 Biomass production in CW Figure 20 Kim Nguu river (Toan, 2008).
threshold value. Fish production in this pond of 4–5 ton ha1 yr1 has been reported and the harvested fish are being sold for public consumption. With the above information, these fish should be well cooked before eating and preventive measures such as depuration should be implemented to minimize health risks from pathogen infection.
The growth of emerging aquatic plants in a CW unit can be described by daily increment factor as proposed by Mitchelle (1974):
Nt ¼ N0 xti
ð14Þ
where Nt, N0 are the final and initial numbers of plants, ti is the time interval for plant growth, day and x is the daily incremental factor, dimensionless.
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Constructed Wetlands and Waste Stabilization Ponds 800 History
Projections 608 563
600 Quadrillion Btu
695 652
512 462 398
400 284
309
347
365
200
20 30
20 25
20 20
20 15
20 10
20 05
20 00
19 95
19 90
19 85
19 80
0
Figure 21 World marketed energy consumption, 1980–2030 (EIA, 2008). Note: 1 quadrillion Btu ¼ 2.928 1011 kilowatt-hour (kWh).
The area doubling time (td) is the time when Nt/N0 ¼ 2 and, from Equation (14), can be determined from
td ¼
ln 2 ðln Nt ln N0 Þ=ti
Table 13 Proximate compositions (% dry weight) of some aquatic plants and alfalfa hey Species
Ash
Crude protein
Fat
Cellulose
Typha latifolia Justicia americana Sagittaria latifolia Alternanthera philoxeroides Orontium aquaticum Alfaalfa hay
6.9 17.4 10.3 13.9 14.1 8.6
10.3 22.9 17.1 15.6 19.8 18.6
3.9 3.4 6.7 2.7 7.8 2.6
33.2 25.9 27.6 21.3 23.9 23.7
ð15Þ
The value of td varies with plant species, climates, water quality, and other environmental conditions. Under tropical conditions, the td of several emerging plants can range from 4 to 8 weeks; the td values are longer for CW units located in temperature climates. The knowledge of td is useful to determine the frequency of plant harvesting without interfering with treatment performance or the biomass productivity. If necessary, the td value can be determined from experiments with pilot-scale CW units in order to obtain accurate biomass productivity. Westlake (1963) reported the productivity ranges of emerging aquatic plants as 27–77 ton of dry organic matter ha1 yr 1. The proximate composition of some emerging aquatic plants, shown in Table 13, indicates their crude protein contents to be 10–20%, similar to alfalfa hay. A promising technique to use the harvested aquatic plants from CW as animal food is to convert them to silage (NAS, 1976). This is accomplished by chopping the aquatic plants into small pieces (1–2 cm) and packing them in a silo under anaerobic conditions. The organic acids such as acetic acids and lactic acids produced in the silo will keep the pH at about 4 and avoiding putrefaction of the biomass. After about 20 days of silaging, the silaged products could be used as supplemental feeds to animals such as cattle. The harvested aquatic plants can be used as a raw material to make compost fertilizer. However, since the cell walls contain high cellulose content (Table 13), which is not easily biodegradable, the aquatic plants should be chopped into small pieces first (1–2 cm preferable) before composting. Because aquatic plants have a high C content, they should be mixed with other wastes that have high N contents (such as sludge or manure) so that the mixture C/N ratio becomes 30/1, optimum for composting process. Further details
From Boyd CE 1974. Utilization of aquatic plants. In: Mitchell DS (ed.) Aquatic Vegetation and Its Use and Control, pp. 107–115. Paris: UNESCO.
of aquatic plants composting can be found in Polprasert (2007).
4.10.5.1.2 Biomass production in WSP Due to the long HRT and without media, there appears to be more potential for biomass production in WSP in the form of algal and fish protein biomass, as described in the following: Algal protein production. Algal cells have high protein content (about 50%) and subsequent harvesting of algae growing in HRAP for human or animal consumption will be a financial incentive for wastewater treatment. The data of average algal productivity presented in Table 14 suggested that protein production through algal cultivation is far more effective than cultivating other conventional crops. The chemical compositions of different algae are presented in Table 15, which show them to have higher protein contents than soya bean. Due to the rapid generation of algae, HRAP treating wastewaters and operating outdoors under ambient conditions will have growth of mixed algal species such as those shown in Table 15. The harvested algal cells from HRAP can be fed to animals (such as cattle, pigs, poultry, and fish) and these animals are used as food for humans. This strategy would lengthen the food chains, minimizing health risks and
Constructed Wetlands and Waste Stabilization Ponds Table 14
Comparative protein productivity
Crop
Biomass productivity (tonnes ha1 yr1)
Protein content (%)
Protein productivity (tonnes ha1 yr1)
Algae Wheat Rice Potato
70 3 5 40
50 12 12 20
35 0.36 0.60 0.80
From Becker EW (1981). Algae mass cultivation – production and utilization. Process Biochemistry 16: 10–14.
Table 15 Chemical composition of different algae compared with soya (% dry matter) Component
Scenedesmus
Spirulina
Chlorella
Soya
Crude protein Lipids Carbohydrates Crude fiber
50–55 8–12 10–15 5–12
55–65 2–6 10–15 1–4
40–55 10–15 10–15 5–10
35–40 15–20 20–35 3–5
From Becker EW (1981). Algae mass cultivation – production and utilization. Process Biochemistry 16: 10–14.
promoting better social acceptance. Except for Spirulina sp. which has soft cell walls, most of the waste-grown algae have thick cell walls which are not easily digestible by nonruminant animals (such as poultry). Therefore, these cell walls need to be ruptured by heat or acid treatment before feeding to the nonruminants. Hintz et al. (1966) reported that the waste-grown algae (Chlorella and Scenedesmus) were 73% digestible when fed to ruminant animals such as cattle and sheep, and were only 54% digestible when fed to pigs. The digestible energy content for the ruminants was 2.6 kcal g1 algae. These algae were found to supply adequate protein to supplement barley for pigs. Alfalfa–algae pellets, when fed to lambs, resulted in higher weight gains than alfalfa pellets alone. Algae are basically not palatable to most livestock, but this may be overcome by pelletizing the processed algae with usual feed of the particular animal, such as steam barley in the case of cattle. In general, the waste-grown algae appear to have potential as a livestock feed because of the high contents of protein and other valuable substances (Table 15). Algal cells may also serve as a source of steroids. The concentration of steroids in algae is variable but significant amount may be found in some algal species. Algae may also contain up to 0.2% dry weight as carotenoids (Paoletti et al., 1976). Some medicinal products have been isolated from algae (Volesky et al., 1970). Despite the high potentials of applying HRAP for wastewater treatment and protein biomass production, the key challenge to this technology is the selection of an algal harvesting technique that is effective and economical. The algal biomass growing in HRAP is usually in the microscopic unicellular forms with sizes ranging from 10 to 100 mm. These algal cells need to be removed or harvested from the HRAP water so that the treated effluent will have low SS that meets the discharge standards; the harvested algal cells could be
295
processed for further reuse. Polprasert (2007) listed several algal harvesting technologies such as micro-straining, paperprecoated belt filtration, flocculation, flotation and centrifugation, and so on. Due to the wide range of conditions and objectives encountered, the selection of an algal-harvesting technology and its application should be approached on a case-by-case basis, considering the advantages and disadvantages of each technology and the intended uses of the algal biomass. Fish protein production. The use of waste-grown algae as food for herbivorous fish (feeding on algae and plants) or omnivorous fish (feeding on plants or animals) will minimize the problems of algal harvesting and algal digestibility mentioned above. Common herbivores and omnivores that could be fed with waste-grown algae are: Tilapia (Oreochromis niloticus), Chinese common carp (Cyprinus carpio), grass carp (Ctenopharyngoden idelia), and so on. These fish could be raised in ponds fed with harvested algal cells from HRAP or they could be raised directly in FP or MP to graze on the algae present in the pond water. The reuse of municipal and agricultural wastes through the production of algal and fish protein has been practiced in several countries (Table 16). Fish yields from this practice vary widely ranging from 1 to 10 ton ha1 yr1 depending on wastewater characteristics, climates, water quality, and fish pond management. To minimize health risks, the World Health Organization proposed that wastewaters to be fed to fish ponds should be pre-treated to have the microbiological quality as shown in Table 17 (Mara and Cairncross, 1988). Another measure believed to reduce the transfer of pathogens is to raise carnivorous fish for human consumption using the herbivorous fish from waste-fed ponds as feed. The waste-grown fish can be used as feed for carnivorous fish or shrimps, which have high market value when sold for human consumption. The biological value of protein in fish meal is 75–90%, which is quite high (Williamson and Payne, 1978) and is suitable for feeding to pigs and chicken. For this purpose, fish in waste-fed ponds can be reared at a higher stocking density and shorter growing period to obtain high fish yield. Fish can be sundried, grounded, and mixed with other food stuffs to increase the value of fish meal. Experiments were conducted at Asian Institute of Technology (AIT) to investigate the feasibility of using Tilapia grown in septage-fed ponds as fish feed ingredients for carnivorous fish of high market value such as snakehead (Channa striata) and walking catfish (Clarias batrachus and C. macrocephalus) (Edwards et al., 1988). Tilapia harvested from the septage-fed ponds was used either fresh, or processed as silage and mixed with other feed ingredients prior to feeding to the carnivores. A silage was produced by adding 20% carbohydrate (cassava) to minced Tilapia in the presence of Lactobacillus casei. It was found that growth of these carnivorous fish in ponds fed with Tilapia fish meal, fresh Tilapia, or silaged Tilapia was comparable with that fed with marine trash fish (conventional fish feed). In many cases, depuration, which is a natural process to remove some microorganisms or toxic and malodorous compounds from the fish body, can be applied to improve quality of the waste-grown fish. By putting contaminated fish
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Constructed Wetlands and Waste Stabilization Ponds
Table 16
Fish yields in waste-fed ponds
Country
Type of waste
Yield (kg ha1 yr1)
Remark
Israel Philippines Indonesia Poland India China Munich, Germany
Cattle manure Biogas slurry Livestock waste Sugar beet wastes Wastewater Wastewater Wastewater
10 950 8000 7500 400–500 kg ha1 – growing season 958–1373 6000–10 000 500 kg ha1
Extrapolated from experiment
Hungary
Wastewater
1700 kg ha1
USA
Wastewater effluent, 37% wastewater Septage, loading ¼ 150 COD ha1 d1
126–218 5000 5000–6000
AIT, Thailand
Indian carp Common carp, per growing season Polyculture, Chinese carp and common carp, per growing season Channel catfish Silver carp, bighead carp Tilapia (extrapolation)
Modified from Polprasert C (2007). Organic Waste Recycling, Technology and Management, 3rd edn. London: IWA Publishing.
Table 17 Tentative microbiological quality criteria for the aquacultural use of wastewater and excreta Reuse process
Fish culture Aquatic macrophyte culture
Viable trematode eggsa (arithemetic mean number per liter or kg)
Fecal coliforms (geometric mean number per 100 ml or per 100 g)b
0 0
o104 o104
a
Clonorchis, Fasciolopsis, and Schistosoma. Consideration should be given to this guideline only in endemic area. b This guideline assumes that there is a one log 10 unit reduction in fecal coliforms occurring in the pond, so that in-pond concentrations are o1000 per 100 ml. If consideration of pond temperature and retention time indicates that a higher reduction can be achieved, the guideline may be relaxed accordingly. From Mara DD and Cairncross AM (1988). Guidelines for the Safe Use of Wastewater and Excreta in Agriculture and Aquaculture: Measures for Public Health Protection. Geneva: World Health Organization.
in clean water for 1–2 weeks, most of the contaminants retained on or in the fish body could be cleansed out. The nutritional value of waste-grown fish has shown to be better than those grown using other sources of feed. Most of the nutrition applied in the form of waste is converted to protein rather than fat as reported by Moav et al. (1977) and Wohlfarth and Schroeder (1979) . The use of waste-grown algae as feed for herbivorous fish (Tilapia) was reported by Edwards et al. (1987). Extrapolated fish yields approaching 20 ton ha1 yr1 were obtained in the 4 m3 concrete pond system based on 3-month growing periods under ambient, tropical conditions. A linear relationship was established between fish yields and means algal concentration in the fish ponds, in which an algal concentration of 70 mg l1 in the pond water was considered to be high enough to produce good fish growth. Higher algal concentrations were not recommended since it might lead to zero DO concentrations in the early morning hours. Another study by Edwards et al. (1988) applied septic tank sludge to 200 m2 earthen ponds to grow algae, which were subsequently grazed by
Tilapia. At the organic loading rate of 250 kg COD ha1 d1 and stocking density of 20 fish m2, the fish productivity of about 11 ton ha1 yr1 was obtained.
4.10.5.2 Energy and Water Conversation and Climate Change Mitigation The increasing uses of fossil fuels have resulted in global warming and climate change such as rising seawater levels, severe drought, and flooding. Natural systems such as CW and WSP are cheaper to build and operate and consume less energy in operation than conventional treatment systems (Table 1). Accordingly, these natural systems use less fossil fuels, generate less green house gases (GHGs), and contribute to climate change mitigation. A case study on GHG reduction and water conversation was conducted by Kittipongvises (2008) at the Sanguan Wongse Industries (SWI) factory in Nakorn Ratchasima, northeastern Thailand, which employed a covered AP to treat 7000 m3 d1 of its starch-processing wastewater. There was more than 90% reduction of COD and the produced CH4 gas of 50 000–80 000 m3 d1 was being used as heat in tapioca processing and in electricity generation (Figure 22). This CH4 utilization resulted in mitigation of GHG emission of 300 000 ton CO2eq-yr1 which, based on the Clean Development Mechanisms (CDM) and Certified Emission Reductions (CER) rate of Euro 10 per ton CO2eq, would yield an additional income of Euro 3 000 000 (or 120 million baht). Furthermore, financial gains from replacement of heavy fuel oil and grid-fed electricity were about US$1000 000 and US$250 000 per annum, respectively. The treated wastewater from the covered lagoon was used to irrigate nearly tapioca crops, hence contributing to water conservation. A recent report by Haag (2007) indicated that the potential of biodiesel production of about 90 000 l yr1 from algae cells grows in a 1-ha HRAP; soya could produce biodiesel to only 450 l ha1 yr1, canola about 1200 l ha1 yr1and oil palm about 6000 l ha1 yr1. Algae cells could yield about 50% of their weight in oil, higher than oil palms, which typically yield about 20%. Oils squeezed from algal cells are biocrude, the
Constructed Wetlands and Waste Stabilization Ponds
297
Figure 22 Anaerobic ponds with captured biogas for electricity generation (Kittipongvises, 2008).
Figure 24 A biosphere on Mars utilizing CW and HRAP. Figure 23 International Space Station (www.science.national geographic.com/science, cited on 19 January 2009).
precursor of biodiesel, which needs to be processed further to convert them to become biodiesel. The rest of the algal biomass can be converted into ethanol and animal feed. Besides the economic advantages, converting waste-grown algae to
biodiesel would contribute to global warming mitigation (algae need CO2 for photosynthesis, Equation (1)) and the produced biodiesel should not pose health and/or social problems.
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Constructed Wetlands and Waste Stabilization Ponds
4.10.5.3 CW and HRAP as Future Life-Support Technology The current progress in space exploration should eventually lead to a possibility that human beings can stay in space, for example, at the International Space Station (ISS, Figure 23) for long duration. To minimize launch costs and re-supply requirements, the treatment and recycling of water by growing algal cells and aquatic plants for use as human foods at the ISS may be necessary, and the CW and HRAP technologies as described in this chapter appear relevant. Since there is more than 95% content of CO2 in the air of the Mars’ atmosphere, the potential of applying CW and HRAP to produce algal cells and food production in a biosphere on Mar could be considered, as shown in Figure 24.
4.10.6 Summary Increased global population and industrialization have become a major challenge for government and private sectors to adequately manage the generated wastewaters. The current problems of energy and food crisis and global warming have made the effort to treat the wastewaters more difficult. Natural systems such as CW and WSP are alternative technologies which are low cost, but can provide high treatment efficiencies comparable to or even better than those of conventional wastewater treatment systems. The principal mechanisms responsible for wastewater degradation are the interactions between emergent aquatic plants (in case of CW) or algae (in case of WSP) and bacteria growing in these systems. Through photosynthetic reactions, O2 produced by the emerging aquatic plants or algal cells are used by the bacteria in biodegrading the incoming organic matters and the biodegradation by-products such as CO2 and NH3 are used by the plants and algae for their photosynthesis and growth. The relatively long HRT occurring in these natural systems results in effective sedimentation, adsorption, and plant uptake of SS and nutrients. Pathogen die-offs in WSP are due to UV light exposure and high pH conditions prevailing during photosynthesis, resulting in acceptable concentration of fecal microorganisms in MP effluent for discharge into a receiving water body or reuses in irrigation and aquaculture. Pathogen die-offs in CW are less effective because the emerging aquatic plants prevent sunlight from reaching the wastewaters. Since there are physical, chemical, and biological reactions occurring in these natural systems and their performance is dependent to a large extent on climates, the available design criteria are based on empirical or rationale approaches. The empirical method for CW and WSP design is based mainly on relationships between organic loading rates, HRT, and their treatment performance. The rationale approach for design of a CW unit is the plug-flow model and assuming first-order kinetics in which the reaction rate is temperature dependent. Since there are emerging aquatic plants and media in the CW bed, the effect of bed porosity on the actual HRT and, consequently, the CW volume and surface area has to be taken into consideration. In WSP design, the completely mixed model assuming first-order kinetic is commonly used for FP and MP design for both organic matter and fecal coliform reductions. The
first-order reaction rates are temperature dependent, similar to that of CW. The operation of CW and WSP results in the production of biomass, which could be converted to or utilized as food for human and animals or other useful products. The harvested aquatic plant can be used as a raw material in making compost fertilizer or silage for use as animal feeds. The nutritional contents of these aquatic plants are similar to those of alfalfa hay. The algal productivity from HRAP is estimated to be 70 ton (dry weight) ha1 yr1, much higher than other terrestrial crops. Since algal biomass contains about 50% protein, they can be used as food for animal and herbivorous fish, which could consequently serve as human foods. Some medicinal and chemical products have been isolated from algal cells for pharmaceutical and industrial purposes. Due to the presence of algae in FP and MP water, herbivorous fish can be raised in these ponds and the harvested fish, if properly processed, could be used as animal or human foods. The AP system offers high opportunity for energy recovery through the production of CH4 gas. A case study on AP treatment of a tapioca-processing wastewater showed that the captured CH4 gas could be converted to heat for tapioca processing and electricity generation, hence, a financial return to the industry and contributing to the reduction of global warming problems. With the current progress in the space exploration, human beings could stay in space or eventually on the Moon or Mars for a long duration. The natural systems such as CW and WSP as discussed in this chapter could be considered as possible technologies for life-support systems in the production of O2 and biomass.
References Becker EW (1981) Algae mass cultivation – production and utilization. Process Biochemistry 16: 10--14. Boyd CE (1974) Utilization of aquatic plants. In: Mitchell DS (ed.) Aquatic Vegetation and Its Use and Control, pp. 107--115. Paris: UNESCO. Edwards P, Polprasert C, Rajput VS, and Pracharaprakiti C (1988) Integrated biogas technology in the Tropics. 2. Use of slurry for fish culture. Waste Management and Research 6: 51--61. EIA (2008) International Energy Outlook. Washington, DC: Energy Information Administration, US Department of Energy. http://www.eia.doe.gov/oiaf/ieo/ index.html Girts MA and Knight RL (1989) Operations optimization. In: Hammer DA (ed.) Constructed Wetlands for Wastewater Treatment: Municipal, Industrial, and Agricultural. Chelsea, MI: Lewis. Goldman JC (1979) Outdoor algal mass cultures. II. Photosynthetic yield limitation. Water Research 13: 119--136. Gomez E, Wang X, Dagnino S, et al. (2007) Fate of endrocrine disrupters in waste stabilization pond systems. Water Science Technology 55: 157--163. Haag AL (2007) Algae bloom again. Nature 447: 520--521. Hintz HF, Heitman H Jr., Weir WC, Torell DT, and Meyer JH (1966) Nutritive value of algae grown on sewage. Journal of Animal Science 25: 675--681. IPCC (2007) An Assessment of the Intergovernmental Panel on Climate Change. Geneva, Switzerland: Intergovernmental Panel on Climate Change. Kittipongvises S (2008) Potential of Clean Development Mechanism (CDM) Activities for Greenhouse Gases Reduction at a Starch-Processing Factory in Thailand. Master’s Thesis, Asian Institute of Technology, Thailand. Koottatep T, Polprasert C, and Laugesen C (2007) Integrated eco-engineering design for sustainable management of fecal sludge and domestic wastewater. Journal of Korean Wetland Society 9: 69--78. Koottatep T, Polprasert C, Oanh NTK, et al. (2001) Septage dewatering in vertical-flow constructed wetlands located in the tropics. Water Science and Technology 44: 181--188.
Constructed Wetlands and Waste Stabilization Ponds Koottatep T, Surinkul N, Polprasert C, Kamel AS, and Strauss M (2005) Treatment of septage in constructed wetlands in tropical climate-lessons learnt after seven years of operation. Water Science and Technology 51: 119--126. Levenspiel O (1972) Chemical Reaction Engineering, 2nd edn. New York, NY: Wiley. Mara DD, Alabaster GP, Pearson HW, and Mills S (1992) Waste Stabilisation Ponds: A Design Manual for Eastern Africa. Leeds, UK: Lagoon Technology International. Mara DD and Cairncross AM (1988) Guidelines for the Safe Use of Wastewater and Excreta in Agriculture and Aquaculture: Measures for Public Health Protection. Geneva: World Health Organization. Marais G (1974) Fecal bacteria kinetics in waste stabilization ponds. Journal of the Environmental Engineering Division, ASCE 100: 120--139. Metcalf and Eddy Inc. (2003) Wastewater Engineering: Treatment and Reuse, 4th edn. New York, NY: McGraw-Hill. Middlebrooks EJ, Middlebrooks CH, Reynolds JH, et al. (1982) Water Stabilization Lagoon Design, Performance, and Upgrading. New York, NY: Macmillan. Mitchelle DS (1974) Aquatic Vegetation and Its Use and Control. Paris: UNESCO. Moav R, Wohlfarth G, and Schroeder GL (1977) Intensie polyculture of fish in freshwater ponds. I. Substitution of expensive feeds by liquid cow manure. Aquaculture 10: 25--43. MOI (2005) Wastewater Treatment Standards. Bangkok, Thailand: Ministry of Industry. NAS (1976) Making Aquatic Weeds Useful: Some Perspectives for Developing Countries. Washington, DC: US National Academy of Sciences. Pano AE and Middlebrooks J (1982) Ammonia nitrogen removal in facultative wastewater stabilization ponds. Journal of the Water Pollution Control Federation 54: 344--351. Park WH and Polprasert C (2008) Roles of oyster shells in an integrated constructed wetland system design for P removal. Ecological Engineering 34: 50--56. Paoletti C, Phushparaj B, Florenzano G, Capella P, Lercker G (1976) Unsaponifiable matter of green and blue-green algal lipids as a factor of biochemical differentiation of their biomass: I. Total unsaponifiable and hydrocarbon fraction: Lipids 11: 258-–265. Polprasert C (2006) Design and operation of constructed wetlands for wastewater treatment and reuses. In: Ujang Z and Henze M (eds.) Municipal Wastewater Management in Developing Countries: Principles and Engineering. London: IWA Publishing. Polprasert C (2007) Organic Waste Recycling, Technology and Management, 3rd edn. London: IWA Publishing. Polprasert C and Agarwalla BK (1994) A facultative pond model incorporating biofilm activity. Water Environment Research 66: 725--732. Polprasert C and Bhattarai KK (1985) Dispersion model for waste stabilization ponds. Journal of Environmental Engineering, ASCE 111: 45--49. Polprasert C, Khatiwada NR, and Bhurtel J (1998) Design model for COD removal in constructed wetlands based on biofilm activity. Journal of Environmental Engineering, ASCE 124: 838--843. Reed SC, Middlebrooks EJ, and Crites RW (1988) Natural Systems for Waste Management and Treatment. New York, NY: McGraw-Hill.
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Sawaittayothin V and Polprasert C (2007) Nitrogen mass balance and microbial analysis of constructed wetlands treating municipal landfill leachate. Bioresource Technology 98: 565--570. Silva SA, De Oliveira R, Soares J, Mara DD, and Pearson HW (1995) Nitrogen removal in pond systems with different configurations and geometries. Water Science and Technology 31: 321--330. Stern N (2006) The Economics of Climate Change. London: HM Treasury. Stowell R, Ludwig R, Colt J, and Tchobanoglous G (1980) Towards the Rational Design of Aquatic Treatment Systems. Paper presented at the ASCE Convention, Portland, OR, USA, 14–18 April. Department of Civil Engineering, University of California, Davis, CA, USA. Toan TQ (2008) Potential of Domestic Wastewater Reuse in Vietnam: A Case Study in Wastewater Fed Fish Ponds in Yenso Commune, Hanoi City. Master’s Thesis, Asian Institute of Technology, Thailand. UN (2005) Monitoring Progress Towards the Achievements of Millennium Development Goals. http://unstats.un.org/unsd (accessed February 2010). UN (2008) The Millennium Development Goals 2008 Report. New York, NY: United Nations. US Census Bureau (2007) International Data Base. July 2007 version. Washington, DC: US Census Bureau. US EPA (1988) Design Manual – Constructed Wetlands and Aquatic Plant Systems for Municipal Wastewater Treatment. EPA/625/1-88/022. Cincinnati, OH: United States Environmental Protection Agency. US EPA (2000) Free Water Surface Wetlands for Wastewater Treatment: A Technology Assessment. Washington, DC: Office for Water Management, US Environmental Protection Agency. Visesmanee V, Polprasert C, and Parkpian P (2008) Long-term performance of subsurface-flow constructed wetlands treating Cd wastewater. Journal of Environmental Science and Health, Part A – Toxic/Hazardous Substances and Environmental Engineering A43: 765--771. Volesky B, Zajic JE, and Knettig E (1970) Algal products. In: Zajic JE (ed.) Properties and Products of Algae, pp. 49--82. New York, NY: Plenum. W.E.F (2001) Natural Systems for Wastewater Treatment, Manual of Practice FD–16, 2nd edn. Alexandria, VA: Water Environment Federation. Wehner JF and Wilhelm RH (1956) Boundary conditions of flow reactor. Chemical Engineering Science 6: 89--93. Westlake DF (1963) Comparison of plant productivity. Biological Review 38: 385--425. WHO (2000) Global Water Supply and Sanitation Assessment 2000 Report. Geneva: World Health Organization. Wieder RK, Tchobanoglous G, and Tuttle RW (1989) Preliminary considerations regarding constructed wetlands for wastewater treatment. In: Hammer DA (ed.) Constructed Wetlands for Wastewater Treatment: Municipal, Industrial, and Agricultural. Chelsea, MI: Lewis. Williamson G and Payne WJA (1978) Animal Husbandry in the Tropics, 3rd edn. London: Longman. Wohlfarth GW and Schroeder GL (1979) Use of manure in fish farming – a review. Agricultural Wastes 1: 279--299.
4.11 Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis AG Fane, CY Tang, and R Wang, Nanyang Technological University, Singapore & 2011 Elsevier B.V. All rights reserved.
4.11.1 4.11.1.1 4.11.1.2 4.11.1.2.1 4.11.1.2.2 4.11.1.2.3 4.11.1.2.4 4.11.1.3 4.11.1.3.1 4.11.1.3.2 4.11.1.3.3 4.11.1.3.4 4.11.2 4.11.2.1 4.11.2.2 4.11.2.3 4.11.3 4.11.3.1 4.11.3.2 4.11.3.2.1 4.11.3.2.2 4.11.3.3 4.11.3.4 4.11.4 4.11.5 4.11.5.1 4.11.5.2 4.11.5.2.1 4.11.5.2.2 4.11.5.2.3 4.11.5.2.4 4.11.6 4.11.6.1 4.11.6.2 4.11.6.2.1 4.11.6.2.2 4.11.6.3 4.11.6.4 4.11.6.5 4.11.7 4.11.7.1 4.11.7.2 4.11.7.3 4.11.8 References
Introduction The Range of Membrane Processes Role in Water Supply, Sanitation, and Reclamation Water treatment Desalination Water reclamation Wastewater MBR Status of Development Seawater desalination Water reclamation Water treatment Membrane bioreactors Membrane Types and Properties Membrane Types and Important Membrane Properties Membrane Properties for RO and NF Membranes Membrane Properties for MF and UF Membranes Membrane Materials and Preparation Polymeric Membrane Materials Hollow Fiber Preparation Mechanism of membrane formation Fabrication of hollow fiber membranes TFC Membrane Preparation Ceramic Membrane Preparation Membrane Characterization Membrane Modules The Role of the Module Module Types Spiral-wound module Tubular module Hollow fiber module (contained) Submerged module Basic Relationships and Performance Membrane Flux and Rejection Transport Inside a Membrane – Basic Relationships Transport models for MF and UF membranes Transport models for RO membranes Transport toward a Membrane – Concentration Polarization Factors Affecting Membrane Performance Membrane Fouling Membrane Process Operation Crossflow versus Dead-End Operation System Components Energy and Economic Issues Conclusions
Nomenclature a A Am
B molecular radius water permeability coefficient (solutiondiffusion model) membrane area
C C
301 302 302 302 302 304 304 305 305 305 305 306 306 306 308 309 312 312 312 312 317 319 319 319 321 321 321 322 322 322 323 324 324 326 327 327 328 329 330 332 332 332 333 333 333
solute permeability coefficient (solutiondiffusion model) solute concentration average solute concentration inside the membrane
301
302
Cb Cf Ci Cm Cp Cwm dh D Dsm Dwm Jcrit Js Jw K KKC Ksm lm Lp Ls m˙ s Mw P Pm Pp
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
bulk concentration feedwater concentration molar concentration of dissolved species i solute concentration near the membrane surface solute concentration in the permeate water concentration of water inside the RO rejection layer hydraulic diameter diffusion coefficient of solute diffusion coefficient of solute inside the RO rejection layer diffusion coefficient of water inside the RO rejection layer critical flux solute flux water flux mass transfer coefficient Kozeny–Carman coefficient solute partitioning coefficient into the RO rejection layer thickness of the rejection layer water permeability solute permeability coefficient solute mass flow rate molecular weight hydraulic pressure hydraulic pressure near the membrane surface hydraulic pressure of the permeate water
4.11.1 Introduction Membrane technology is used in the water industry to improve the quality of water for use, reuse, or discharge to the environment. Membranes range from finely porous structures to nonporous and can remove contaminants such as bacteria and protozoa down to ions. The advantages of membrane technology include its modular nature, allowing application at very large or small scale, the quality of the product water, the relatively small footprint, and, in some cases, the lower energy usage. Increased water scarcity, coupled with steady improvements in membrane performance, costs, and energy demand, will see a steady growth in membranes in the water industry into the foreseeable future.
4.11.1.1 The Range of Membrane Processes Membranes used for purification and separation can be defined as semipermeable thin films. The semipermeable property means that membranes may be able to transport water but not bacteria (microfilters) or salts (reverse osmosis, RO). Other membranes are able to transport salts but not water (electrodialysis). The family of membrane processes is depicted in Figure 1 which shows the driving force for transport
Qp rp R Rapp Re Rf Rg Rint Rm Rsys S Sc Sh T u v Vw Y e k g d p pm pp r s
volumetric flow rate that permeates through the membrane pore radius rejection apparent rejection of a membrane Reynolds number foulant hydraulic resistance universal gas constant (R ¼ 8.31 J mol1 K1) intrinsic rejection of a membrane membrane hydraulic resistance overall rejection – or rejection at a system level specific surface area Schmidt number Sherwood number absolute temperature (K) flow velocity kinetic viscosity molar volume of water recovery membrane porosity ratio of solute (or particle) diameter to the pore diameter viscosity of water boundary layer thickness osmotic pressure osmotic pressure at the membrane surface osmotic pressure of the permeate water reflection coefficient tortuosity of the membrane
and the size range of the species involved. For driving forces DC and DE, the species transported are solutes and water transport is low. For the liquid-phase pressure-driven processes (DP), water is transported and other species are partially or wholly retained. Gas-phase separations are also possible, such as N2/O2 or CO2/CH4, and details can be found elsewhere (Hagg, 2008). Membrane distillation (MD) is a process driven by temperature difference (DT) using hydrophobic membranes with vapor-filled pores. The DT provides a vapor pressure driving force for water vapor transport; a detailed review can be found elsewhere (Khayet, 2008). The membrane processes of interest in the water industry are the pressure-driven liquid-phase processes summarized in Table 1. The microfiltration/ultrafiltration (MF/UF) range can be regarded as a continuum. These membranes are typically produced by phase inversion, and small changes in preparation technique adjust the nominal pore size. The nanofiltration/reverse osmosis (NF/RO) membranes also represent a continuum and are usually produced as thin-film composite (TFC) structures. Section 4.11.2 gives details of the various types of membrane and their properties. Section 4.11.3 describes membrane preparation and Section 4.11.4 shows how membrane properties are characterized. Section 4.11.5 explains the role of the membrane module (housing) and
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303
Applications Organics Driving force
Ionic Macromolecular Colloidal Fine Dialysis
ΔC
Liquid membranes Pervaporation
ΔE
Electrodialysis
Reverse osmosis Nanofiltration Ultrafiltration ΔP Microfiltration Gas sep.
ΔT
Filtration
Membrane distillation
0.01 nm
1 nm
10 nm
100 nm
1000 nm 104 nm 1 μm 10 μm
105 nm 100 μm
Figure 1 The family of membrane processes (driving forces and applications size range).
Table 1
Typical properties of pressure-driven membranes Microfiltration
Ultrafiltration
Nanofiltration
Reverse osmosis
Pore size (nm)
50–10 000
1–100
B2
o2
Water permeability (l m2 h1 bar1) Operating pressure (bar) MWCO (Da)
4500
20–500
5–50
0.5–10
0.1–2.0
1.0–5.0
2.0–10
10–100
Not applicable
1000–300 000
4100
410
Bacteria, algae, suspended solids, turbidity Polymeric, inorganic
Bacteria, virus, colloids, macromolecules
Di- and multivalent ions, natural organic matter, small organic molecules Thin-film composite polyamide, cellulose acetate, other materials (Schafer et al., 2005)
Dissolved ions, small molecules
Targeted contaminants in water Membrane materials
Polymeric, some inorganic
Adapted from Winston and Sirkar (1992) and Mulder M (1996) Basic Principles of Membrane Technology, 2nd edn. Dordrecht: Kluwer.
Thin-film composite polyamide, cellulose acetate
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introduces the various module types. Section 4.11.6 provides the basic relationships describing membrane performance.
4.11.1.2 Role in Water Supply, Sanitation, and Reclamation The needs of the water industry (water supply, sanitation, and reclamation) are to remove various contaminants to make the water fit for purpose. Potable water requires removal of pathogens, organic species, and salts to specify limits. Some industrial waters have even tighter limits, for example, boiler feedwater, or ultrapure water for microelectronics. Wastewater treatment, such as sanitation, involves biological treatment and subsequent polishing. Water reclamation takes treated wastewater and upgrades it to high quality for industry or indirect potable reuse (IPR). Industrial water applications are very broad but could encompass water treatment to industry standards, water recovery and recycle, and treatment for discharge. It will be evident that the pressure-driven liquid-phase membrane processes, with the properties given in Table 1, provide the means to achieve the above separations. In many cases, the membrane process is used in combination with another unit operation (such as a bioreactor) in a hybrid membrane process. In other cases, low-pressure (MF/UF) membranes are used as pretreatment to high-pressure (NF/ RO) membranes in dual membrane processes. Figure 2 is a simplified generic flow sheet of a membrane process as applied in the water industry. In addition to the
Feedwater
Pretreatment
membrane separation, it is not uncommon to have both pretreatment and posttreatment steps. It should also be noted that contaminant removal inevitably results in reject (waste) streams that have to be dealt with. Input streams may also be involved in pre- and posttreatment. The following are brief descriptions of water industry applications of membranes. Table 2 summarizes the membrane process configurations for the range of applications and water sources.
4.11.1.2.1 Water treatment Table 2 identifies five water treatment applications (WT.1– WT.5), based either on low-pressure MF/UF membranes in hybrid processes or on tighter NF membranes. The water sources involved are low salinity. WT.1 is the most typical configuration for membrane-based water treatment plant (Kennedy et al., 2008); the posttreatment may be simpler depending on the natural organic matter (NOM) removal achieved upstream. Disinfection by chlorine is usually applied to maintain a residual in the distribution system. The lowpressure membranes are predominantly hollow fibers, either contained or submerged (see Section 4.11.5). In most applications the suspended solid content of the source is relatively low and this allows operation in dead-end mode (see Section 4.11.7), with cycles of filtration and backwash, typically over 30–60 min. The WT.2 option, combining membranes and powdered activated carbon (PAC), is used in cases where the
Membrane separation
Posttreatment
Product water
Reject Figure 2 Generic flow sheet of membrane process.
Table 2 Application
Membrane process configurations in the water industry Source water
Membrane process
Pretreatment or hybrid
Posttreatment
Target removals
Water treatment WT.1 Surface WT.2 Surface WT.3 Surface WT.4 Surface WT.5 Ground
MF/UF MF/UF NF NF NF
Coagulation PAC Filtration Coagulation þ filtration Filtration
AOT, BAC, Dis Dis Dis AOT Dis
NOM, turbidity, pathogens Taste/odor, trace organics NOM Trace organics, taste/odor Hardness
Desalination D.1 D.2
Brackish ground Seawater
RO RO
Filtration Media filtration or MF/UF
Dis Ca addition
Salinity Salinity
Reclamation R.1 R.2
Treated wastewater Wastewater
RO RO
MF/UF MBR
AOT AOT
Pathogens, trace organics Pathogens, trace organics
MF/UF
Screening
Dis
BOD, turbidity, pathogens
Membrane bioreactor MBR.1 Wastewater
AOT, advanced oxidation treatment (UV, etc.); BAC, biologically active carbon; BOD, biochemical oxygen demand; Dis, disinfection (chlorination, etc.); MBR, membrane bioreactor; MF, microfiltration; NOM, natural organic matter; PAC, powdered activated carbon; UF, ultrafiltration.
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
water may encounter trace organics or taste and odor (Lebeau et al., 1998). Applications WT.3–5 use NF membranes on low-salinity surface- or groundwaters. NF operates at higher pressures than MF/UF and also requires crossflow to control fouling, rather than dead-end operation, and consequently has a higher energy demand. WT.3 is popular in Norway (Wittmann and Thorsen, 2005), where raw waters are high in NOM and energy is relatively plentiful. WT.4 makes use of the ability of NF to remove organics of 100–200 Da size, and is exemplified by the treatment of river water in France to remove trace herbicides (Wittmann and Thorsen, 2005). In some locations, NF is also used on groundwater to remove calcium hardness.
4.11.1.2.2 Desalination Saline water sources range from brackish groundwater to seawater. In these applications, the key membrane process is RO. The role of pretreatment is to protect the RO membranes from various foulants (see Section 4.11.6.5), and posttreatment prepares the product water for discharge. Application D.1 (Table 2) is brackish water desalination. These are usually modest-size plants for local water supplies. A major challenge for this application is how to dispose of the plant reject and which is a high-salinity stream. Various options are available (Voutchkov and Semiat, 2008), including evaporation ponds and subsurface disposal. Application D.2 is desalination of seawater, which is effectively a limitless source. In this case, the high salinity and high osmotic pressure require an operating pressure of 60–70 bar, making seawater reverse osmosis (SWRO) more energy intensive than the other membrane/water options. Pretreatment has to be able to minimize various forms of fouling (inorganic, organic, and biofouling). Most SWRO plants with feedwater from ocean intakes opt for either media filtration, or increasingly use low-pressure membranes for pretreatment. In the case of SWRO, the reject (brine) stream is typically 50% of the intake flow, and is discharged back into the ocean, subject to adequate arrangements for rapid dispersion of salinity. Prior to discharge, the pressurized brine passes through energy recovery devices which can recover 495% of the pressure energy in the brine for pressurizing a portion of the feed. The product water is usually conditioned by addition of calcium ions to satisfy World Health Organization (WHO) requirements (Cote et al., 2008). A detailed description of desalination by RO is given in Voutchkov and Semiat (2008).
4.11.1.2.3 Water reclamation Municipal wastewater provides the second limitless source of water. Reclamation processes convert secondary effluent (R.1 in Table 2) or raw wastewater (R.2) into water of exceptionally high quality. Process R.1 is more common as it builds on the existing municipal wastewater infrastructure. Secondary effluent has relatively low suspended solids, total organic carbon (TOC), and salinity, which makes it very attractive as a source water. It also tends to be located close to where it could be reused. All major reclamation plants use dual membrane arrangements with the low-pressure (MF/UF) pretreatment membranes operating in dead-end cycles, similar
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to MF/UF water treatment plant, and providing very low solids feed to the RO. Due to the low-salinity feed, the RO operates at much lower pressures and with higher recoveries (75% vs. 50%) than SWRO. This means that the reclamation plant can produce high-quality water at approximately half the energy and costs of SWRO (see Section 4.11.7.3). A comprehensive review of membrane reclamation plant and comparison with SWRO can be found elsewhere (Cote et al., 2008). Posttreatment in R.1 is typically ultraviolet (UV) which provides an added barrier to virus and also oxidizes trace organic compounds, possibly present at ppb levels in the RO permeate. Flow sheet option R.2 uses a membrane bioreactor (MBR, see Section 4.11.1.2.4) in place of the conventional activated sludge processes (CASPs) combined with MF/UF. This option would probably be favored in a green field site due to the smaller foot print and the reported better-quality feed (lower TOC) to the RO (Cote et al., 2008). The high-quality water produced by the dual membrane reclamation process is suitable for demanding industrial applications and for IPR.
4.11.1.2.4 Wastewater MBR The CASP combines a wastewater bioreactor and a settling tank. The membrane bioreactor (MBR) replaces the settling tank with low-pressure membranes. The advantages of the MBR include an improved effluent quality (solids-free, potentially lower TOC) and significantly smaller foot print. The membrane separation allows the biomass mixed liquor to be increased to 10–20 g l1, compared with the o5 g l1 in the CASP. This improves organics removal and can reduce the excess waste sludge for disposal. Typically, MBRs operate with mixed liquors of 10–12 g l1, which is a compromise to avoid raised viscosity that would impair oxygen transfer. MBRs have either submerged (immersed) membranes in the bioreactor, or external side-stream membranes connected so that mixed liquor can be cycled through the modules. Submerged membranes use either hollow fibers in bundles (or curtains) or flat sheets, vertically aligned, operated under suction. Side-stream modules use large bore hollow fibers. Pretreatment for MBRs is usually fine screening to eliminate sharp objects that could damage the membranes. Posttreatment depends on the fate of the permeate, but at a minimum it would involve disinfection. Membrane fouling (see Section 4.11.6.5) is a major challenge in MBRs, due to the complex nature of the mixed liquor with its biomass floc, colloids, and macrosolutes. Fouling is mitigated by careful selection of the operating flux and by maintaining a vigorous crossflow induced by air sparging below the membranes or in the membrane loop. A comprehensive review of MBR fouling is available (Le-Clech et al., 2006). More detailed information on MBRs can be found in Judd (2006) and Lieknes (2009).
4.11.1.3 Status of Development Membrane technology in the water industry has grown from a research curiosity to mainstream applications over a brief period of 50 years. This section outlines the major developments and milestones and summarizes the status of the various applications.
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4.11.1.3.1 Seawater desalination The application of membranes in the water industry started in the 1960s, following the invention of the cellulose acetate (CA) RO membrane by Loeb and Sourirajan (1964). The CARO membrane was an alternative to thermal desalination of seawater and brackish water to provide drinking water. Over the next two decades, seawater RO steadily developed, but was considered relatively costly and energy intensive, suitable for niche applications. A major advance occurred in the late 1970s with the invention of the TFC membrane (see Section 4.11.3.3) by Cadotte (1977). The TFC significantly improved both water flux and salt retention and has become the basis for modern SWRO membranes. The ability to tune the chemistry of the TFC separating layer has allowed steady improvements in performance. Both CA-RO and TFC-RO membranes are produced in continuous flat sheets. The housing, or module (see Section 4.11.5) developed to package these membranes, is the spiralwound module (SWM). This is produced in standard sizes, typically 8 in (203 mm) diameter and 40 in (1016 mm) long, which allows interchange between products. Over the 30-year period from 1978 to 2008, the SWM has steadily improved, with a drop in real cost (1/12), an increased life (2.3 ), an improved water production rate (2.5 ), and a reduced salt transmission (1/7) (Birkett and Truby, 2007). The SWM is now the module of choice for RO, NF (and some large-scale UF in other industries). However from the mid-1960 s to the late 1990s, there were two SWR options, one the SWM, the other the hollow fiber RO membrane produced by Dupont and others. The HFRO was initially predominant in larger-scale SWRO plant. However, the SWM became more competitive because of the better performance of the TFC membrane, its slightly lower pretreatment requirements, and its evolution into an interchangeable standard module. The SWM is now the predominant option for large SWRO and reclamation plant. Current (2010) trends are to larger 16-in-diameter SWMs with 4 the water output per element. Membrane developments include the thin-film nanocomposite (TFNC), which incorporates highly permeable zeolite nanoparticles into its separation layer (Jeong et al., 2007). Energy demand and costs for SWRO have steadily declined (see Section 4.11.7), largely due to the introduction of high-efficiency energy recovery devices that recover the pressure energy in the RO concentrate stream. Desalination capacity by SWRO now exceeds that of thermal processes. Plant capacities of several hundreds of ML d1 are not uncommon and applications continue to grow due to increasing water scarcity and population growth.
4.11.1.3.2 Water reclamation Water reclamation with membranes commenced at a significant scale in the 1970s with Water Factory 21 in Orange County California. This plant took secondary-treated municipal effluent and used physical/chemical pretreatment prior to RO. Today, this plant uses low-pressure membrane (UF) pretreatment as do all the major reclamation plants. Water reclamation processes provide significant augmentation of water supplies in some locations. For example, the Singapore NeWater plants (Seah et al., 2008) provide about 20% of
Singapore’s water supply. The high-quality water produced by water reclamation plant tends to go to industrial usage, with a portion going to IPR. The likely trend will be for more membrane-based IPR schemes, due to lower cost and energy demand than SWRO and increased confidence in its water quality.
4.11.1.3.3 Water treatment Water treatment with low-pressure membranes is now well established, but the growth only started in the early 1990s. Prior to this, membranes were considered too expensive for the production of a low-cost product, such as water – the exception being SWRO in niche areas. Several factors have contributed to the rapid growth in low-pressure membrane water treatment, including tighter treatment standards in the US prompted by a serious cryptosporidium outbreak in 1993, a concerted effort by manufacturers to reduce membrane plant costs and the adoption of dead-end with backwash operation (see Section 4.11.7.1) with lower energy demand. The current low-pressure membrane options are hollow fibers either in submerged or in pressurized modules (see Section 4.11.5), and both concepts appear to be equally popular. There is a small, but growing, interest in the use of ceramic membranes for water treatment, particularly in Japan. Claimed benefits are higher fluxes and longer membrane lifetimes. Higher-pressure NF membranes are also well established for effective removal of hardness and organic contaminants, including trace pollutants such as herbicides and endocrine disruptors (EDCs). NF is a more expensive option than low-pressure UF or MF, but has application in special cases (Wittmann and Thorsen, 2005).
4.11.1.3.4 Membrane bioreactors MBRs for wastewater treatment have been around since the late 1960s. In the early period, a major incentive was to increase the biomass mixed liquor (MLSS) to high levels (420 g l1) to reduce excess sludge production. However, this exacerbated fouling and reduced oxygen transfer. In addition, the high cost of membranes favored relatively high fluxes that required vigorous crossflow and high-energy input to control fouling. As a result, the MBR was another niche membrane application until the early 1990s. The significant growth in MBRs has been due to innovations in design and operation. First, the lower cost of membranes allowed a lower design flux with less intrinsic fouling. Second, the use of a two-phase flow, such as bubbling, has been found to control fouling with substantially reduced energy costs. Third, the use of submerged (immersed) membranes provided an alternative MBR configuration, besides allowing retrofitting to existing plants. The current status is that MBR applications are becoming widespread in industry and increasingly in municipal use. There tend to be more, but smaller, plant in industry and some large (4100 ML d1) municipal plant (Lieknes, 2009). MBR designs are not standardized and popular options include submerged membranes, with either hollow fibers or flat sheets, or pressurized side-stream vessels with hollow fibers. All applications use the two-phase flow to control fouling. While current MBRs are all aerobic processes, there is a significant R&D effort in anaerobic MBRs (AnMBRs), with the
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
incentive of energy recovery in biogas. Future applications of AnMBRs can be anticipated.
4.11.2 Membrane Types and Properties A pressure-driven membrane is a selective barrier for separation. Its selectivity and permeability depend strongly on its pore characteristics (pore size, pore-size distribution, and porosity). Thus, membrane pore structure is one of the most important properties affecting membrane performance. Depending on the pore structure, pressure-driven membranes can be classified into RO, NF, UF, or MF membranes. Other important membrane properties include membrane hydrophilicity, surface charge, roughness, etc. These properties are discussed in this section.
•
4.11.2.1 Membrane Types and Important Membrane Properties Membrane pore structure is one of the most important properties of a membrane, as this largely determines the selectivity as well as the permeability of a membrane. Pressuredriven membranes can be classified into porous and nonporous membranes on the basis of membrane pore size (Table 1). MF and UF membranes are porous membranes that can be operated at low pressures. For this reason, MF and UF membranes are also frequently referred to as low-pressure membranes. The pore structure of MF and UF membranes can be observed using an electron microscope. In contrast, the rejection layer of a typical RO membrane does not appear to have any visible pores under an electron microscope. RO membranes are believed to be nonporous. In some literature, subnanometer pore size is reported for RO membranes based on their rejection properties of dissolved ions and small organic molecules. Finally, an NF membrane is an intermediate between a tight UF membrane and a loose RO membrane. Both RO and NF membranes require relatively high operating pressure, and they are also referred to as high-pressure membranes. The performance parameters and pore size range for MF, UF, NF, and RO membranes are summarized in Table 1:
•
MF membranes typically have pore sizes ranging from 0.05 to 10 mm. Corresponding to their relatively large pore sizes, MF membranes have high permeability (4500 l1 m2 h1 bar1) and can be operated in a low-pressure range (typically from 0.1 to 2.0 bar). MF membranes are used to retain particulates whose size is greater than membrane pore size. They can be fabricated from both polymeric and
Symmetric MF membrane
307
inorganic materials with either symmetric or asymmetric structures (Figure 3). For symmetric MF membranes, the pore diameters do not vary over the entire cross section of the membrane, and the thickness of the membrane determines its flux. Depending on the manufacturing method used, MF can also be asymmetrically structured, but the pores of the active layer are not much smaller than those of the supporting substructure. UF membranes have pore sizes ranging from 1 to 100 nm. This pore size range allows them to be used for removing bacteria, viruses, colloids, and macromolecules from a feedwater. The selectivity of a UF membrane is commonly represented by its molecular weight cutoff (MWCO), which is defined as the molecular weight of the solute that achieves a 90% rejection by the membrane. The MWCO of typical UF membranes is in the range of 1–300 kDa. A larger MWCO indicates that the membrane has a lower rejection ability and that it has a larger pore size. The MWCO of a membrane can be used to determine the pore size of the membrane by relating the molecular radius (a) of the solute molecules to its molecular weight (Mw):
a ¼ 0:33M0:46 w
ð1Þ
(valid for dextran, a in A˚ and Mw in Da, from Aimar et al. (1990)). UF membranes typically have an asymmetric structure (Figure 3) to maximize its membrane permeability. A very thin (0.1–1 mm) active or selective skin layer with fine pores is supported by a highly porous 100–200-mm-thick substructure. The pore diameters may increase from one side of the membrane (the skin layer) to the other (supporting sublayer) by a factor of 10–1000 (Strathmann, 1990). Its separation characteristics and mass flux are determined mainly by the feature of the skin layer (pore size, pore-size distribution and thickness, etc.), while the porous sublayer serves only as a mechanical support. Because of such an asymmetric structure, UF membranes gain excellent separation performance and considerable strength from the fine pore size of the thin active layer, and encounter little mass transfer resistance from the open supporting substructure. Typical permeability ranges from 20 to 500 l m2 h1 bar1. The normal operating pressure ranges from 1.0 to 5.0 bar.
•
RO membranes are able to remove small organic molecules and dissolved ions, including monovalent ions such as Naþ and Cl. These membranes have subnanometer pores and their separation properties are generally reported in terms of water permeability and sodium chloride rejection. They
Integral asymmetric UF membrane
Figure 3 Structures of symmetric and asymmetric membrane. MF, microfiltration; UF, ultrafiltration.
308
•
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
can be further subdivided into SWRO membranes and brackish water RO (BWRO) membranes. SWRO membranes are used for seawater desalination, and they generally have high sodium chloride rejection (499%). Such high rejection is typically achieved using a highly crosslinked dense rejection layer which tends to have a low water permeability (o1 l m2 h1 bar1). High pressure (460 bar) is required for SWRO operation to overcome the osmotic pressure of seawater as well as the large hydraulic resistance of the membrane. Compared to SWRO membranes, BWRO membranes have lower sodium chloride rejection (495%) but higher water permeability (1–10 l m2 h1 bar1). Relatively lower operating pressure (10– 20 bar) is required. Some BWRO membranes are marketed as low-energy or high-flux RO membranes. NF membranes are similar to RO membranes in that they can retain dissolved ions as well as some small organic molecules. The difference between NF and RO is that NF membranes typically have low rejection to monovalent ions such as Naþ (10–90%). The rejection of dissolved ions depends strongly on their valence, and divalent and multivalent ions tend to be better rejected. Due to their ability to effectively remove calcium and magnesium ions, NF membranes can be used for water softening. Many NF membrane manufacturers also provide rejection data on solutes other than sodium chloride, such as magnesium chloride, magnesium sulfate, or glucose. NF membranes can be operated at significantly lower pressure levels (o10 bar) compared to those for RO membranes due to their higher water permeabilities (5–50 l m2 h1 bar1).
There are two types of structure for RO and NF membranes. One is an integral asymmetric structure formed by the phaseinversion method, and the other is a TFC structure formed by interfacial polymerization method (see Section 4.11.3.3). An integral asymmetric RO/NF membrane is made of one polymer material and has a thin, permselective skin layer with a thickness of 0.1–1 mm supported by a more porous sublayer. The membrane’s flux and selectivity are determined by the dense skin layer, while the porous sublayer has little impact on the membrane separation properties. In contrast, a TFC RO/ NF membrane is made of two or more polymer materials. The most important TFC composite membranes are made from crosslinked aromatic polyamide by the interfacial polymerization method, on a microporous polymer such as polysulfone (PS) support layer, followed by a reinforcing fabric (see Section 4.11.2.2). In addition to the membrane separation properties, other important membrane properties include the following. Chemical, mechanical, and thermal stability. A good membrane shall be mechanically stable. For many applications, pH, temperature, and chlorine tolerance are important considerations. For example, RO membranes synthesized from CA are susceptible to hydrolysis and biodegradation. These membranes can only be used within a narrow pH and temperature range, which is one of the major causes for phasing out of this type of RO membrane. Modern RO membranes are typically based on polyamide chemistry with a TFC structure. Unfortunately, TFC RO membranes have low chlorine resistance (Kwon et al., 2006, 2008). Where chlorine disinfection is
needed for biofouling control, the free chlorine level needs to be carefully controlled to prevent unacceptable membrane damage. Hydrophilicity. A hydrophilic membrane is water like, that is, water has a strong affinity to its surface and it has a tendency to wet the surface. In contrast, a hydrophobic membrane does not interact favorably with water. The relative hydrophilicity/hydrophobicity can be determined by contactangle measurements (Section 4.11.4). In general, a hydrophilic membrane surface is preferred as it tends to enhance water permeability and reduces membrane fouling propensity. A hydrophobic membrane is preferred for some special applications (such as MD and some membrane contactors) where transmission of liquid water through the membrane needs to be prevented. Surface charge. A membrane surface can gain surface charge due to either its charged functional groups or preferential adsorption of some specific ionic species. For example, most TFC polyamide RO membranes are negatively charged at neutral pH due to the presence of carboxylic groups (–COO) (Tang et al., 2007a). Membrane surface charge plays an important role in fouling. A positively charged particle may have a strong tendency to deposit on a negatively charged membrane surface due to electrostatic attraction (Jones and O’Melia, 2000). In addition, rejection of ionic species as well as membrane permeability can be affected by surface charge for RO, NF, and tight UF membranes as a result of electrostatic interaction (Donnan exclusion effect; see Schafer et al. (2005)). For example, Childress and Elimelech (2000) reported that both salt passage and water flux were maximum at the pore isoelectric point (BpH 5) for an NF membrane. Membrane surface charge has also been used to correlate the transport of some trace organic solutes (Kimura et al., 2003). Surface roughness. Surface roughness has been reported as an important parameter for RO and NF membrane fouling. A rougher membrane surface tends to promote fouling likely due to reduced shear force over the membrane surface and increased membrane nonhomogeneity (e.g., nonuniform flux distribution over a membrane surface) (Vrijenhoek et al., 2001; Tang and Leckie, 2007).
4.11.2.2 Membrane Properties for RO and NF Membranes The physiochemical properties, such as surface roughness, hydrophobicity, and rejection properties, of RO and NF membranes strongly depend on the chemistries forming these membranes (Petersen 1993; Tang et al., 2007a, 2009a, 2009b). There are mainly two types of RO membranes in the market: (1) asymmetrical CA-RO membranes formed by phase inversion and (2) TFC polyamide RO membranes formed by an interfacial polymerization process (Petersen, 1993). CA membranes, though they have hydrophilic and smooth membrane surfaces, have low resistance to hydrolysis and biodegradation. In addition, their separation properties (permeability and rejection) are inferior to modern TFC-PA membranes. A comparison between typical CA and polyamide RO membranes is presented in Table 3. Both CA and polyamide can be used to form NF membranes. In addition, other polymers (e.g., polyvinyl alcohol (PVA) and sulfonated PS) and inorganic materials (e.g., some metal oxides) can also be
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Table 3
309
Comparison between cellulose acetate (CA) and thin-film composite (TFC) reverse osmosis (RO) membranes
Parameter
CA RO membrane
TFC polyamide RO membrane
Permeability
Low
High (B5 l m2 h1 bar1 for brackish water RO)
NaCl rejection (%)
85–98
95–99.9
Surface hydrophilicity
Very hydrophilic
Less hydrophilic
Surface roughness
Smooth
Rough surface with valley-and-ridge structures
Maximum temperature (1C)
30
45
Stable pH range
4.5–6.5
3–10 (2–11 for some membranes, e.g., with a polyvinyl alcohol surface coating)
Resistance to hydrolysis
Good
Chlorine resistance
Low. Unstable at pHo4.5 or pH46.5, and accelerated hydrolysis at high temperature Stable at low levels (o1 ppm)
Low tolerance to free chlorine (o0.1 ppm)
Resistance to biodegradation
Low
Relatively good
Polyamide ~ 0.05 – 0.3 μm
Polysulfone ~ 20 – 50 μm
Backing layer ~ 200 μm Figure 4 Schematics of a thin-film composite polyamide membrane. The composite PA membrane typically comprises three distinct layers – a thin, dense selective layer (the polyamide layer) of 50–300 nm in thickness, a polysulfone support layer of 20–50 mm in thickness, and a nonwoven fabric backing layer.
used for NF synthesis. A thorough review of RO and NF chemistry is available elsewhere (see Petersen (1993)) for RO and (Schafer et al., 2005) for NF membranes. The current section focuses on the structure and properties of TFC polyamide RO and NF membranes (Petersen, 1993). A typical TFC polyamide membrane comprises a dense polyamide rejection layer of B100 nm in thickness on top of a microporous PS or polyethersulfone (PES) support (Figure 4). The PS/PES layer, typically casted on a nonwoven fabric layer of 100–200 mm in thickness, provides a mechanical support to the rejection layer. Among the three layers, the polyamide rejection layer is the most critical one, as most of the membrane properties (e.g., permeability, rejection, surface charge, roughness, and surface hydrophilicity) are determined by this ultrathin dense rejection layer (Tang et al., 2007a; Petersen, 1993). In addition, the mechanical properties of the different layers are important, as RO and NF membranes need to withstand high pressures. The PS or the polyamide layer may deform mechanically and become more compact under high pressure. This reduces the membrane permeability with time, a phenomenon known as membrane compaction. Most commercial TFC-RO membranes for water treatment are formed by interfacial polymerization of aromatic amine
monomers (such as m-phenylenediamine (MPD) in an aqueous solution) and aromatic acid chloride monomers (such as trimesoyl chloride (TMC) in an organic solvent). Since the discovery of this reaction scheme by Cadotte and the development of the first commercial fully aromatic TFC RO membrane FT30 by FilmTec&, the reaction scheme and its variations have been widely used to prepare most commercial TFC-RO membranes (Petersen, 1993). In general, fully aromatic TFC RO membranes formed in this way have a high degree of crosslinking (i.e., the fraction of crosslinked repeating units; see Figure 5) (Tang et al., 2007a). Increased crosslinking tends to increase salt rejection but decrease water permeability. Commercial BWRO membranes (e.g., XLE from Dow FilmTec and ESPA3 from Hydranautics) have a water permeability in the range of 4–8 l m2 h1 bar1 and sodium chloride rejection 495% (Table 4). Seawater RO membranes have lower water permeability but much higher salt rejection (499%). It is worth noting that rejection of a solute (R) is not an inherent property of an RO membrane, as it also depends on the operating conditions such as applied pressure. A more fundamental property of a nonporous RO membrane is the solute permeability coefficient (B). The NaCl permeability coefficient is also tabulated in Table 4 for some commercial RO membranes. The solute permeability coefficient can be determined from rejection test results via
B ¼ AðDP DpÞðR 1 1Þ
ð2Þ
Typical fully aromatic polyamide RO membranes formed by TMC and MPD have rough membrane surfaces, with a rootmean-square (RMS) roughness (ridge-and-valley) on the order of 100 nm (Table 4). The surface is negatively charged at neutral pH due to the deprotonation of carboxylic (–COOH) functional group (Tang et al., 2007a; Childress and Elimelech, 1996). The typical contact angle for an unmodified membrane is 40–501. Some posttreatment steps, such as the application of a coating layer or additives, might be involved to protect the
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O
O
C
C
C
H N
H N
O
O
C
C
O
H N
NH
COOH
NH
n
1–n
(a)
O
O
C
C
C
N
N
O
O
C
C
O
N
N
COOH n
(b)
1–n
Figure 5 Typical chemistry for interfacially formed TFC-RO and NF membranes. (a) Fully aromatic polyamide based on trimesoyl chloride and m-phenylenediamine. (b) Semi-aromatic polyamide based on trimesoyl chloride and piperazine. Modified from Tang CY, Kwon YN, and Leckie (2009b) Effect of membrane chemistry and coating layer on physiochemical properties of thin film composite polyamide RO and NF membranes. I. FTIR and XPS characterization of polyamide and coating layer chemistry. Desalination 242: 149–167; and Petersen RJ (1993) Composite reverse-osmosis and nanofiltration membranes. Journal of Membrane Science 83: 81–150.
Table 4
Physiochemical properties of thin-film composite (TFC) polyamide reverse osmosis (RO) and nanofiltration (NF) membranes
Chemistry
Type
Membrane
A (l m2 h1 bar1)a
MPD þ TMC, no coating
SWRO BWRO BWRO BWRO
SWC4 XLE LE ESPA3
0.80 6.04 4.29 7.52
MPD þ TMC, PVA coating
SWRO BWRO BWRO BWRO
SW30HR LFC1 LFC3 BW30
0.85b 3.96 2.81 3.96
MPD þ TMC, no coating
NF NF
NE90 NF90
9.04 11.2
PIP þ TMC, no coating
NF NF
HL NF270
12.8 14.5
B for NaCl (l m2 h1)
R for NaCl (%)a
RMS roughness (nm)
Contact angle (1 )
Zeta potential (mV) at pH 9
0.11 3.02 2.60 5.58
99.0 96.5 95.8 94.9
135.6 142.8 95.7 181.9
48.8 46.4 47.2 43.1
–20.9 –27.8 –26.1 –24.8
0.11b 1.52 0.59 1.17
99.6b 97.3 98.5 97.9
54.4 135.8 108.4 68.3
30.9 20.1 22.8 25.9
–1.7 –13.2 –6.5 –10.1
91.5 94.4
72.4 129.5
46.3 44.7
–21.0 –37.0
21.3 56.9
7.2 9.0
27.5 32.6
–26.0 –41.3
11.6 9.17 653 152
a
The permeability of all membranes except BW30 was evaluated using ultrapure water at an applied pressure of 1380 kPa (200 psi). Rejection of all membranes except BW30 was determined using a 10 mM NaCl solution at pH 7. b The permeability and rejection for BW30 was calculated from the membrane manufacturer’s specification. The testing pressure for BW30 was 55 bar for a feedwater containing 32 000 mg l1 NaCl. MPD, m-phenylenediamine; PIP, piperazine; RMS, root-mean-square; TMC, trimesoyl chloride. Adapted from Tang CY, Kwon YN, and Leckie JO (2009a) Effect of membrane chemistry and coating layer on physiochemical properties of thin film composite polyamide RO and NF membranes II. Membrane physiochemical properties and their dependence on polyamide and coating layers. Desalination 242: 168–182.
membrane surface and/or improve membrane properties. Membrane surface properties, such as hydrophilicity and surface charge density, are largely determined by the top-most layer. Consequently, the application of a surface coating can greatly affect these properties. For example, PVA, a neutral and hydrophilic polymer, has been widely used for coating RO membranes to achieve improved membrane properties. The
PVA-coated RO membranes are less rough and less charged (Table 4). They are significantly more hydrophilic, with a contact angle of only 20–301 (Table 4 and Figure 6), and thus they are expected to be less prone to membrane fouling. Many commercial membranes marketed as low-fouling composite RO membranes are PVA coated. Some examples are the low fouling composite (LFC) series from Hydronautics and
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
BW30 from Dow FilmTec (Tang et al., 2009a). Commerical PVA-coated membranes tend to have better rejection than uncoated ones, but their permeability is lower probably due to the additional hydraulic resistance from the coating layer (Tang et al., 2009a). Other commonly used coating materials and additives reported in the literature include poly(ethylene oxide-b-amide) (Pebaxs), chitosan, and sodium alginate (Schafer et al., 2005; Louie et al., 2006). Similar to TFC RO membranes, TFC NF membranes can be formed by the fully aromatic polyamide chemistry (such as MPD reacting with TMC). Some commercial examples include NF90 from Dow FilmTec and NE90 from Saehan Industries. With a less crosslinked polyamide compared to typical RO membranes, these TFC-NF membranes have lower sodium chloride rejection (B90–95%) and higher water permeability of about 10 l m2 h1 bar1. Similar to fully aromatic RO membranes, MPD þ TMC-based NF membranes have negatively charged membrane surfaces with ridge-and-valley type of roughness. Their contact angle is also similar to MPD þ TMC-based RO membranes. Some other TFC-NF membranes are semi-aromatic where an aliphatic amine monomer is used with the aromatic TMC (Petersen, 1993). The most widely used amine monomer for semi-aromatic TFC-NF membranes is piperazine (PIP) (Petersen, 1993; Schafer et al., 2005). Some commercial examples of PIP þ TMC-based NF membranes include NF270 from Dow FilmTec and HL from GE Osmonics. Poly (piperazinamide) NF membranes have higher fluxes, lower
311
rejections, and smoother and more hydrophilic membrane surfaces than fully aromatic NF membranes. In Table 4, it is apparent that there is a strong trade-off between water permeability and salt rejection – a more water permeable membrane tends to have lower rejection and higher solute permeability.
4.11.2.3 Membrane Properties for MF and UF Membranes Porous MF and UF membranes can be formed by a wide range of structures, materials, and formation methods (refer to Section 4.11.3). Therefore, the properties of MF and UF membranes can vary significantly. By far, the most important consideration for porous membrane selection is the pore structure (e.g., pore size, pore-size distribution, porosity, tortuosity, and thickness of the active separation layer), as their flux and retention properties are mainly determined by these properties. In addition, membrane flux and rejection can be marginally affected by some material properties, including hydrophilicity. In general, hydrophilic membranes are preferred due to their lower fouling tendency and higher water permeability. There are many different pore structures for porous membranes (Figure 7). A track-etched MF membrane typically has straight cylindrical pores. These pores are perpendicular or slightly oblique to the membrane surface. The membrane permeability for cylindrical-pore membranes can be
BW30 membrane
EPSA3 membrane
Figure 6 Contact-angle measurement for membranes BW30 and ESPA3. The PVA-coated membrane BW30 is much more hydrophilic than the uncoated membrane ESPA3. Unpublished photos.
Cylindrical pores
Stacked spheres
Spongy structure
Cylindrical pores perpendicular or oblique to membrane surface (e.g., track-etched MF) Pores typically not interconnected
Interconnected pores form by the space between nearly spherical particles (inorganic membranes, foulant layer)
Interconnected pores. Most phaseinversion polymeric membranes have this type of pore structure
Transport equation:
Transport equation:
Hagen–Poiseuille Equation
Kozeny–Carman equation
Transport equation: Hagen–Poiseuille equation, or Kozeny–Carman equation
Figure 7 Types of membrane pore structures. Modified from Mulder M (1996) Basic Principles of Membrane Technology, 2nd edn. Dordrecht: Kluwer.
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Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis 4.11.3.1 Polymeric Membrane Materials
determined by the Hagen–Poiseuille equation:
Lp ¼
er2p 8Ztlm
ð3Þ
where Lp is the water permeability of porous membranes, e the membrane porosity, rp the pore radius, Z the viscosity of water, t the tortuosity of the membrane, and lm the thickness of the rejection layer. Another type of well-defined pore structure closely resembles stacked spheres (Figure 7), a structure typical for inorganic porous membranes prepared from spherical particles. Interconnected pores are formed by the space between the particles. The same pore structure can also be used to describe foulant cake layers formed by the deposition of colloids or suspended particles. Stacked-spheres pore structure can be modeled by the Kozeny–Carman equation:
Lp ¼
e3 Kð1 eÞ 2 S 2 Zlm
ð4Þ
where K is the Kozeny–Carman coefficient and S the specific surface area of the particles. For perfect spherical particles, S ¼ 3/rp, K ¼ 5, and e ¼ 0.4. Many other porous membranes have much more complicated pore structures. For example, polymeric membranes prepared via phase inversion may have a spongy structure. The pores tend to be highly interconnected with a high tortuosity and a wide size distribution. A lognormal pore-size distribution is sometimes assumed for modeling purpose (Aimar et al., 1990). The Hagen–Poiseuille and/or Kozeny–Carman equation are widely used to approximate the transport properties of this type of membrane. The pore structural parameters of porous membranes prepared by different methods are summarized in Table 5.
4.11.3 Membrane Materials and Preparation Polymers are the most popular materials used for membrane fabrication. Being membrane materials, the polymers should demonstrate thermal and chemical stabilities, good mechanical strength, and ability to form flat sheet or hollow fiber membranes easily. The two major techniques for membrane preparation include the phase-inversion process and interfacial polymerization, which are widely used for commercial membrane productions.
Table 5
Most of materials such as polymer, ceramic, metal, carbon, and glass can be used to make membranes. Among these, polymeric materials are the most popular ones used for membrane fabrication. Being membrane materials, the polymers should demonstrate thermal stability over a wide range of temperatures and chemical stability over a range of pH, and possess good mechanical strength. In addition, they can also be processed into flat sheet or hollow fiber membranes easily. Some commercial and representative polymeric membrane materials are introduced in Table 6 (Ren and Wang, 2010). Commercial UF/MF membranes can be made by various materials ranging from fully hydrophilic polymers, such as CA, to fully hydrophobic polymers, such as polypropylene (PP). PS, PES, polyacrylonitrile (PAN), and polyvinylidene fluoride (PVDF) are between the two extremes. A hydrophilic surface tends to resist attachment due to absorption by organics, but it has the disadvantage of being less robust compared with hydrophobic membranes. In order to reduce membrane fouling tendency, the hydrophobic polymers are modified through various approaches such as the use of additives as pore formers, or blending with a hydrophilic polymer, or posttreatment. Figure 8 provides the relative hydrophilicity of commonly used polymeric materials. The pros and cons of different membranes made by different polymers are summarized in Table 7 (Pearce, 2007).
4.11.3.2 Hollow Fiber Preparation 4.11.3.2.1 Mechanism of membrane formation Hollow fiber membranes can be made either by a phase-inversion process or by a melt spinning plus stretching process (Mulder, 1996; Strathmann, 1990). Normally, UF membranes are produced by the phase inversion, which makes the membrane a precisely controlled asymmetric structure by varying the pore size over a wide range, while MF membranes can be fabricated by either process. Phase inversion refers to the process by which a polymer solution (in which the solvent system is the continuous phase) inverts into a swollen three-dimensional macromolecular network or gel (where the polymer is the continuous phase) (Kesting, 1985). The essence of phase inversion is the appearance in a polymer solution of two interdispersed liquid phases (a polymer-rich phase and a solvent-rich phase) due to the change of the state of the polymer solution caused by the
Pore structure of different types of porous membranes
Pore structure
Pore size (mm)
Pore-size distribution
Porosity
Tortuosity
Symmetry
Fabrication method
Microfiltration Stacked spheres Stretch pores Cylindrical pores Spongy structure
0.1–20 0.1–3 0.05–5 0.1–10
Narrow Wide Nearly uniform Wide, log-normal
0.1–0.2 High, up to 0.9 Low, up to B0.1 0.3–0.7
Low Straight pores Straight pores Tortuous pores
Symmetrical Symmetrical Symmetrical Symmetrical/asymmetrical
Sintering Stretching Etching Phase inversion
Ultrafiltration Spongy structure
0.001–0.1
Wide, log-normal
0.01–0.2
Tortuous pores
Asymmetrical
Phase inversion
Adapted from Mulder M (1996) Basic Principles of Membrane Technology, 2nd edn. Dordrecht: Kluwer.
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Table 6
313
Commercial and representative polymeric membrane materials
Polymer Cellulose and cellulose acetate (CA)
Structure
Properties
Cellulose is highly hydrophilic and OH
OH
O
HO HO
OH
HO O
O HO
O
OH
HO O
OH O
OH
OH
OH
O
CH
CH
O
CH
CH
CH
O n
CH CH
OCCH3
CH2
OCCH3
CH2
O
CH3CO O
O
membrane preparation.
Cellulose acetate, diacetate, triacetate and
O
CH
n
CH CH
OH OCCH3
Cellulose is mainly used for dialysis their blends are widely used to make MF, UF, and RO membranes. Cellulose and cellulose acetate membranes are susceptible to hydrolysis and microbial attack. They are only stable over limited pH range between 4 and 6.5.
n
Cellulose O
crystalline.
OH
O
OCCH3 O
Cellulose triacetate
Cellulose acetate
PS is an amorphous polymer. It belongs to the group of high-
Polysulfone (PS) O
O
SO2
performance polymers with excellent chemical and thermal stability. PS is mainly used to form UF, MF, and gas separation membranes. PS is also used to form the porous support layer of many RO, NF, and some gas separation membranes.
Udel polysulfone
PES membranes have very high chemical
Polyethersulfone (PES)
and thermal stability.
Like PS material, PES membranes are affected by aromatic hydrocarbons or ketones. PES membranes are slightly less hydrophobic than PS membranes. PES membranes are mainly used in UF, MF, and dialysis.
Radel A polyethersulfone
Radel H polyethersulfone Polyacrylonitrile (PAN)
CH2
CH
PAN possesses superior resistance to
n
C
hydrolysis and oxidation.
It is mainly used to prepare UF membranes
N
and porous supports of composite membranes. Polyetherimide (PEI)
PEI is an amorphous thermoplastic with
O O N
H3C
O
CH3
N
O
O O
n
characteristics similar to the related plastic polyether ether ketones (PEEKs). PEI cannot be used in contact with chloroform and dichloromethane. It is a good material for the fabrication of the integrally asymmetric membranes for gas separation and pervaporation. PEI is also used to fabricate the support of composite flat sheet membranes. (Continued )
314 Table 6
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Continued
Polymer Polyamide (PA)
Structure
Properties
H N
CH2
O
H
C
N
x-1
CH2
O
O
N
C
CH2
x
n
O
A PA is a polymer containing monomers of
C
The basic aliphatic polyamides are referred
y-2
amides. n
as nylons which possess good thermal stability and mechanical strength, and are resistant to many organic solvents. PA is used as the thin dense layer for RO and NF, but it has lower chlorine tolerance The porous polyamide membranes have been commercialized for many years.
Aliphatic polyamides (nylon x or nylone x,y)
Polyimide (PI)
O
O
O
C
C
C
N
Polyimides exhibit excellent thermal and
CH3
C
C
O
O
chemical stability because of their high glass transition temperature. P84 is an amorphous commercial copolyimide with excellent resistance to many organic solvents. Matrimide 5218 is another commercial polyimide. Both polymers can be used as a material for making gas separation and NF membranes.
H N
C H 80%
n
20%
Copolymide P 84 O
O
O
C
C
C
N
CH3 N
C
C
O
O
CH3
CH3
n
Matrimid 5218
Polyether ether ketones (PEEKs)
PEEK: O
O
C
O n
Polycarbonate (PC)
CH3 C CH3
Polyvinylidene fluoride (PVDF)
PEEK has exceptional heat and chemical
SPEEK: O
stability.
C n
SO3H
The high insolubility in common solvents makes PEEK membranes successfully being used in chemical processes as a solvent-resistant membrane. The sulfonated PEEK (SPEEK) is soluble in common solvents, which can be used for the preparation of hydrophilic membranes or ion-exchange membranes.
PC is a transparent thermoplastic with
O
high-performance properties.
O C O n
It is mainly used for track-etched membranes with well-defined pore structures and very good mechanical strength. PC can also be used to make UF and MF membranes by the phase-inversion process.
H
F
PVDF is semi-crystalline with a very low
C
C
It is the most popular and available
H
F n
hydrophobic membrane material to be used for making MF by phase-inversion process. It has excellent chemical resistance and thermal stability. It is resistant to most inorganic and organic acids and tolerant to a wide pH range.
glass transition temperature.
(Continued)
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315
Continued
Polymer
Structure
Polypropylene (PP)
Properties
CH2
CH
PP membrane is normally hydrophobic
n
with high chemical stability.
CH3
Hydrophobic PP membrane is ideal for the filtration of aggressive solvents.
PP membranes are much cheaper than PTFE membranes. Polytetrafluoroethylene (PTFE)
F
F
C
C
F
F n
PTFE is highly crystalline and demonstrates a very high resistance to chemical attack. It cannot be formed by phase-inversion techniques. PTFE membranes are formed by melt extrusion followed by stretch cracking. The membranes are hydrophobic and need pre-wetting with a nonpolar solvent before use.
Adapted from Mulder M (1996) Basic Principles of Membrane Technology, 2nd edn. Dordrecht: Kluwer, and from Ren JZ and Wang R (2010) Preparation of polymeric membranes. In: Wang LK, Chen JP, Hung YT, and Shammas NK (eds.) Handbook of Environmental Engineering, vol. 13, ch. 2. Totowa: Humana Press.
Hydrophilic
CA
Hydrophobic
PES
PAN
PS/PVDF
PP
PTFE
Table 7
Pros and cons of different membranes
Polymer
Properties
CA
Good permeability and rejection characteristics Susceptible to hydrolysis Limited pH resistance Chlorine tolerant and fouling resistant
PES, PVDF, PS, PAN
Ability to modify properties through polymer blend Good strength and permeability PVDF best for flexibility and use with air scour PES best for polymer blending and UF rating
PP
Susceptible to oxidation Limited blend capability
CA is naturally hydrophilic PS, PES, PAN, and PVDF are naturally quite hydrophobic, but can be blended with additives and pore formers to make a moderately hydrophilic membrane PP and PTFE are hydrophobic, and are difficult to modify
Figure 8 Relative hydrophilicity of commonly used polymeric materials. CA, cellulose acetate; PAN, polyacrylonitrile; PES, polyethersulfone; PS, polysulfone; PVDF, polyvinylidene fluoride; PP, polypropylene; PTFE, polytetrafluoroethylene. Modified from Pearce G (2007) Introduction to membranes: Membrane selection. Filtration and Separation 44: 35–37.
alteration of its surrounding environment or operating conditions, followed by crystallization, gelation, or vitrification. In other words, a liquid polymer solution is precipitated into two phases: (1) a polymer-rich phase that will form the matrix of the membrane; (2) a polymer-poor phase that will form the membrane pores in an unstable nascent membrane structure. The porous asymmetric membrane morphology is then fixed according to the subsequent solidification process. There are different approaches to make the polymer solution precipitate, such as cooling, immersion in a nonsolvent coagulant bath, evaporation, and vapor adsorption. Depending on the change of the operating parameters that induce the phase inversion, two different separation mechanisms are involved:
•
Thermally induced phase separation (TIPS). The precipitation is achieved by decreasing the temperature of the polymer
CA, cellulose acetate; PAN, polyacrylontrile; PES, polyethersulfone; PP, polypropylene; PS, polysulfone; PVDF, polyvinylidene fluoride. Adapted from Pearce G (2007) Introduction to membranes: Membrane selection. Filtration and Separation 44: 35–37.
•
solution. This process can be used for PVDF membrane preparation. Diffusion-induced phase separation (DIPS). Diffusional mass exchange, because of the contact of the polymer solution with a nonsolvent, leads to a change in the local composition of the polymer film and then precipitation is induced. CA, PS, PES, PVDF, and PAN membranes are made by this method.
TIPS. It is one of the main approaches for the preparation of microporous membranes. In the TIPS process, a polymer is dissolved into a solvent or a mixture of solvent and nonsolvent, and a homogeneous solution is formed only at elevated temperature. By cooling down the homogeneous
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solution, the phase separation is induced. Once the polymerrich phase is solidified, the porous membrane structure can be created by removing the solvent via extraction. Membrane formation by TIPS can be illustrated by the phase diagram of a polymer solution as a function of temperature from the basic point of thermodynamics, as shown in Figure 9. From the phase diagram, it can be seen that there exists three different phase areas of homogeneous solution phase I, the liquid–liquid demixing (metastable areas II, III, and unstable area IV) and one crystalline phase V, which are separated by a binodal curve, a spinodal curve, and a crystallization curve. If a homogeneous polymer–solvent mixture at a temperature TA, as indicated by point A in Figure 9, is cooled to the point M, the solution separates spontaneously into two phases after it crosses the spinodal curve. Upon further cooling to the temperature Tgel, as indicated by point B, the composition of the polymer-rich phase reaches point B00 , and the point B’ represents the solvent-rich, the polymer-lean
liquid phase. At this moment, the polymer-rich phase starts to solidify and forms the solid membrane structure, and the polymer-lean phase forms the pores. An interpenetrating three-dimensional network can be obtained. DIPS. The invention of the first integral asymmetric membranes by DIPS was a major breakthrough in the history of RO and UF membrane development (Loeb and Sourirajan, 1964; Kesting, 1985). DIPS can be realized through immersing the casting solution in a nonsolvent coagulant bath, evaporating the solution and using vapor adsorption. Figure 10 illustrates the concepts of three DIPS processes. For an immersion precipitation process, at least three components of polymer, solvent, and nonsolvent are involved. The membrane formation can also be illustrated with a ternary phase diagram as shown in Figure 11. In the diagram, four different regions are shown: one solution phase (region I), liquid–liquid two phases (region II), liquid–solid two phases (region III), and one solid phase (region IV).
Unstable Liquid phase I
TA
Metastable
A
Critical point
Temperature
Binodal curve Spinodal curve IV
III
II
M
B
B′
Tgel
cur
ve
iz
tall
s Cry
n atio V
B″
Liquid–solid demixing
0 A solvent-rich phase
Polymer composition A polymer-rich phase
1
An interpenetrating three-dimensional network Figure 9 Schematic phase diagram of thermally induced phase separation. Modified from van de Witte P, Dijkstra PJ, van de Berg JWA, and Feijen J (1996) Phase separation process in polymer solutions in relation to membrane formation. Journal of Membrane Science 117: 1–31.
Immersion precipitation (coagulation bath) NS
Vapor adsorprtion
Solvent evaporation
NS
S S Casting solution (polymer + solvent + nonsolvent + additives) Support Figure 10 Representation of three DIPS processes (S: solvent; NS: nonsolvent). Modified from Ren JZ and Wang R (2010) Preparation of polymeric membranes. In: Wang LK, Chen JP, Hung YT, and Shammas NK (eds.) Handbook of Environmental Engineering, vol. 13, ch. 2. Totowa, NJ: Humana Press.
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P (polymer)
IV
1
Vitrification S0
2-1: Path 2
S2 1 S1 K >1 2-1
2-3
III 2-3:
K <1 2-2
A
2-2:
S1 K <1
II
I
S1
S3
N (nonsolvent)
S (solvent) Critical point Binodal
Spinodal
Gelation boundary
Figure 11 Schematic diagram of membrane formation process for DIPS. Modified from Ren JZ and Wang R (2010) Preparation of polymeric membranes. In: Wang LK, Chen JP, Hung YT, and Shammas NK (eds.) Handbook of Environmental Engineering, vol. 13, ch. 2. Totowa, NJ: Humana Press.
Bore fluid Dope inlet
Nitrogen
Spinneret
Dope fluid
Water spray
Bore fluid
Syringe pump
Coagulation bath
Flushing bath
Figure 12 Schematic diagram of a hollow fiber spinning line.
Point A represents the initial composition of a casting solution. If the polymer solution is immersed into a nonsolvent bath, there are two possible composition paths, 1 and 2 (2-1, 22, 2-3). For path 1, the polymer solution undergoes a glass transition and goes into the solid phase IV directly. Consequently, the solution becomes a homogeneous glassy film. For the path 2, gelation dominates the formation of the porous membrane morphologies. When the composition path crosses the binodal curve and reaches point S1, liquid–liquid phase separation (S1 - S2 þ S3) occurs. The polymer-rich phase is represented by point S2 and polymer lean phase is represented by point S3. Depending on the location of point S1, three different nascent membrane morphologies are formed by (a) the nucleation and growth of the polymer-poor phase (path 2-
1), which leads to a morphology with dispersed pores; (b) the spinodal decomposition (path 2-2), which leads to a bi-continuous network of the polymer-poor and polymer-rich phases without any nucleation and growth due to instantaneous demixing; and (c) the nucleation and growth of the polymerrich phase (path 2-3), which leads to low-integrity powdery agglomerates. Such membrane morphology is not practical and thus it rarely happens in membrane formation.
4.11.3.2.2 Fabrication of hollow fiber membranes Polymeric hollow membranes were first introduced in 1966 (Mahon, 1966). A schematic diagram of a hollow fiber spinning line is shown in Figure 12. Hollow fiber membranes can
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be fabricated by either wet or dry-jet wet spinning process, where an air gap between the spinneret and the coagulation bath can be zero or a certain value. Basically, the polymer dope container is connected to a N2 gas cylinder or a pump. The dope is dispensed under pressure or pumped through a spinneret at a controlled rate, and goes through an air gap before immersing into a coagulation bath. Water is (typically) used as external coagulant, while a mixture of milli-Q Water and a solvent with varying ratios is used as the bore fluid. The nascent hollow fiber is taken up by a roller at a free falling or controlled velocity and stored in a water bath to remove
residual solvent for further characterization (Strathmann, 1990; Shi et al., 2008). Figure 13 shows the morphology of poly(vinylidene fluoride-co-hexafluropropylene) (PVDF-HFP) asymmetric microporous hollow fiber membranes made by the DIPS process (Shi et al., 2008). Being an extremely complex process, the fabrication of hollow fiber membranes requires highly sophisticated mechanical, thermodynamic, and kinetic considerations. The spinning parameters involved are summarized in Figure 14. Mckelvey et al. (1997) have provided detailed guidance on how to control macroscopic properties of hollow fiber
Figure 13 Cross-section morphology of the hollow fibers spun from the PVDF-HFP/NMP dopes without an additive. Reproduced from Shi L, Wang R, Cao YM, Liang DT, and Tay JH (2008) Effect of additives on the fabrication of poly(vinylidene fluoride-co-hexafluropropylene) (PVDF-HFP) asymmetric microporous hollow fiber membranes. Journal of Membrane Science 315: 195–204, with permission from Elsevier.
Formula Bore fluid
Dope
Approaching ratio Thermodynamic
Temperature Dope solution
Shear
Bore fluid (approaching coagulation ratio) Shear rate Spinneret
Shear stress
Shear flow
Velocity distribution
Die swell
Temperature, humidity Die swell
Elongation
Relaxation Air gap
Evaporation Elongation rate
Dynamic Elongational
Elongation stress flow Temperature Approaching coagulation ratio Coagulation
Stretching Solidification Posttreatment
Figure 14 Parameters involved in a dry-jet wet spinning process for hollow fiber membranes. Modified from Ren JZ and Wang R (2010) Preparation of polymeric membranes. In: Wang LK, Chen JP, Hung YT, and Shammas NK (eds.) Handbook of Environmental Engineering, vol. 13, ch. 2. Totowa, NJ: Humana Press.
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
membranes using dominant process parameters, including spinneret design, dope extrusion rate, air gap distance, bore fluid extrusion rate, solvent concentration in the bore fluid, vitrification kinetics, etc., and how to determine the optimal macroscopic properties. The effect of shear rate on the performance and morphology of hollow fiber membranes can also be found in the literature (Qin et al., 2001).
4.11.3.3 TFC Membrane Preparation In addition to the integral asymmetric membranes, which are produced by the phase-inversion process, composite membranes are also widely used in industry. For instance, many NF and RO membranes are formed with a composite structure (see also Section 4.11.2). A typical composite membrane is shown schematically in Figure 4. Normally, composite membranes are made in a two-step process: (1) fabricating a microporous support and (2) depositing/casting a barrier layer on the surface of the microporous support layer. This approach provides great flexibility for selecting different materials to tailor the membrane structure and properties. Today, the most important technique for manufacturing composite membranes is the interfacial polymerization of reactive monomers on the surface of a microporous support membrane. This technique was developed in the mid-1970s (Cadotte, 1977). A PS microporous membrane is soaked in an aqueous solution containing 0.5–1% polyethyleneimine, and then brought in to contact with a 0.2–1% solution of toluene diisocyanate in hexane. These two reagents react rapidly on the membrane surface, forming the selective layer of the composite membrane. A heat curing step leads to further crosslinking of the polyethyleneimine, which extends into the pores of the support membrane (Strathmann, 1990). More information on TFC membrane properties is given in Section 4.11.2.2.
4.11.3.4 Ceramic Membrane Preparation Similarly to polymeric membranes, ceramic membranes have been developed for many process applications, including MF and UF in the water industry. They are more thermally and chemically stable than polymer membranes with much greater mechanical strength and higher structural stability, which allows ceramic membranes to be used in harsh environments. Generally, the structure of ceramic membranes is asymmetric, consisting of a macroporous support, one or two mesoporous intermediate layers, and a microporous (or a dense) top layer. The support layer provides mechanical strength, the middle layers bridge the pore-size differences between the support and the top layers, and the thin top layer determines the separation. The preparation of a ceramic membrane support normally involves the following steps: (1) forming particle suspensions; (2) packing the particles in the suspensions into a membrane precursor with a certain shape using various shaping techniques such as slip casting, tape casting, extrusion, and pressing; and (3) sintering the membrane precursor at elevated temperature (Li, 2007). The separation layers of composite ceramic membranes are made of SiO2, Al2O3, ZrO2, and TiO2 materials, which can be formed on a membrane support via dip-coating, sol–gel, chemical
319
vapor deposition (CVD), or electrochemical vapor deposition (EVD), followed by repeated firing steps. A detailed description of ceramic membrane preparation methods can be found in Li (2007). The phase-inversion method which is commonly employed to spin polymeric hollow fiber membranes can also be used to prepare inorganic hollow fibers in combination with sintering step. The protocol of preparation is depicted in Figure 15. Basically, a desired spinning dope (a mixture of ceramic powder and polymer solution) is formed and passed through a spinneret to enter a water bath for precipitation. After posttreatment, the hollow fiber precursors are heated in a furnace to remove the organic polymer binder, and then calcined at a high temperature to allow the fusion and bonding to occur. Figure 16 shows the morphology of Zirconia hollow fibers membranes made by this method (Liu et al., 2006). Various factors such as the sintering temperature and the ratio of the selected ceramic powders to the polymer binders will affect the structure and performance of the resultant membranes.
4.11.4 Membrane Characterization Membrane characterization is critical to the understanding of the chemistry and structure of membranes and to the identification of causes of membrane failures. This section also briefly reviews a wide range of commonly used characterization techniques, including pore-structure characterization, microscopic methods, as well as spectroscopic methods for information on membrane chemistry. Membrane characterization is the basis for establishing the chemistry–structure–properties relationship that is a critical aspect in any new membrane development process. This can guide subsequent membrane optimization to achieve excellent separation efficiency and fouling resistance. Characterizing and understanding membrane chemistry and properties is also the key for selecting suitable membranes for a given application. A wide range of characterization methods have been applied to membrane characterization. This section briefly summarizes some of the most commonly used techniques (Table 8): Performance tests. This involves measurements of membrane permeability and rejection of various solutes using a filtration setup. The performance data (e.g., rejection of probe molecules such as dextran or polyethylene glycol) can be used to determine pore-size distribution of a membrane (Aimar et al., 1990; Schafer et al., 2005). Membrane porometry. Various porometry methods have been used to determine the pore-size distribution of MF and UF membranes. The bubble point method, which measures the pressure at which air bubbles start to pass through a wetted MF membrane, can be used to characterize pores 40.15 mm. Another method commonly used for MF membrane pore characterization is the mercury intrusion method, which is suitable for pores ranging from 3.5 to 1000 nm. Other methods include gas adsorption, permporometry, and thermoporometry, which are suitable for characterizing smaller pores. A comprehensive review of these methods can be found in Nakao (1994).
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Mixing of inorganic powder, polymer binder, and solvent
Suspension
Spinneret Dope fluid
Wet spinning
Sintering Figure 15 Protocol of preparing inorganic hollow fibers using modified phase-inversion method.
Figure 16 Morphology of zirconia hollow fiber membranes: (a) cross section and (b) outer surface Reproduced from Liu LH, Gao SJ, Yu YH, Wang R, Liang DT, and Liu M (2006) Bio-ceramic hollow fiber membranes for immunoisolation and gene delivery. I: Membrane development. Journal of Membrane Science 280: 762–770, with permission from Elsevier.
Microscopic methods. Microscopic methods are useful for the visualization of membrane morphology and structure. They can be used to study the surface features of a membrane as well as its cross section. Microscopic methods are also widely used for membrane pore-size characterization (Kim et al., 1990). Where a foulant cake layer is present, microscopic investigation can provide great details about the morphology, structure, and properties of the layer. Conventional visible light microscopy is routinely used for membrane visualization due to its low cost and easy operation. However, its resolution is usually limited due to the relatively large wavelength of visible light. In contrast, electron microscopic methods, such as scanning electron microscopy (SEM) and transmission electron microscopy (TEM), offer much better resolution. SEM has been widely used to characterize the surface and cross section of both clean and fouled membranes at a resolution as good as 5 nm for polymeric samples. Where ultrathin (o100 nm in thickness) sections can be prepared, TEM can
provide a large amount of detail on the structural information of membranes and foulants (Tang et al., 2007a; Freger, et al., 2005). Other commonly used microscopic methods include atomic force microscopy (AFM) and confocal laser scanning microscopy (CLSM). AFM can provide valuable information on the surface features of a membrane, and it has become the standard method for characterizing membrane roughness. On the other hand, CLSM is a powerful method for biofilm characterization. Spectroscopic methods. Spectroscopic methods provide essential information on the structure and chemistry of a membrane. For example, Fourier transform infrared spectroscopy (FTIR) is able to identify various chemical bonds in a membrane based on its adsorption of infrared irradiation (Tang et al., 2007a). Both X-ray photoelectron spectroscopy (XPS) and energy dispersive spectroscopy (EDX) are widely used for identifying elements present in membranes. XPS is a highly surface-sensitive technique, with the ability to measure
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Table 8
321
Commonly used membrane characterization methods
Type
Instrument
Information
Performance test
Membrane filtration setup
Permeability, rejection, and pore-size distribution inferred from transport models
Membrane porometry
Bubble point, mercury intrusion, gas adsorption, permporometry, thermoporometry
Information on membrane pore structure
Microscopic methods
SEM TEM AFM CLSM
Surface/cross-section features Cross section of membrane/foulant Roughness, surface morphology Foulant structure/composition
Spectroscopic methods
FTIR XPS EDX EIS
Membrane/foulant functional groups Elements/chemical binding Elemental mapping of foulants Structural information of sublayers
Other methods
Goniometer Streaming potential AFM force measurement
Hydrophobicity Surface charge Interaction force
AFM, atomic force microscopy; CLSM, confocal laser scanning microscopy; EDX, energy dispersive spectroscopy; EIS, electrical impedance spectroscopy; FTIR, Fourier transform infrared spectroscopy; SEM, scanning electron microscopy; TEM, transmission electron microscopy; XPS, X-ray photoelectron spectroscopy.
elemental composition and chemical binding information for the top 1–5 nm depth of the surface region (Tang et al., 2007a). The technique is able to detect all elements except hydrogen with detection limits around 0.01 monolayer or 0.1% of the total elemental concentration. Compared to XPS, EDX is less sensitive. Nevertheless, EDX offers a unique advantage as it can be coupled to an SEM or TEM, which allows it to analyze microscale features at specific locations and to construct elemental mapping. EDX has been widely used in membrane autopsy to analyze chemical compositions of membrane foulants (Khedr, 2003). Recently, electrical impedance spectroscopy (EIS) has been used to characterize the layered structures of TFC-RO and NF membranes. This method is able to provide structural information (e.g., thickness and electrical properties) of each sublayer of a composite membrane (Coster et al., 1996). Other surface characterization techniques. Goniometer and streaming potential analyzer are widely used to characterize membrane surface hydrophobicity (via contact-angle measurements) and surface charge (via zeta potential measurements), respectively (Tang et al., 2009a). AFM interaction force measurement is an emergent technique for membrane surface and foulant layer characterization (Bowen et al., 1999; Lee and Elimelech, 2006; Tang et al., 2009c). Using a colloidal cantilever probe with well-defined surface chemistry and calibrated spring constant, the AFM force measurement technique can be used to measure tiny interaction forces (B1 nN) between a membrane surface (either clean or fouled) and the probe. Such interaction forces correlate well with membrane fouling behavior (Tang et al., 2009 c; Lee and Elimelech, 2006).
4.11.5 Membrane Modules The membranes described in Sections 4.11.2 and 4.11.3 are produced as hollow fibers, tubes, and flat sheets. To use these membranes in large-scale processes, it is necessary to
incorporate them into a membrane module. Important features of modules include packing density, ease of cleaning, and flow distribution. Several module geometries have been developed suited to the range of membranes and their applications in the water industry.
4.11.5.1 The Role of the Module The membrane module, or element, has two major roles: (1) supporting the membrane and (2) providing efficient fluid management. Membranes are typically produced as flat sheets, tubes, or hollow fibers. The flat sheet and tubular forms are not self-supporting and the membranes must be placed on a porous support able to withstand the applied pressure and also facilitate permeate removal. Hollow fibers can be selfsupporting, and operate outside-to-in or inside-to-out. The latter is also called lumen feed, where the lumen is the bore of the hollow fiber. Good fluid management is vital for efficient membrane processing. The hydrodynamic conditions in the boundary layer at the membrane surface control the concentration polarization (CP) (see Section 4.11.6.3), which directly influences membrane performance. The various module designs deal with feed-side flow in different ways, attempting to balance boundary-layer mass transfer and the feed channel pressure losses. Fluid management also pertains to the downstream, permeate side of the membrane, because resistance to flow determines the downstream pressure losses and the net transmembrane pressures (TMPs). Several characteristics are potentially important in module design and are summarized in Table 9. In what follows, the various modules are described (Section 4.11.5.2), and their characteristics are summarized in Table 10.
4.11.5.2 Module Types This section describes the most common modules used in the water industry. One early approach, not described, was the
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Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
plate-and-frame module which used stacks of flat sheet membranes on porous supports separated by flow channel spacers. Limitations due to packing density, pressure containment, and labor-intensive membrane replacement lead to the development of the SWM (see below).
4.11.5.2.1 Spiral-wound module The SWM is the predominant design for RO and NF applied to the water industry (see Section 4.11.1.3 for some of the historical developments); as such, it is the workhorse for SWRO and water reclamation plant. The SWM uses flat sheet membranes sealed by gluing on three sides to form leaves attached to a permeate channel (tube) along the unsealed edge of the leaf. Inside each leaf is a permeate spacer which is a porous matrix designed to support the membrane without compression, and to have a high hydraulic conductivity for permeate flow to the permeate tube (see Figure 17(a)). A net-like feed channel spacer fits between the leaves and defines the channel height (typically B1 mm). Several leaves are fixed to and then wound around the permeate tube and given an outer rigid casing (Figures 17(b) and 17(c)). The module has an anti-telescoping end cap which provides support to counter axial pressure drops. It should be emphasized that the feed channel spacer plays an important role as it enhances the effect of crossflow and promotes boundary layer mass transfer that controls CP (see Section 4.11.6.3). Table 12 (Section 4.11.6) provides information about the mass transfer correlations for different spacer geometries.
Table 9
The SWM comes in a standard diameter of 8 in (203 mm), but 2.5 and 4 in are also used for pilot or small scale, and 16 in SWMs are being introduced. The SWM is fitted into standard pressure vessels which can take several elements connected in series with O-ring seals to prevent bypassing and feed-to-permeate flow. Up to eight modules could be present in a pressure vessel, and many vessels are connected in an array (examples given in Section 4.11.7.2). A large desalination plant could have 20 000–50 000 modules.
4.11.5.2.2 Tubular module Tubular modules have the membrane surface on the inside of the tubes. They have several niche applications at the medium scale. Diameters are in the range 5–25 mm. The modules are similar to the shell and tube heat exchanger (Figure 18) with tubes connected in parallel and series. Some designs have the membrane tubes inserted into porous metal support tubes, and are able to withstand pressure for RO and NF. In other cases, the tubes are self-supporting and the burst pressure of the tubes limits it to UF/MF applications. Tubular modules are also produced in ceramic materials as multichannel monoliths with UF or MF capability; there are reported applications in water treatment. For RO and NF tubular modules are operated with crossflow in the turbulent flow regime which provides good control of CP, but at a relatively high energy cost. This type of module is suitable for feeds with high turbidity. An interesting example is remote-area water treatment which uses tubular NF with automatic foam ball cleaning for chemicalfree water treatment of colored waters (see reference 12 in Fane (2005)).
Module characteristics of importance
Characteristic Packing density Energy use
Fluid management
Standardization Replacement Cleaning
Table 10
Significant influence on
4.11.5.2.3 Hollow fiber module (contained)
System size, footprint, and (probably) cost Costs ¼ f {operating pressure, flow rate, flow resistance, flow regime} Concentration polarization, flux/ pressure relationship, fouling, and cleaning Flexibility in terms of choice of membrane supplier Maintenance and labor costs System availability, downtime, and time-averaged production
Hollow fiber membranes are self-supporting, that is, the walls can be strong enough to avoid collapse or bursting. Outer diameters are in the range of 0.5–1.0 mm with inner lumen diameters of o0.3 to 0.8 mm. Hollow fiber modules (HFMs) are either contained (filtration under pressure) or submerged (filtration under suction); Section 4.11.5.2.4 deals with submerged modules. Contained HFMs involve thousands of fibers arranged in a bundle and potted by epoxy in an outer shell (Figure 19). The design is similar to the shell and tube design for tubular membranes, but can be operated with feed in the shell side (out-to-in) or feed in the lumen (in-to-out), depending on the membranes and the application. HFMs with shell-side feed are externally pressurized and some RO hollow
Characteristics of different module concepts
Characteristic
Spiral wound
Tubular
Hollow fiber
Submerged
Packing density (m2 m3) Energy use Fluid/fouling management
High (500–1000) Moderate (spacer losses) Good (no solids) Poor (solids) Yes Element Can be difficult (solids)
Low–moderate (70–400) High (turbulent) Good
High (500–5000) Low (Laminar) Moderate (in-to-out) Poor (out-to-in) No Element Backflush (MF/UF)
Moderate Low Moderate
Standardization Replacement Cleaning
No Tubes (or element) Good – physical cleaning possible
No Element (or bundle) Backflush (HF) (MF/UF)
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Permeate spacer
Membrane leaf
Feed spacer
Permeate tube
Flux (a)
(b)
Feed flows axially Feed Feed-side spacer
Permeate
Membranes
(c)
Permeate flows inward to collector tube
Membrane support and permeate-side spacer
Figure 17 (a) Spiral-wound module showing membrane leaves and spacers. (b) Spiral-wound module with leaves wrapped around permeate tube. (c) Spiral-wound module showing flow paths. (a–c) Reproduced from Fane AG (2005) Module design and operation. In: Schaefer AI, Fane AG, and Waite TD (eds.) Nanofiltration – Principles and Applications, pp. 67–88. Oxford: Elsevier, with permission from Elsevier.
‘Shell’
Membranes (‘tubes’)
Retentate
Feed
Permeate Figure 18 Tubular module (shell-and-tube arrangement). Reproduced from Fane AG (2005) Module design and operation. In: Schaefer AI, Fane AG, and Waite TD (eds.) Nanofiltration – Principles and Applications, pp. 67–88. Oxford: Elsevier, with permission from Elsevier.
fibers can withstand high pressures up to the level of SWRO. In some special cases, seawater applications with hollow fibers of cellulose triacetate are used (Kumano and Fujiwara, 2008); their major advantage is the ability to withstand chlorine to control biofouling. However, in the vast majority of seawater RO desalination plants the HFM has been superseded by the SWM.
HFMs with shell-side feed are commonly used in lowpressure membrane applications (UF and MF), such as water treatment and pretreatment (see Table 2). These applications often use dead-end operation (see Section 4.11.7.1) with intermittent backwash from the lumen to the shell. Compared with submerged HFMs, the contained modules have a wider range of TMPs available. HFMs with lumen-side feed are also used for water treatment and pretreatment. Operation is either with crossflow or with dead-end flow, depending on the solids content (low solid favors use of dead-end with backwash). Intermittent two-phase (air–liquid) flow is often applied to HFMs during the backwash cycle. Continuous two-phase flow may be implemented in cases where the HFM (lumen-side feed) is used with high solids, such as MBRs (see Section 4.11.1.2.4).
4.11.5.2.4 Submerged module Submerged (or immersed) modules involve membranes positioned in a flooded tank at atmospheric pressure typically open at the top. The liquid to be filtered is fed to the tank and permeate is removed from the module under suction, either
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Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Hollow fibers
Feed
Retentate
Permeate
Shell
Lumen
Figure 19 Contained hollow fiber module. Reproduced from Fane AG (2005) Module design and operation. In: Schaefer AI, Fane AG, and Waite TD (eds.) Nanofiltration – Principles and Applications, pp. 67–88. Oxford: Elsevier, with permission from Elsevier.
Feed
Permeate
Hollow fibers or flat sheets
intermittent air scour. Flat sheets are not used in these applications because they cannot be backwashed. In these low solid operations both submerged and contained HFMs appear to be equally popular. In some cases, submerged HFMs may offer marginal cost advantages, but they have less turn-up/turndown capability and are heavier than contained HFMs. For high solid content feeds, such as MBRs, submerged modules are either hollow fibers or flat sheets, with continuous or rapidly intermittent air scour, and occasional backwash (hollow fibers). Submerged modules are more popular than contained modules for MBRs. There is no standardization in submerged membranes and there are many commercial suppliers; for example, Judd (2006) describes 12 different MBRs using submerged membranes. A more detailed account of submerged membranes can be found elsewhere (Fane, 2008).
4.11.6 Basic Relationships and Performance
Waste
Air
Figure 20 Submerged membrane module.
by a pump or by gravity. The concentrate is removed continuously or intermittently from the tank. Figure 20 depicts the general features of a submerged membrane system, which include: 1. an open tank (no pressure vessel), 2. modules in bundles of fibers or vertically aligned flat plates, 3. permeate removed by suction, and 4. TMPs o1 atm. Submerged modules use low-pressure MF and UF membranes; they are unsuitable for NF or RO due to the limited TMP. For low solid feeds, such as water treatment or pretreatment to RO, submerged hollow fibers are commonly used. Operation is usually dead-end cycles with regular backwash and
Membrane performance (such as flux and rejection) is determined by the mass transport inside a membrane as well as the transport toward the membrane surface. The mass transport inside a membrane defines the basic relationship between flux and the driving force. In addition, it determines the intrinsic retention properties of the membrane. Due to its retentive nature, solutes transported toward a membrane will tend to accumulate near the membrane surface, leading to a higher solute concentration near the surface compared to the bulk concentration. This phenomenon is known as CP. Another important phenomenon in pressure-driven membrane processes is membrane fouling, that is, the deposition of contaminants on a membrane surface and/or inside membrane pores. Both CP and fouling can adversely affect membrane flux and rejection. Thus, they need to be carefully controlled in membrane operation.
4.11.6.1 Membrane Flux and Rejection Flux and rejection are among the most important performance parameters for any membrane process. Membrane flux of a given species can be defined as the mass (or volume) of that species passing through a unit membrane area within a given duration. For applications in water and wastewater treatment,
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
the water flux is of particular interest, as this directly relates to the membrane productivity and thus process economics (refer to Section 4.11.7 for more details). Water flux Jw is typically defined as
ð5Þ
where Qp is the volumetric flow rate that permeates through the membrane and Am is the membrane area. Typical units for water flux in the literature include m3 m2 s1, m d1, mm s1, l m2 h1 and (US) gallons per ft2 per day (gfd). Similar to the definition of water flux, the mass flux of a solute Js can be defined as the mass flow rate m˙ s passing through the membrane normalized by membrane area:
Js ¼
m˙ s Am
ð6Þ
The solute flux is commonly given in kg m2 s1, kg m2 h1, mol m2 s1, or mol m2 h1. The retention ability of a membrane in water applications is expressed by the membrane rejection. The intrinsic rejection of a membrane Rint can be defined as (Ho and Sirkar, 1992)
Rint ¼ 1
Cp Cm
ð7Þ
where Cm is the solute concentration near the membrane surface and Cp is the solute concentration in the permeate water. Here, Cp can be determined from the ratio of the solute mass flux to the volumetric water flux by
Cp ¼
Js Jw
ð8Þ
The intrinsic rejection Rint defined in Equation (7) relates the solute concentration in the permeate water Cp to that near the membrane surface Cm. Usually, Cm is not known as a priori.
Table 11
Thus, a more commonly used rejection parameter in practice is the apparent rejection Rapp, which relates Cp to the bulk feed concentration Cb (Ho and Sirkar, 1992):
Rapp ¼ 1
Qp Jw ¼ Am
325
Cp Cb
ð9Þ
In a similar fashion, an overall observed rejection Rsys (rejection at the module or system level) can be defined based on the feedwater concentration Cf :
Rsys ¼ 1
Cp Cf
ð10Þ
The overall rejection is usually lower than the intrinsic rejection. This can arise due to two main reasons: (1) concentration polarization which leads to a higher concentration near the membrane surface compared to the average bulk concentration (Cm Z Cb), and/or (2) high membrane recovery so that the average bulk concentration experienced by a membrane is greater than the feedwater concentration (Cb ZCf). Thus, Rint Z Rapp Z Rsys. The concentration polarization phenomenon will be discussed in greater detail in Section 4.11.6.3, and the effect of membrane recovery is discussed in Section 4.11.6.4. It is also worth noting that the retention mechanism for porous membranes (MF and UF) is different from that for nonporous membranes (RO) (Table 11). Particles and solutes are retained by porous MF or UF membranes by size discrimination, that is, a sieving mechanism (Mulder, 1996). Particles larger than membrane pore size are completely retained, while smaller particles are less retained. In contrast, selectivity of an RO membrane is based the solution-diffusion mechanism (Mulder, 1996). Solute or solvent absorbs into the nonporous membrane on the feedwater side, diffuses through the rejection layer under a chemical potential gradient, and desorbs on the permeate water side. Separation of different species is achieved based on their different ability to partition into the rejection layer as well as their different ability to
Rejection mechanisms for porous and nonporous membranes
Membrane type
Rejection layer
Rejection mechanism(s)
Water flux and solute rejection model(s)
MF
Porous
Sieving
Water flux: Hagen–Poiseuille equation; Kozeny–Carman equation Rejection: Ferry equation; Zeman and Wales equation; other pore models (Nakao, 1994)
UF
Porous
Sieving
Water flux: Hagen–Poiseuille equation; Kozeny–Carman equation Rejection: Ferry equation; Zeman and Wales equation; surface forcepore flow model; hindered transport models
NF
In between tight UF and loose RO
Solution-diffusion, sieving, Donnan exclusion
Solution-diffusion model; solution-diffusion-imperfection; preferential sorption-capillary flow model; surface force-pore flow model; Donnan equilibrium model; extended Nernst–Planck model
RO
Nonporous
Solution-diffusion
Solution-diffusion model; solution-diffusion-imperfection; preferential sorption-capillary flow model; surface force-pore flow model
MF, microfiltration; NF, nanofiltration; RO, reverse osmosis; UF, ultrafiltration. Adapted from Schafer AI, Fane AG, and Waite TD (2005) Nanofiltration – Principles and Applications. Oxford: Elsevier; Ho WS and Sirkar KK (1992) Membrane Handbook. New York: Chapman and Hall; and Nakao S (1994) Determination of pore size and pore size distribution. 3. Filtration membranes. Journal of Membrane Science 96: 131–165.
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diffuse through the rejection layer. RO membranes generally allow relatively high absorption of water molecules and faster diffusion of water molecules through the rejection layers (Mulder, 1996). In comparison, a typical solute (such as sodium chloride) has lower sorption onto RO membranes and slower diffusion through them, which results in a lower solute concentration in the permeate compared to that of the feedwater. As NF membranes are in between tight UF membranes and loose RO membranes, transport inside NF membranes may be governed by both solution-diffusion and sieving mechanisms (Schafer et al., 2005). In addition, charge repulsion (Donnan exclusion) can be important for the rejection of charged species (Schafer et al., 2005). Both porous membrane models and the solution-diffusion model are discussed in more detail in Section 4.11.6.2.
4.11.6.2 Transport Inside a Membrane – Basic Relationships
ð11Þ
where Lp is the water permeability coefficient of the pressuredriven membrane. In Equation (11), Dp represents the osmotic pressure difference between the membrane surface pm and permeate water pp. The osmotic pressure of dilute solutions can be determined by the van’t Hoff equation:
p ¼ Rg T S Ci
•
applying a positive Pm while maintaining Pp around atmospheric pressure (Pp B 0); this is typically done for most RO and NF applications, as well as for many MF and UF applications and applying a negative Pp (suction or partial vacuum) while maintaining an atmospheric Pm, which is widely used for submerged membranes (see Section 4.11.5.2.4).
The Darcy’s law for pressure-driven membranes is also commonly presented in terms of membrane hydraulic resistance Rm and dynamic viscosity of the permeating water Z by
Jw ¼
DP Dp ZRm
1 ZLp
ð14Þ
The Darcy’s law (Equation (11) or Equation (13)) is applicable for both porous and nonporous membranes. A more sophisticated model available in the membrane literature is the irreversible thermodynamics model (Mulder, 1996; Bitter, 1991), which recognizes that membrane processes are not under thermodynamic equilibrium due to the continuous free energy dissipation and entropy production. According to the irreversible thermodynamics model, the volumetric flux Jv and the solute flux Js are given by the following equations, respectively (Mulder, 1996; Bitter, 1991):
Jv ¼ Lp ðDP sDpÞ
ð15Þ
sÞJv þ Ls DC Js ¼ Cð1
ð16Þ
where Lp is the water permeability, s the reflection coefficient, the average solute concentration inside the membrane, Ls C the solute permeability coefficient, and DC the solute concentration across the membrane (DC ¼ Cm – Cp). Equation (15) takes a similar form to that of Equation (11), except a reflection coefficient s is introduced in the former. For s ¼ 1, two equations become identical. In effect, s is an indicator of a membrane’s ability to separate a solute from the solvent, and its value is usually between 0 and 1:
•
ð12Þ
where Rg is the universal gas constant (R ¼ 8.31 J mol1 K1), T the absolute temperature in kelvin, and Ci the molar concentration of dissolved species i. Thus, for a 0.01 M NaCl, SCi ¼ 0.02 M as there are 0.01 M sodium ions (Naþ) and 0.01 M chloride ions (Cl). The osmotic pressure term appears in Equation (11) only if the solute under concern is retained by the membrane. Similar to the osmotic pressure difference, DP in Equation (11) is the TMP, that is, the difference between the pressure near membrane surface Pm (which is identical to the applied pressure on the feedwater side) and that in the permeate water Pp. A positive TMP can be achieved by:
•
Rm ¼
and
The permeate water flux of a membrane can be related to its driving force (i.e., the net pressure difference across the membrane) following the phenomenological Darcy’s law:
Jw ¼ Lp ðDP DpÞ
where the membrane hydraulic resistance Rm is related to its water permeability by
ð13Þ
•
•
For s ¼ 0, the membrane has no selectivity with respect to the solute. One example is the rejection of dissolved salts by porous membranes. As MF and (most) UF membranes do not retain dissolved salts, no osmotic pressure difference will be developed across these membranes (i.e., Dp ¼ 0 and ¼ Cp ¼ Cm). This leads to Jv ¼ LpDP and Js ¼ Cm Jv. As a C result of the complete leakage of solute, the volumetric flux only depends on the hydraulic pressure difference and the solute flux arises solely from convective transport. For s ¼ 1, the membrane has ideal separation properties so that the solute and the solvent transport through the membrane are independent and uncoupled to each other. As a result, the convective transport term (the first term in Equation (16)) is zero, and solute transport through the membrane is purely by diffusion. This leads to Jv ¼ Lp(DP Dp) and Js ¼ LsDC. A special example of this is rejection by high-retention RO membranes (Section 4.11.6.2.2). A real membrane typically has a s between 0 and 1, which indicates that the solute transport is partially coupled to the solvent transport (Bitter, 1991).
The Darcy’s law and the irreversible thermodynamic model treat a membrane as a black box. The effect of membrane structure and properties on the transport parameters (e.g., Lp and Ls) is not reflected in these models. For this reason, mechanistic models are preferred. Section 4.11.6.2.1 discusses transport models for porous MF and UF membranes, whereas Section 4.11.6.2.2 briefly reviews the solution-diffusion model commonly applied to RO membranes.
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis 4.11.6.2.1 Transport models for MF and UF membranes The water flux of MF and UF membranes can be described by the Hagen–Poiseuille equation for cylindrical-pore membranes or the Kozeny–Carman equation for stacked-sphere pore structure (refer to Section 4.11.2.3):
Jw
er2p DP 8Ztlm
ðHagen2Poiseuille equationÞ
327
potential gradient. According to the solution-diffusion model, the water flux through an RO membrane is proportional to the net applied pressure (DP Dp), whereas the solute flux is proportional to the concentration difference across the membrane (DC):
ð17Þ
Jw ¼ AðDP DpÞ
ð21Þ
Js ¼ BDC
ð22Þ
and
or
e 3 DP Jw ¼ Kð1 eÞ 2 S 2 Zlm ðKozeny2Carman equationÞ
ð18Þ
The Hagen–Poiseuille equation and the Kozeny–Carman equation state that the water flux of a porous membrane is proportional to the applied pressure difference. The proportionality constant (i.e., the water permeability Lp) is a function of membrane pore structure (porosity, pore size, type of pores, etc.), the thickness of the rejection layer, and the viscosity of the permeating solution. The osmotic pressure difference Dp does not appear in these models because MF and UF membranes do not retain dissolved salts so that their reflection coefficient s is zero (refer to the irreversible thermodynamics model and Equation (15)). For some special cases where s is not zero (e.g., osmotic pressure due to macromolecules that can be retained by UF membranes), the osmotic pressure difference term may need to be considered as well. Rejection of solutes (or particles) by porous membranes is based on the sieving mechanism. A simple rejection equation based on Poiseuille flow for cylindrical-pore membranes was derived by Ferry (1936):
Rint ¼ ½lð2 lÞ2
ðfor lo 1Þ
ð19Þ
and
Rint ¼ 1
ðfor l 1Þ
ð20Þ
where l is the ratio of solute (or particle) diameter to the pore diameter. Ferry’s equation clearly suggests that rejection increases as the size of particle increases relative to the pore size. Strictly speaking, Ferry’s equation is applicable only for solid spherical particles in cylindrical pores. In addition, the interaction between particles in the pores and that between a particle and the pore wall are not considered. More sophisticated models are available in the literature (Ho and Sirkar, 1992; Nakao, 1994). Solute–solute and solute–pore interactions as well as membrane pore-size distribution are considered in some models (e.g., the surface force-pore flow model (Ho and Sirkar, 1992)).
4.11.6.2.2 Transport models for RO membranes One of the most widely used transport models for RO membranes is the solution-diffusion model. This model assumes that (1) both the solvent and the solute absorb into the rejection layer and (2) they diffuse through the nonporous layer independent of each other under their respective chemical
where A and B are the respective water and solute permeability coefficients in the solution-diffusion model. Comparing the solution-diffusion model and the irreversible thermodynamics model shows that the two models take the same form if the reflection coefficient s is set to unity in Equations (15) and (16). The advantage of the solution-diffusion model is that the transport coefficients (A and B) in this model can be linked to membrane properties:
A¼
Dwm Cwm Vw Rg Tlm
ð23Þ
Dsm Ksm lm
ð24Þ
and
B¼
where Dwm and Dsm are the diffusion coefficient of water and that of solute inside the rejection layer, respectively; Cwm the concentration of water inside the rejection layer; Vw the molar volume of water; and Ksm the solute partitioning coefficient into the rejection layer. The solution-diffusion model suggests that a high-flux RO membrane shall have higher water absorption and also allow fast diffusion of water molecules (Equation (23)), which requires a lower degree of crosslinking of the rejection layer. However, reduced crosslinking will lead to a significantly enhanced diffusion of solutes and thus a much greater B value. This explains the strong trade-off relationship between water permeability and salt permeability for RO membranes, as discussed in Section 4.11.2.2. The intrinsic rejection of an RO membrane can be determined by
Rint ¼
1þ
B AðDP DpÞ
1 ð25Þ
As both A and B are inversely proportional to the rejection layer thickness lm, the solution-diffusion model suggests that increasing rejection layer thickness alone does not improve membrane rejection. The intrinsic rejection of an RO membrane can be improved by (1) preferential sorption of water molecules compared to solute molecules, (2) enhanced diffusion of water molecules through the rejection layer relative to solute molecules, and (3) increased applied pressure. The solution-diffusion model can be extended to include pore flows due to membrane imperfections (the solution-diffusion-imperfection model). Other models, such as the
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Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
preferential sorption-capillary flow model and the surface force-pore flow model, assume that the rejection layers of RO membranes are microporous. For NF membranes where electrostatic interaction is an important consideration, the Donnan equilibrium model and the extended Nernst–Planck model have also been applied. A review of these models is available (Ho and Sirkar, 1992; Schafer et al., 2005).
4.11.6.3 Transport toward a Membrane – Concentration Polarization The solute concentration near a pressure-driven membrane surface is typically higher than the bulk concentration as a result of rejection by the membrane. The concentration gradient adjacent to the membrane surface leads to a diffusion of solute molecules back to the bulk solution. When the back diffusion balances with convective transport of solutes toward the membrane, a steady concentration polarization profile is established (Figure 21). Based on the mass balance of the solute in the control volume shown in Figure 21, the following equation can be established for describing the solute concentration C as a function of distance x for a one-dimensional problem:
Jw C Jw Cp D
dC ¼0 dx
ð26Þ
with the boundary conditions given by
C ¼ Cb
at
x¼0
ð27Þ
C ¼ Cm
at
x¼d
ð28Þ
and
Solving Equations (26)–(28) leads to
Cm Cp ¼ expðJw =KÞ Cb Cp
ð29Þ
where K is the mass transfer coefficient (K ¼ D/d), and exp (Jw/K) is the concentration polarization modulus. Equation (29) is the boundary layer film model. By substituting Equation (7) into Equation (29), we have
Cm expðJw =KÞ ¼ Cb Rint þ ð1 Rint ÞexpðJw =KÞ
ð30Þ
For the special case where Cp is negligible (i.e., Rint B1), Equation (30) becomes
Cm ¼ expðJw =KÞ Cb
ð31Þ
Equation (29) clearly shows that CP increases at higher water flux and reduced mass transfer coefficient. Thus, the membrane surface concentration Cm can be significantly higher than the bulk concentration at high flux and/or low mass transfer coefficient. The mass transfer coefficient can be determined from the Sherwood number Sh by relating Sh to Reynolds number Re and Schmidt number Sc (Table 12):
Sh ¼ a Re b Sc c ðdh =LÞd
ð32Þ
In Equation (32),
Sh ¼
Kdh D
ð33Þ
Re ¼
dh u v
ð34Þ
v D
ð35Þ
Jw
Sc ¼ Cm
where dh is the hydraulic diameter, u the flow velocity, and v the kinetic viscosity. Equation (32) can be rearranged to give the following form (assuming c ¼ 1/3):
JwC
D
dC dx
JwCp
Cb
Boundary layer thickness
Membrane
X
Cp
Figure 21 Concentration polarization over a membrane surface.
K ¼ a D 2=3 u b v 1=3b d bþd1 Ld h
ð36Þ
According to Equation (36), the mass transfer coefficient is proportional to D2/3. This suggests that bigger molecules are more likely to suffer from severe concentration polarization as a result of their lower diffusion coefficient. Another important point is that the mass transfer coefficient can be enhanced at larger flow velocity in a crossflow module. However, this is usually at the expense of increased pressure drop across a membrane module (pressure difference between module inlet and outlet) (Ho and Sirkar, 1992; Schafer et al., 2005; Schock and Miquel, 1987). In addition, large crossflow may damage the membrane surface (such as formation of wrinkle structures). Typical crossflow velocities for SWMs are 10–90 cm s1,
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Table 12
329
Mass transfer correlations
Geometry
Flow region
Correlation
Notes
Channel or tube
Laminar
Sh ¼ 1:62Re 0:33 Sc 0:33 ðd h =LÞ0:33
Turbulent
Sh ¼ 0:644Re 0:5 Sc 0:33 ðd h =LÞ0:5 Sh ¼ 0:023Re 0:8 Sc 0:33 Sh ¼ 0:023Re 0:875 Sc 0:25
Fully developed flow (L40.029dh), 100oRe Sc dh/Lo5000 Developing flow (Lr0.029dh) Scr1 1rScr1000
Stirred cell
Laminar Turbulent
Sh ¼ 0:285Re 0:55 Sc 0:33 Sh ¼ 0:044Re 0:75 Sc 0:33
8000rRe r 32 000 32 000rRe r 82 000
Spacers filled channels
Laminar
Sh ¼ 0:644Re 0:5 Sc 0:33 ðd h =LÞ0:5 Sh ¼ 0:644k dc Re 0:5 Sc 0:33 ð2d h =LÞ0:5
Turbulent
Sh ¼ 0:065Re 0:875 Sc 0:25
Ladder type spacer (Da Costa et al., 1994) Diamond-type spacer, correction factor kdc is a function of spacer geometry (Da Costa et al., 1994) Schock and Miquel (1987)
Adapted from Ho WS and Sirkar KK (1992) Membrane Handbook. New York: Chapman and Hall; and Fane AG (2005) Module design and operation. In: Schaefer AI, Fane AG, and Waite TD (eds.) Nanofiltration – Principles and Applications, pp. 67–88. Oxford: Elsevier.
while much higher crossflow velocities may be used for tubular modules due to their large tube diameter.
4.11.6.4 Factors Affecting Membrane Performance The transport toward a membrane surface is discussed in Section 4.11.6.3 and that inside a membrane has been discussed in Section 4.11.6.2. By combining these two aspects together, the water flux and the apparent rejection of a membrane can be determined by
Jw ¼
Rapp ¼
DP expðJw =KÞðpb pp Þ ZRm
ð37Þ
Rint Rint þ ð1 Rint Þ expðJw =KÞ
ð38Þ
Clearly, concentration polarization has a negative effect on both water flux and apparent rejection of a membrane; thus, Jw and Rapp can be significantly reduced at higher CP modulus, exp(Jw/K). This effect is more severe for larger molecules at lower crossflow and higher flux (or higher applied pressure). For systems or modules operated at low recovery (say recovery Yo0.1), the average bulk concentration Cb in the membrane system can be approximated by the feedwater concentration Cf . However, Cb can be significantly larger than Cf for systems with high recovery. The reject water (e.g., brine in RO) concentration Cc can be determined by
Cc ¼ Cf ð1 YÞRapp
ð39Þ
while the average bulk concentration in the system is approximated by
Z Cb D
Cf
y
ð1 YÞRapp dY
0
Y
¼ Cf
1 ð1 YÞ 1Rapp Yð1 Rapp Þ
ðfor 0r Rapp o 1 and 0o Y r 1Þ
ð40Þ
Rsys D 1
1 ð1 YÞ 1Rapp Y
ð41Þ
For typical applications, the feedwater concentration is given. Both the reject stream concentration and the average bulk concentration increase at higher recovery. Figure 22 shows the increase in rejection (brine) concentration and the correspondence osmotic pressure as a function of recovery for different feed concentration. For a feedwater containing 35 000 ppm NaCl (typical seawater conditions), the osmotic pressure of the brine is about 5.6 MPa at 50% recovery. Even for a feedwater with only moderate salt concentration (1000 ppm NaCl, typical wastewater reclamation conditions), this osmotic pressure can be substantial (B0.8 MPa at a recovery of 90%). As higher recovery increases the average bulk concentration and the corresponding osmotic pressure, this leads to reduced average flux as well as reduced system rejection (Equation (41)). For this reason, recovery for typical seawater RO desalination plants is limited to 50%, and that for wastewater reclamation plants is below 80%. The effect of operating conditions (applied pressure, crossflow velocity, recovery, and temperature) on membrane performance is summarized below (Figure 23):
•
Applied pressure. At low applied pressure (thus low flux level), concentration polarization is not significant. As applied pressure increases, water flux increases linearly initially according to the Darcy’s law (Equation (37)). For a porous MF membrane, the solute rejection remains constant based on the Ferry model (Equation (19)). In contrast, the solute rejection also increases for RO based on the solution-diffusion model (Equation (25)). However, at high applied pressure and water flux, concentration polarization becomes important. Further increase in applied pressure will lead to a significant concentration polarization and thus significant increase in osmotic pressure. Such an increase in osmotic pressure can offset the increase in applied pressure, resulting in a significant deviation from the linear flux–pressure relationship. The increased membrane surface concentration will also result in lower apparent membrane rejection.
330
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis 10 000 Osmotic pressure
1000 ppm feed
Brine concentration (g l–1)
5000 ppm feed 1000
35 000 ppm feed
85 MPa
100
8.5 MPa 5.6 MPa
10
1.4 MPa 0.85 MPa
1 0
100
50 Recovery (%)
Figure 22 Brine concentration and osmotic pressure as a function of recovery for a high-retention RO membrane.
Water flux
Solute rejection (for MF) Δ
Water flux Performance
Performance
Solute rejection (for RO)
Applied pressure
Solute rejection
Cross-flow
Water flux Solute rejection
Recovery
Performance
Performance
Water flux
Solute rejection
Temperature
Figure 23 Effect of operating conditions on RO membrane performance. Adapted from Ho WS and Sirkar KK (1992) Membrane Handbook. New York: Chapman and Hall.
•
• •
Crossflow velocity. Increasing crossflow tends to improve both water flux and apparent rejection of a membrane as a result of reduced CP. A plateau is usually observed at high crossflow where further increase in crossflow velocity is less effective (mass transfer is no longer a limiting factor). Recovery. Increasing recovery leads to an increase in average bulk concentration. This reduces both water flux (due to increased osmotic pressure) and system rejection. Temperature. Higher operating temperature tends to increase both water flux and solute flux due to improved diffusion through the membrane rejection layer. However, the increase in solute flux is usually more drastic compared to the enhancement in water flux. Consequently, membrane rejection tends to decrease.
Besides the operating conditions mentioned above, membrane fouling can also have profound effect of the performance of a membrane. Fouling is discussed in more detail in Section 4.11.6.5.
4.11.6.5 Membrane Fouling Membrane fouling is the deposition of contaminants on a membrane surface or inside membrane pores (Figure 24). According to the nature of the foulants, fouling can be classified into scaling (precipitation of insoluble salts), colloidal fouling, organic fouling, and biofouling (formation of a biofilm). Fouling leads to an additional hydraulic resistance (foulant resistance Rf) and therefore a lower water
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis
331
Pressurized vessel Crossflow velocity (sweeping, limiting boundary-layer thickness and CP control cake-layer formation)
Feedwater
Insoluble salts
Concentrate
Microorganisms
Organic macromolecules
Inorganic colloids
Membrane Permeate flux Figure 24 Illustration of membrane fouling in a pressurized crossflow module.
permeability of the fouled membrane:
DP Dp ZðRm þ Rf Þ
ð42Þ
The net effect of fouling is either reduced water flux at constant applied pressure or increased TMP to maintain a constant water flux. In either way, the energy demand to treat a unit volume of water can be increased significantly. Both CP and fouling can reduce water flux during constant pressure operation. CP happens within the boundary layer near the membrane surface, and it is fully reversible. Once water flux is reduced to a low level, CP disappears. The timescale for CP to reach a stable condition or to disappear is usually very short (in seconds to a fraction of a minute (Chong et al., 2007)). In contrast, foulants attach onto a membrane during membrane fouling. Membrane fouling typically occurs over longer timescales (hours to days or months), although rapid fouling can happen under some unfavorable conditions. Although CP and fouling are two different phenomena, they are closely related to each other. Severe CP can accelerate membrane fouling as a higher foulant concentration is experienced by the membrane surface. On the other hand, the formation of a cake layer on membrane surface can potentially reduce the mass transfer coefficient which results in a severe cake enhanced concentration polarization (Chong et al., 2007). Membrane fouling can be affected by many different factors, such as feedwater characteristics, membrane properties and module/system design, and hydrodynamic conditions over a membrane surface. In general, hydrophilic membranes with smooth surfaces have lower tendencies for fouling. A good module and system design improve mass transfer over the membrane surface (such as the use of spacer in SWMs and aeration in submerged membrane bioreactors (see Section 4.11.5)). Membrane fouling can be strongly affected by feedwater solution chemistry such as pH and ionic composition (Tang et al., 2007b). Unfavorable solution conditions (such as high ionic strength and hardness) can lead to severe colloidal and organic fouling by making membrane–foulant and foulant–foulant interactions less repulsive. Feedwater
Strong form Jcrit Flux
Jw ¼
Pure water flux
Weaker form Jcrit
Transmembrane pressure Figure 25 Critical flux in membrane operation. Modified from Bacchin P, Aimar P, and Field RW (2006) Critical and sustainable fluxes: Theory, experiments and applications. Journal of Membrane Science 281: 42–69.
contains high levels of sparingly soluble salts which are more susceptible to scaling formation, while the presence of microorganism and nutrients may promote biofouling. Pretreatment (such as removal of certain contaminants and pH adjustment) can be used to condition the feedwater for minimizing its fouling potential. Finally, hydrodynamic conditions are important for membrane fouling. Increased crossflow, thus enhanced mass transfer, helps minimize membrane fouling as well as CP. On the other hand, high membrane permeate flux tends to promote both CP and fouling problems. The concept of critical flux has been widely used in the membrane fouling literature. The critical flux concept states that membrane fouling is minimal below a threshold flux value (the critical flux Jcrit). Above the critical flux, significant fouling occurs. The theoretical basis has been extensively discussed in a review paper by Bacchin et al. (2006). In essence, the critical flux is the minimum flux needed to overcome the surface force and back diffusion of foulants such that fouling occurs. The critical flux can be classified into the strong form and weak form (Figure 25). In the strong form, the experimental flux versus TMP curve for a feedwater is compared to
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its pure water flux curve, and the critical flux is the flux at which the experimental flux starts to deviate from the pure water flux. In its weak form, it is recognized that rapid surface or pore adsorption of macromolecules may occur which reduces the membrane permeability. Thus, the experimental flux in these cases is always below the pure water line. Nevertheless, the experimental flux is still linear with respect to the TMP over a wide range. The weaker form of critical flux is the flux at which the experimental flux starts to deviate from its linear trend (Figure 25). Subcritical flux operation is usually preferred to avoid successive membrane fouling. Membrane critical flux can be increased by increasing mass transfer rate (such as increasing crossflow velocity, bubbling, and vibration) and avoiding unfavorable solution conditions.
(42), neglecting osmotic pressure):
dDP dRf ¼ ZJi dt dt
ð43Þ
After a specified period, tc, or at a predetermined maximum pressure drop, the flux is stopped and the deposit is removed by backwashing and (usually) vigorous aeration. The cycle times would typically be about 30 min and the backwash o5 min. This mode of operation is batch-continuous, and net flux would be slightly less than the imposed flux, that is, for tc of 30 min and for tBW of 5 min, the net flux is about 80% of imposed flux (allowing for loss of product in backwash water). Over time, due to fouling, a residual resistance may build up. This can usually be controlled by a cleaning cycle, such as chemically enhanced backwash.
4.11.7 Membrane Process Operation 4.11.7.2 System Components Membrane process operation requires consideration of whether the feed is delivered in crossflow or dead-end mode, and this depends on the nature and concentration of the contaminants to be removed. While the membranes (Sections 4.11.2 and 4.11.3) and the modules (Section 4.11.5) are the key components, the overall system includes pre- and posttreatment processes and various options for the arrangement of modules. The energy demand in most cases is directly related to the required input pressure and the fractional recovery (product/feed). The potential for energy recovery is significant in the high-pressure SWRO process. In the MBR, energy for air scour is important. The cost of water production using membranes has steadily fallen, and, in some cases, is equivalent to conventional processes.
Membrane process systems comprise membranes and modules as key components. Important additional components are the intake systems, the pretreatment steps, the feed pumps, the posttreatment steps, the energy recovery devices, and concentrate disposal method. Figure 2 is a simplified flow sheet of membrane process configurations in the water industry and Table 2 summarizes the pre- and posttreatment steps involved. Other ancillary components could be chemical addition to control fouling, membrane cleaning systems, and
4.11.7.1 Crossflow versus Dead-End Operation In many membrane applications in the water industry, such as SWRO, RO reclamation, NF water treatment, and MBRs, the aim is to operate at a steady-state production rate with continuous crossflow for controlling concentration polarization (Section 4.11.6.3) and fouling (Section 4.11.6.5). In these applications, an important consideration is how the boundary layer is influenced by crossflow velocity which depends on the flow rates and the design of the module. Any membrane application where the feed fluid is caused to move tangentially to the membrane surface is in crossflow mode. For example, MBRs with submerged membranes are operated in crossflow mode, due to the effect of continuous air scouring that creates a two-phase flow across the membrane surface. However, in the water industry, some applications are not operated in the crossflow mode. These include water treatment and pretreatment prior to RO. These processes use low-pressure membranes (MF and UF) and the feed streams have relatively low levels of suspended solids or turbidity. For these feeds, it is feasible to operate without continuous crossflow or surface shear, and this can reduce energy costs. This mode of operation is called dead-end filtration (or frontal filtration), and the key feature is that the deposition of retained species is allowed to grow. A typical cycle commences with a clean membrane (after backwash) and at constant imposed flux (Ji) the TMP rises according to Equation (43) (from Equation
(a)
(b)
(c)
Stage 1
1st pass
Stage 2
Stage 3
2nd pass
Figure 26 (a) Parallel connection. (b) Tapered cascade 3:2:1 array. (c) Two pass connection.
Membrane Technology for Water: Microfiltration, Ultrafiltration, Nanofiltration, and Reverse Osmosis Table 13
333
Approximate energy demand and costs for membrane applications to water industry
Membrane application
Energy demand (kWh m3)
Production cost (USD m3)
Reference
Seawater RO RO reclamation MBR Water treatment
3.2–3.8 1.0–1.5 o0.8 o0.3
0.5–0.75 B0.3 Similar to conventional Similar to conventional
Voutchkov and Semiat (2008) Cote et al. (2008) Cornel and Krause (2008)
MBR, membrane bioreactor; RO, reverse osmosis.
integrity testing facilities. Integrity tests are important in water treatment to check if damaged fibers are present. One popular test is the air pressure decay test (a damaged fiber shows rapid pressure loss) (Kennedy et al., 2008). Membrane modules can be connected in various ways, in series and parallel, as depicted in Figure 26. The low-pressure applications tend to have modules connected in parallel with permeate taken to a common header (Figure 26(a)). The feed is similar to all modules, although in a large system the feed may be staged. In a typical SWRO plant, the modules are arranged in both series and parallel (Figure 26(b)). The firststage pressure vessels are connected in parallel, the number of paths depending on the maximum allowable flow per module. Within the pressure vessel there would be 6–8 SWMs connected in series, such that toward the vessel outlet, the concentration builds up and the flow drops due to permeate removal. The process is continued in the second and possibly third-stage pressure vessels as shown in Figure 26(b). In this example, the second and third stages have fewer vessels in parallel as the net volumetric flow has dropped; this is known as a tapered cascade and the example is a 3:2:1 cascade. The permeate from the stages is often blended. Feed pumps may be augmented by interstage pumps to maintain pressuredriving force along the cascade. Another option is depicted in Figure 26(c), which shows a two-pass arrangement with permeate from the first set of membranes having further treatment in a second set of membranes. This approach is used if there is a need for greater removals of specific contaminants, such as boron.
1.56 kWh m3 reported (Truby, 2008). It should be noted that these values are for the RO stage only and it is usual to report plant data including seawater intake pumps, pretreatment, and other miscellaneous plant energy use. These add 0.6– 1.0 kWh m3 to energy demand (Voutchkov and Semiat, 2008). Table 13 summarizes typical energy data. The energy demand for RO reclamation is significantly less than SWRO, due to the much lower pressures required (about one-fourth of SWRO). Based on differences in O and M costs (Cote et al., 2008), the energy demand would be less than 50% of SWRO. For the MBR, a major energy demand is air scour to control fouling. Typical energy demand for MBR processing municipal wastewater is 0.75–1.0 kWh m3 (Cornel and Krause, 2008) and developments promise lower energy usage. Finally, treatment of surface water by membranes, using dead-end with backwash, has a modest energy demand of typically o0.3 kWh m3. Production costs follow similar trends to the energy demands. Table 13 gives indicative costs. It should be noted that the range could be considerable and depends on scale of operation (small plant typically have more costly product). SWRO is most costly, but the cost of production is only o0.1 cents US per liter. RO reclamation delivers water at about 50% of the cost of SWRO. For the low-pressure processes, it is now evident that both the MBR and membrane water treatment have similar costs to conventional processes for green field sites. The marginally higher energy costs are offset by the smaller foot print and infrastructure costs.
4.11.7.3 Energy and Economic Issues
4.11.8 Conclusions
Energy demand and production costs are important parameters for the water industry. As can be anticipated the greater the required pressure or the more fouling the feed, the greater the energy demand and cost. This means that energy and cost ranking is in the order, SWRO 4 RO reclamation 4 MBR 4 water treatment. A guide to the intrinsic energy demand can be obtained by noting for a flow of Q (m3 s1), with feed pressure P (Pa or N m) and a recovery Y (volume product/volume feed) the energy demand is QP/QY ¼ (P/Y) (1/3.6 106) (correcting W s m3 to kWh m3). For SWRO operating at a feed pressure of 70 bar and recovery of 0.5, the intrinsic energy usage can be estimated as 3.9 kWh m3. However, modern RO plants use pressure energy recovery on the brine stream which could return about 1.7 kWh m3, giving a net energy of about 2.2 kWh m3. Even lower values have been achieved for SWRO under carefully optimized conditions, with a value of
Membrane technology is playing an increasingly significant role in the water industry. Membranes are applied across the spectrum from seawater desalination, through wastewater treatment and reclamation, to surface water treatment. The technology continues to advance with improved membranes and processes. The energy demands and production costs have steadily declined and, in some cases, are similar to conventional processes, but with better-quality water products.
References Aimar P, Meireles M, and Sanchez V (1990) A contribution to the translation of retention curves into pore size distributions for sieving membranes. Journal of Membrane Science 54: 321--338. Bacchin P, Aimar P, and Field RW (2006) Critical and sustainable fluxes: Theory, experiments and applications. Journal of Membrane Science 281: 42--69.
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Birkett J and Truby R (2007) A figure of merit for appreciating improvements in RO membrane performance. International Desalination Association News 16(1–2): 2--3. Bitter JGA (1991) Transport Mechanisms in Membrane Separation Processes. New York: Plenum. Bowen WR, Hilal N, Lovitt RW, and Wright CJ (1999) Characterization of membrane surfaces: Direct measurement of biological adhesion using an atomic force microscope. Journal of Membrane Science 154: 205--212. Cadotte JE (1977) Reverse Osmosis Membranes. US Pat. 4,039,440, 2 August 1977. Childress AE and Elimelech M (1996) Effect of solution chemistry on the surface charge of polymeric reverse osmosis and nanofiltration membranes. Journal of Membrane Science 119: 253--268. Childress AE and Elimelech M (2000) Relating nanofiltration membrane performance to membrane charge (electrokinetic) characteristics. Environmental Science and Technology 34: 3710--3716. Chong TH, Wong FS, and Fane AG (2007) Fouling in reverse osmosis: Detection by non-invasive techniques. Desalination 204: 148--154. Cornel P and Krause S (2008) Membrane bioreactors for wastewater treatment. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 217--238. Hoboken, NJ: Wiley. Coster HGL, Chilcott TC, and Coster ACF (1996) Impedance spectroscopy of interfaces, membranes and ultrastructures. Bioelectrochemistry and Bioenergetics 40: 79--98. Cote P, Liu M, and Siverns S (2008) Water reclamation and desalination by membranes. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 171--188. Hoboken, NJ: Wiley. Da Costa AR, Fane AG, and Wiley DE (1994) Spacer characterization and pressure drop modelling in spacer-filled channels for ultrafiltration. Journal of Membrane Science 87: 79--98. Fane AG (2005) Module design and operation. In: Schaefer AI, Fane AG, and Waite TD (eds.) Nanofiltration-Principles and Applications, pp. 67--88. Oxford: Elsevier. Fane AG (2008) Submerged membranes. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 239--270. Hoboken, NJ: Wiley. Ferry JD (1936) Statistical evaluation of sieve constants in ultrafiltration. Journal of General Physiology 20: 95--104. Freger V, Bottino A, Capannelli G, Perry M, Gitis V, and Belfer S (2005) Characterization of novel acid-stable NF membranes before and after exposure to acid using ATR-FTIR, TEM and AFM. Journal of Membrane Science 256: 134--142. Hagg MB (2008) Membranes in gas separations. In: Pabby AK, Rizvi SSH, and Sastre AM (eds.) Handbook of Membrane Separations: Chemical, Pharmaceutical and Biotechnological Applications, pp. 65--107. Boca Raton, FL: CRC Press. Ho WS and Sirkar KK (1992) Membrane Handbook. New York: Chapman and Hall. Jeong B-H, Hoek EMV, and Yan Y (2007) Interfacial polymerization of thin film nanocomposites: A new concept for RO membranes. Journal of Membrane Science 294(1–2): 1--7. Jones KL and O’Melia CR (2000) Protein and humic acid adsorption onto hydrophilic membrane surfaces: Effects of pH and ionic strength. Journal of Membrane Science 165: 31--46. Judd S (2006) The MBR Book. Oxford: Elsevier. Kennedy MD, Kamanyi J, Salinas SS, Lee NH, Schippers JC, and Amy G (2008) Water treatment by microfiltration and ultrafiltration. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 131--170. Hoboken, NJ: Wiley. Kesting RE (1985) Phase inversion membranes. In: Lloyd DR (ed.) Materials Science of Synthetic Membranes, ACS Symposium Series, vol. 269, ch. 7, pp. 131--164. Washington, DC: American Chemical Society. Khayet H (2008) Membrane distillation. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 297--370. Hoboken, NJ: Wiley. Khedr MG (2003) Development of reverse osmosis desalination membranes composition and configuration: Future prospects. Desalination 153: 295--304. Kim KJ, Fane AG, Fell CJD, Suzuki T, and Dickson MR (1990) Quantitative microscopic study of surface characteristics of ultrafiltration membranes. Journal of Membrane Science 54: 89--102. Kimura K, Amy G, Drewes JE, Heberer T, Kim TU, and Watanabe Y (2003) Rejection of organic micropollutants (disinfection by-products, endocrine disrupting compounds, and pharmaceutically active compounds) by NF/RO membranes. Journal of Membrane Science 227: 113--121. Kumano A and Fujiwara N (2008) Cellulose triacetate membranes for reverse osmosis. In: Li NN, Fane AG, Ho WSW, and Matsuura T (eds.) Advanced Membrane Technology and Applications, pp. 21--46. Hoboken, NJ: Wiley.
Kwon YN, Tang CY, and Leckie JO (2006) Change of membrane performance due to chlorination of crosslinked polyamide membranes. Journal of Applied Polymer Science 102: 5895--5902. Kwon YN, Tang CY, and Leckie JO (2008) Change of chemical composition and hydrogen bonding behavior due to chlorination of crosslinked polyamide membranes. Journal of Applied Polymer Science 108: 2061--2066. Lebeau T, Lelievre C, Buisson H, Cleret D, De Venter LWV, and Cote P (1998) Immersed membrane filtration for the production of drinking water: Combination with PAC for NOM and SOCs removal. Desalination 117: 219--231. Le-Clech P, Chen V, and Fane AG (2006) Fouling in membrane bioreactors used in wastewater treatment: A review. Journal of Membrane Science 284(1–2): 17--53. Lee S and Elimelech M (2006) Relating organic fouling of reverse osmosis membranes to intermolecular adhesion forces. Environmental Science and Technology 40: 980--987. Li K (2007) Ceramic Membranes for Separation and Reaction. Chichester: Wiley. Lieknes TO (2009) Wastewater treatment by membrane bioreactors. In: Drioli E and Giorno L (eds.) Membrane Operations, Innovative Separations and Transformations, pp. 363--396. Weinheim: Wiley-VCH. Liu LH, Gao SJ, Yu YH, Wang R, Liang DT, and Liu M (2006) Bio-ceramic hollow fiber membranes for immunoisolation and gene delivery – I: Membrane development. Journal of Membrane Science 280: 375--382. Loeb S and Sourirajan S (1964) High Flow Semipermeable Membrane for Separation of Water from Saline Solutions. US Pat. 3,133,132, 12 May 1964. Louie JS, Pinnau I, Ciobanu I, Ishida KP, Ng A, and Reinhard M (2006) Effects of polyether-polyamide block copolymer coating on performance and fouling of reverse osmosis membranes. Journal of Membrane Science 280: 762--770. Mahon HI (1966) Permeability Separatory Apparatus and Membrane Element, Method of Making the Same and Process Ultilizing the Same. US Pat. 3228876, 11 January 1966. Mckelvey SA, Clausi DT, and Koros WJ (1997) A guide to establishing hollow fiber macroscopic properties for membrane applications. Journal of Membrane Science 124: 223. Mulder M (1996) Basic Principles of Membrane Technology, 2nd edn. Dordrecht: Kluwer. Nakao S (1994) Determination of pore size and pore size distribution. 3. Filtration membranes. Journal of Membrane Science 96: 131--165. Pearce G (2007) Introduction to membranes: Membrane selection. Filtration and Separation 44: 35--37. Petersen RJ (1993) Composite reverse-osmosis and nanofiltration membranes. Journal of Membrane Science 83: 81--150. Qin JJ, Wang R, and Chung TS (2001) Investigation of shear stress effect within a spinneret on flux, separation and thermomechanical properties of hollow fiber ultrafiltration membranes. Journal of Membrane Science 175: 197. Ren JZ and Wang R (2010) Preparation of polymeric membranes. In: Wang LK, Chen JP, Hung YT, and Shammas NK (eds.) Handbook of Environmental Engineering, vol. 13, ch. 2. Totowa, NJ: Humana Press (in press). Schafer AI, Fane AG, and Waite TD (2005) Nanofiltration – Principles and Applications. Oxford: Elsevier. Schock G and Miquel A (1987) Mass transfer and pressure loss in spiral wound modules. Desalination 64: 339--352. Seah H, Tan TP, Chong ML, and Leong J (2008) NEWater – multi safety barrier approach for indirect potable use. Water Science and Technology 8(5): 573--588. Shi L, Wang R, Cao YM, Liang DT, and Tay JH (2008) Effect of additives on the fabrication of poly(vinylidene fluoride-co-hexafluropropylene) (PVDF-HFP) asymmetric microporous hollow fiber membranes. Journal of Membrane Science 315: 195--204. Strathmann H (1990) Synthetic membranes and their preparation. In: Porter M (ed.) Handbook of Industrial Membrane Technology, pp. 1--60. Park Ridge, NJ: Noyes Publications. Tang CY and Leckie JO (2007) Membrane independent limiting flux for RO and NF membranes fouled by humic acid. Environmental Science and Technology 41: 4767--4773. Tang CY, Kwon YN, and Leckie JO (2007a) Probing the nano- and micro-scales of reverse osmosis membranes – a comprehensive characterization of physiochemical properties of uncoated and coated membranes by XPS, TEM, ATRFTIR, and streaming potential measurements. Journal of Membrane Science 287: 146--156. Tang CY, Kwon YN, and Leckie JO (2007b) Fouling of reverse osmosis and nanofiltration membranes by humic acid – effects of solution composition and hydrodynamic conditions. Journal of Membrane Science 290: 86--94.
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Tang CY, Kwon YN, and Leckie JO (2009a) Effect of membrane chemistry and coating layer on physiochemical properties of thin film composite polyamide RO and NF membranes II. Membrane physiochemical properties and their dependence on polyamide and coating layers. Desalination 242: 168--182. Tang CY, Kwon YN, and Leckie JO (2009b) Effect of membrane chemistry and coating layer on physiochemical properties of thin film composite polyamide RO and NF membranes. I. FTIR and XPS characterization of polyamide and coating layer chemistry. Desalination 242: 149--167. Tang CY, Kwon YN, and Leckie JO (2009c) The role of foulant–foulant electrostatic interaction on limiting flux for RO and NF membranes during humic acid foulingtheoretical basis, experimental evidence, and AFM interaction force measurement. Journal of Membrane Science 326: 526--532.
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4.12 Wastewater as a Source of Energy, Nutrients, and Service Water P Cornel, A Meda, and S Bieker, Technische Universita¨t Darmstadt, Darmstadt, Germany & 2011 Elsevier B.V. All rights reserved.
4.12.1 4.12.2 4.12.2.1 4.12.2.2 4.12.2.2.1 4.12.2.2.2 4.12.2.3 4.12.3 4.12.3.1 4.12.3.2 4.12.3.3 4.12.4 4.12.4.1 4.12.4.2 4.12.5 4.12.5.1 4.12.5.2 4.12.5.2.1 4.12.5.2.2 4.12.5.2.3 4.12.5.2.4 4.12.5.2.5 4.12.6 4.12.6.1 4.12.6.1.1 4.12.6.1.2 4.12.6.1.3 4.12.6.1.4 4.12.6.2 4.12.6.3 4.12.6.3.1 4.12.6.3.2 4.12.6.3.3 4.12.6.3.4 4.12.6.3.5 4.12.7 4.12.7.1 4.12.7.2 4.12.7.3 4.12.7.4 4.12.8 References
Introduction Resources of Interest Energy Nutrients Nitrogen Phosphorus Water Origin and Amounts of Resources Energy Nutrients Water Energy Caloric Heat Degradable Organic Constituents Nutrients Nitrogen Recovery Phosphorus Phosphorus recovery during wastewater treatment Phosphorus recovery from sewage sludge – wet chemical technology Phosphorus recovery from sewage sludge – thermochemical technologies Products from phosphorus recovery processes Exemplary applications of phosphorus recovery Water Reuse Reuse Options Agricultural reuse Intra-urban reuse as service water Industrial reuse Groundwater recharge Fit for Purpose, Quality Requirements Treatment Options and Energy Requirements Physical and chemical methods Biological treatment Disinfection Other methods Energy requirements Recovery Fosters Decentralization Water Energy System Scale Case Study: Qingdao Summary and Outlook
4.12.1 Introduction With continuously improving analytical techniques, it becomes more and more obvious that municipal wastewater represents a multisubstance mixture containing probably several hundreds of different substances. Normally, human excrements play a major role. Urine and feces contain nonexploited residuals of ingested food and
337 338 339 340 340 341 342 344 344 350 350 352 352 353 356 357 358 358 358 358 358 359 360 361 361 362 363 364 365 366 366 367 367 368 368 368 369 369 369 370 371 372
their degradation products. Fats, proteins, carbohydrates, and other carbon compounds; ammonia, urea, and other organic nitrogen compounds; organically bound phosphorus and dissolved ortho-phosphate; organic sulfur compounds; and potassium are the main components. Their amounts can be estimated and balanced per person. In addition, metals/heavy metals such as iron, copper, and zinc are essential trace elements in human nutrition. They are taken up with food,
337
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excreted via feces/urine, and thereby end up in the wastewater in the same way as salts, sodium chloride in particular, and natural hormones such as estrogens. Human excrements are also the main source for pathogens, coliforms, viruses, protozoans, and helminth eggs. Adding chopped organic (kitchen) waste to wastewater, as is practiced in some cases, increases the load of organic carbon, nitrogen, and phosphorus compounds. Calcium and magnesium salts are introduced via drinking water; and nitrate, sulfate, iron, and manganese are from geogenic sources, for example, via the use of groundwater as drinking water resource. Salt concentrations are increased by washing, dishwashing, and other cleaning agents, disinfectants and salts for water softening, and the regeneration of ion exchangers in domestic appliances and installations (potassium, sodium, calcium, and magnesium). Depending on the composition of the washing/cleaning agents, polyphosphates and zeolith A (source of aluminum) are also introduced into the wastewater. Due to the use of perborate as a bleaching agent, boron can be found in urban wastewater at an average concentration of around 1 mg l1. Boron can become toxic at levels only slightly greater than those required by plants for good growth (Lazarova and Bahri, 2005). Copper and zinc in wastewater mainly originate from domestic installations and roof runoff. In some parts of the world, they are also the source of lead in wastewater. The main component of halogens in wastewater is chlorine. Sources of organic halogens are mainly from commercial and industrial activities. Concentrations can vary significantly, and the typical amount for municipal wastewater in Germany is approximately 100 mg l1 measured as adsorbable organic halogens (AOXs) (Imhoff, 2007). During the last few years, one of the main focuses of research has been on the so-called ‘micro-pollutants’, substances of ecotoxicological relevance which differ considerably in their chemical character and their physical behavior. There are acidic, neutral, as well as alkaline compounds with hydrophilic as well as hydrophobic character. In common, they have a low biological degradability. Here, the term micro-pollutants includes drugs, diagnostics, cleaning and personal care products, industrial chemicals, as well as endocrinically active substances which are found in wastewater in low concentrations, that is, generally in ng l1 to mg l1 levels. Hospitals and nursing homes, industrial and commercial activities are some of the sources; however, the main source is domestic wastewater. In the European Union, approximately 3000 different pharmaceuticals are in use, including analgetics, antirheumatics, antibiotics, antiepileptics, lipid-lowering drugs, beta blockers, cytostatic drugs, radiopaque materials, contraceptives, tranquilizers, and virility drugs (Ternes, 2004). In addition, there are cleaning and personal care products, such as shampoos and other hair care products, bath essences, skin, oral and dental care products, soaps, sunscreen agents, as well as perfumes and aftershaves. These care products normally contain persistent fragrances such as musk, ultraviolet (UV) blockers, and preservatives, which reach the wastewater by human excretion or during/after external use. While the output of pharmaceuticals in Germany varies – depending on
the kind of drug – between a few kilograms per substance and several tens of tons for lipid-lowering drugs, antirheumatics (up to 500 Mg a1), and antibiotics (Forth et al., 1996, cited in Wilken and Ternes, 2001), without considering veterinary medicines, in 1993, more than 550 000 t personal care products were produced (Ternes, 2004). Some pharmaceutical products, for example, contraceptives or antibiotics, belong to the group of so-called endocrine disruptors. This term includes those substances affecting the hormonal balance. They may be not only natural hormones (estrogens), but also substances, which function like hormones without being hormones in the actual sense (xeno-estrogens). The hormones with the highest known activity are the natural estrogen 17b-estradiol and the synthetic estrogens 17a-ethinylestradiol and mestranol (contraceptives). Their concentration in wastewater is in the ng l1 level (Kunst et al., 2002). Besides hormones, a large variety of other substances with undesirable/unintentional endocrine activity gets into the wastewater, from industrial as well as municipal sources. Among many others, relevant substances are nonylphenols, bisphenol A (BPA), organotin compounds, benzylbutylphthalate, dibutylphthalate, as well as some pesticides, pharmaceuticals, polycyclic aromatic hydrocarbons (PAHs), and polychlorinated biphenyls (PCBs). A broad range of very different substances reaches the wastewater via commercial and industrial dischargers. These include the noble metals such as gold and silver, detergents, and disinfectants, partly based on phosphorus, acids, and bases; there is hardly any outline, and as Kroiss states: ‘‘ The only compounds which can reliably be prevented from the wastewater y are those which are not produced’’ (Kroiss, 2004).
However, wastewater consists of H2O to more than 99.5%, a purity which many purchasable products will never reach. This means that, on the other hand, potentially valuable substances are present in extreme dilution, in concentrations of mg l1 down to ng l1, and mixed with all sorts of other substances. This is a result of the conventional flushing sewer concept and makes an economic reuse of individual valuable substances at least questionable. When talking of wastewater as a resource, one has to face the ambivalent questions:
• •
What prerequisite the wastewater has to fulfill to become a resource? Under which conditions wastewater constituents may be called a valuable substance and when should it be named a pollutant? J Does the classification depend on the purpose of use? Nutrients can be valuable in irrigation water; however, they can have harmful impacts on water bodies. J Does the classification depend on the concentration of the substance in the wastewater? J Are bonding form, admixtures, and/or impurities responsible to turn a resource into a pollutant? J Are effort and complexity of separating the substance from the wastewater and its recycling the reason? The energy demand? The emission of greenhouse gases?
Wastewater as a Source of Energy, Nutrients, and Service Water
J
Is the shortage of the substance an issue? That is, the real shortage in the sense of the finite nature of resources, as it seems to be the case for phosphorus and/or the economic shortage.
These questions seem trivial at first sight. However, they can hardly be answered with a universally valid explanation, the more so as the answers seem to be subject to changes during time, as can be demonstrated with the example of phosphorus/phosphate. Until 30 years ago, phosphorus in wastewater did hardly attract any interest. Only later, phosphorus moved into the focus of wastewater treatment, when its eutrophicating impact on freshwaters and saline waters with low water exchange rates became obvious. Methods for phosphorus elimination were developed and were introduced into wastewater treatment plants almost all over Central Europe. The task was to separate phosphorus. Nowadays, hardly 15 years later, efforts have started to reclaim phosphorus from wastewater to be used as resource. What has changed? Neither the quality and the bonding form of phosphorus in wastewater has changed, nor its quantity/concentration. Obviously, our point of view has changed, initiated by new insights and the interdependency with improved technical prospects for recovery. These new methods enable the economic recovery and recycling of phosphorus and thereby transform phosphorus from being a pollutant to a resource. Whether we consider a water constituent as valuable substance or as potentially hazardous material depends on the regional conditions and is subject to temporary changes. In order to justify the efforts for an improved materials flow management and technical measures for concentration/enrichment, purification, and recovery of substances, it is therefore important to critically analyze the relevance of these measures and to evaluate them in comparison with alternative options of energy conversion and fertilizer production.
4.12.2 Resources of Interest The 2000 UN General Assembly Millennium Meeting established eight Millennium Development Goals (MDGs) with targets to be achieved by 2015 starting from the baseline in 1990. One of the water-related targets is ‘‘Halve by 2015 the proportion of people without sustainable access to safe drinking water and basic sanitation.’’ The MDG further states that ‘‘In 2002, nearly half of the developing world (2.5 billion people) had no access to proper sanitation,’’ most of those (1.98 billion) in Asia (UNESCO, 2006). Will we meet the targets? The WHO and UNICEF Joint Monitoring Program 2008 summarizes (WHO and UNICEF, 2008):
•
•
‘‘ The world is on track to meet the MDG target on drinking water. Current trends suggest that more than 90% of the global population will use improved drinking water sources by 2015. The world is not on track to meet the MDG sanitation target. Between 1990 and 2006 the proportion of people without improved sanitation decreased by only 8 percentage points.’’ At least two-thirds of the population in 34 countries are not using improved sanitation facilities.
339
‘‘Without an immediate acceleration in progress, the world will not achieve even half the sanitation target by 2015. Based on current trends, the total population without improved sanitation in 2015 will have decreased only slightly, from 2.5 billion to 2.4 billion. At the current rate, the world will miss the MDG sanitation target by over 700 million people. To meet the target, at least 173 million people on average per year will need to begin using improved sanitation facilities.’’ Wilderer pointed out that in order to meet the MDGs, every day wastewater facilities encompassing collection, transportation in sewers and treatment units must be built serving 900 000 people within the years 2005–15, considering 300 working days per year, (Wilderer, 2005a). This does not seem realistic at all. Against this background, the question is: What would municipal water and wastewater management look like, if we designed it on the drawing board, with our present knowledge and experience, but without including already existing supply and disposal infrastructure? The current concept of urban sanitation is based on the perception to avoid direct contact of humans with their own wastewater and flush away the pathogenic organisms, germs, and viruses potentially contained in the feces together with the urine and wastewater from shower, kitchen, and laundry. The implementation of flushing toilets and the flushing sewer concept is of course a success story in reducing diseases and offers high convenience to the users. To avoid negative ecological and economic impacts to the receiving water bodies, wastewater treatment technology was developed and systematically further developed and implemented. The main objective of wastewater treatment is to convert pollutants into less problematic substances prior to discharge to any surface water body (Wilderer, 2005b). The current system hardly makes use of valuable substances such as nutrients, organic matter as energy source, fatty acids as raw material in industrial chemistry, and of the purified water itself (Wilderer, 2005b). Against this background, the economic, social, and ecological necessity for resource-conserving handling of water and energy as well as the identification and development of potentials for resource recovery from wastewater should be looked at first. From the multitude of compounds and resources existent in water, the following topics – according to the chapter’s title – are focused on: 1. energy, 2. nutrients, phosphorus in particular, and 3. water.
4.12.2.1 Energy Finite fossil fuels as well as the increase in atmospheric concentrations of greenhouse gases due to anthropogenic activities have given priority to the worldwide discussion and efforts to increase energy efficiency in all processes. There are multifold interactions between water and energy:
• •
Electric energy can be generated by using hydro power. Water is necessary for energy production.
340
•
• •
Wastewater as a Source of Energy, Nutrients, and Service Water
Water is used for mining fuels. Taking into account the local conditions of Queensland/Australia, Keller estimates an amount of 2–3 l of water per electric kW h generated from coal (Keller, 2008). Cornel and Meda (2008a) estimate the specific water demand for agricultural energy production via biogas generation in Germany at approximately 300 l kW h.
•
Power plant cooling uses water.
Moreover, vice versa, there is manifold use of energy in water supply and disposal:
•
•
•
Supply and conveyance. If necessary, water is conveyed from large depths or across large distances, for example in southern California which imports approximately 50% of its water supply from the Colorado River and the State Water Project (California Energy Commission, 2005: 144). For California, a range of 0–1.06 kW h m3 for water supply and conveyance is mentioned in the Integrated Energy Policy Report. For southern California, with its particularly unfavorable situation, it is even 2.35 kW h m3. In Germany, the city of Stuttgart is supplied with water from Lake Constance, 150 km to the south, thereby creating an energy demand for pumping of approximately 1.1 kWh m3 (Bodensee Wasserversorgung, 2010). This figure makes clear that, even in a country with overall sufficient water resources such as Germany, water may be scarce on a local scale, requiring energy-intensive transports of water over long distances. Treatment. The energy demand for generating drinking water from existing water resources and eliminating impurities varies from almost zero to several kilowatt-hours per cubic meter. The Integrated Energy Policy Report states it to be up to 4.2 kW h m3. For desalination of brackish water and saltwater, currently one has to calculate approximately 4 kW h m3 (Keller, 2008). In Germany, as a relatively water-rich country, the energy consumption for water catchment and treatment ranges from 0.21 to 0.40 kW h m3 (Ro¨dl and Partner, 2006, 2007). Distribution. The energy demand for distributing water to consumers is mainly caused by costs for pumping and depends on topography, distances, pipe cross sections, water pressure as well as size and age of the distribution system. Between 0.18 and 0.32 kW h m3 are quoted for California (California Energy Commission, 2005: 144). Ha¨hnlein mentions 0.54 kW h m3 for treatment and distribution for a large German city, whereby more than half is caused by energy for pumping (Ha¨hnlein, 2008). Ro¨dl and Partner (2006, 2007) quote 0.06–0.17 kW h m3 as energy consumption for water distribution in Germany. In addition, water losses caused by leakages have to be considered. Water loss rates of up to 50% are not uncommon in urban distribution systems (UNESCO, 2009: 58). High losses require the extraction, treatment, and transport of greater volumes of water than the customer demand requires. As some leakage finds its way into community waste or storm water collection, it will be treated by the local wastewater plant causing additional energy demand with no beneficial use.
•
•
End consumers. At the consumer’s end, the main energy users are posttreatment of water such as softening, filtration, disinfection, as practiced in some places, pressure increase especially in multistory buildings, and – in the first place – hot water production for personal hygiene, washing, and dishwashing. For urban uses (residential, commercial, and industrial) in California in 2001, the Integrated Energy Policy Report states an electricity consumption of 27 887 GW h and a natural gas consumption of 4220 US therms, corresponding to 123 646 GW h (1 US therm ¼ 29.3 kW h). Considering a water use in the urban sector of 11 128 million m3, one can calculate an average end-use energy intensity of 2.5 kW h m3 considering only electricity and 13.6 kW h m3 considering both electricity and natural gas. Neglecting any losses, the latter value is equivalent to an average water temperature increase of almost 12 1C within the household. Wastewater collection and treatment. The energy demand for collecting and treating wastewater depends – on the one hand – on the length of the sewer system and the topography, and – on the other hand – on the treatment requirements, the selected treatment technique, and the size of the wastewater treatment plant. Energy demand values as quoted for California, 0.29–1.2 kW h m3, may increase due to stringent quality requirements for nitrogen and phosphorus discharge values, for helminth eggs, and other hygienic parameters or by the commitment to specific treatment techniques, for example, aerobic membrane-activated sludge systems or desalination membranes. For comparison, in Germany, the average energy demand for wastewater treatment (without collection) amounts to 0.44 kW h m3 including infiltration water and partially storm water (Haberkern et al., 2008). Water discharge. Last but not least, energy is needed for operating the electric pumps required to discharge the effluent of the wastewater treatment plants to the receiving water body (o0.11 kW h m3) or for further water treatment and use.
Table 1 points out the ranges of energy intensities in the water cycle using the example of California and Germany. The data are expressed as energy consumption per water volume (kW h m3) and energy consumption per person and year (kW h(C a)1), assuming a specific water consumption of 100 m3(C a)1 for California according to Asano (2007), and 45 m3(C a)1 for Germany according to the Federal Statistical Office of Germany (FSO) (FSO, 2009). Although the energy intensities are nonadditive, they reflect the ranges of the energy demand for the water use cycle from approximately 0.5 to almost 7 kW h m3 without the waterrelated energy use of the end users. The latter is in average approximately 2.5 kW h m3 in electrical power and 13.6 kW h m3 considering both electricity and natural gas. Altogether, approximately 15% of the total energy consumption of California are classified as water-related, increasing to 19% when including agricultural irrigation (Reiter, 2008). Considering these numbers, the California Energy Commission arrives at the conclusion: ‘‘The link between energy and water use in the state is an important facet of California’s
Wastewater as a Source of Energy, Nutrients, and Service Water Table 1
341
Energy intensities in the water cycle (on the basis of California Energy Commission, 2005, p. 144 and Ro¨dl and Partner, 2006, 2007)
Water cycle segment
Range of energy intensity California 1
kW h
Supply and conveyance Treatment Distribution Wastewater collection and treatment Wastewater discharge
Germany 3
1
m
kW h
(C a)
1a
kW h1 m3
kW h1 (C a)1b
Low
High
Low
High
Low
High
Low
High
0 0.03 0.18 0.29 0
1.06 4.23 0.32 1.22 0.11
0 3 18 29 0
106 423 32 122 11
0.21
0.40
9
18
0.06 0.39
0.17 0.83
3 25c
8 80c
a
Assuming a specific water consumption of 100 m3 (C a)1 according to Asano (2007). Assuming a specific water consumption of 45 m3 (C a)1 according to FSO (2009). c From Keicher K, Krampe J, and Steinmetz H (2008). Eigenenergieversorgung von Kla¨ranlagen. Korrespondenz Abwasser 55(6): 644–650. b
energy system. While the most immediately recognizable aspect of this link is large-scale hydroelectric generation, the amount of energy used by the state’s water infrastructure and water end-users is at least equally significant – and growing fast. The Energy Commission evaluated the relationship between water and energy systems to better understand this link and determine what, if any, mutually beneficial strategies can be developed to improve both the water and energy sectors. As a result of this initial work, the Energy Commission determined that much can be done to improve both systems’’ (California Energy Commission, 2005: 138). It should be mentioned that in other countries water consumption and specific energy intensities might be much lower than in California. In Germany, for example, private households consume only 45 m3(C a)1 compared to 100 m3(C a)1 in California. Combined with the lower specific energy demand due to shorter transport distances, less energy-intensive water-processing techniques (e.g., no seawater desalination plants), and the use of more energy-efficient wastewater treatment processes, the energy demand of the entire water cycle amounts to only 30–120 kWh (C a)1, compared to 50–700 kW h (C a)1 in California. In addition, just how much energy can be produced from the wastewater’s organic matter? With 110–120 g COD (C d)1 (chemical oxygen demand (COD)), the total energy content per person and day is approximately 0.4 kW h. Currently, approximately 0.05–0.1 kW h thereof can be generated as electric energy. That is, depending on the specific water consumption per capita, around 0.15–0.7 kW h m3 and as such, far less than required for water supply, distribution, collection, and treatment (see Section 4.12.4.2 for a detailed description). Thus, wastewater as a source of energy means that besides the potential recovery of energy from wastewater, the entire technology chain starting from water supply, treatment, and distribution to water use at the consumer’s end as far as wastewater collection, treatment, and discharge has to be considered and energetically optimized. The biggest saving potential lies with the reduction of the energy consumption. In Section 4.12.4, the recovery potentials are looked at more specifically and the greenhouse gas emissions will be taken into account.
4.12.2.2 Nutrients The main nutrients in wastewater are nitrogen and phosphorus.
4.12.2.2.1 Nitrogen Nitrogen is available worldwide in sufficient amounts. It is one of the most common elements. By far the largest quantity is found in the atmosphere which consists of approximately 78 vol.% N2. Thus, raw material for nitrogen-based fertilizers is available in sufficient quantities; however, their production from elementary nitrogen or air, respectively, using, for example, the Haber–Bosch process is energy-intensive. In the Haber–Bosch process nitrogen and hydrogen react – at high pressure and increased temperature and in the presence of catalysts – according to the equation
N2 þ 3H2 ¼ 2NH3 þ 92:1 kJ Depending on the process, approximately 9–13 kW h kg1 NH3–N are needed (Mundo, 1970). According to Larsen et al. (2007), a value of 12.5 kW h kg1 NH3–N can be assumed respectively 9–11 kW h kg1 NH3–N according to EFMA (2000). The incentive to recover nitrogen from wastewater, mostly present as the ammonium ion, therefore results from the energy-saving potential rather than a finiteness of nitrogen itself. Considering that the nitrogen excretion amounts to 11 g per person and day, that is, approximately 4 kg per person and year, the energy-saving potential is a maximum of 40 kW h per person and year, in case all the ammonium ions contained in wastewater could replace ammonia produced by the Haber–Bosch process. In addition, around 3.9–6.9 kWh (C a)1 could be saved at the wastewater treatment plant (WWTP) in case nitrification/denitrification could be omitted. (The energy consumption for nitrogen removal on WWTP is discussed in Section 4.12.3.2) Therefore, recovery of nitrogen from wastewater seems substantial and worth to be considered. Besides the general assessment of mass and energy balances, the positive response to a series of pragmatic questions is a precondition for the realization of the direct use of
342
Wastewater as a Source of Energy, Nutrients, and Service Water 4.12.2.2.2 Phosphorus
nitrogen in wastewater as fertilizer or the use of recovery techniques:
•
Phosphorus is an essential element for all organisms. Besides carbon, hydrogen, oxygen, and nitrogen, phosphorus is one of the vital components of the DNA and the key element of the energy supplier adenosine triphosphate (ATP). Phosphorus is an essential nutrient. Already Justus von Liebig (1803–73) identified phosphorus as the limiting factor for plant growth. As a vital cell component, phosphorus cannot be replaced by any other element. This is why phosphorus is different from other resources, such as fossil fuels, where there are potential alternatives, or from nitrogen fertilizers, which can be technically produced from air nitrogen via the Haber–Bosch process. In nature, phosphorus passes through several interconnected cycles (Figure 2). The inorganic cycle describes the cycle from erosion, transport to the oceans, sedimentation, tectonic uplift, and alteration of phosphate-containing rocks into plant-available phosphates in soil (Emsley, 1980, 2001; Filippelli, 2002). The cycle time of this cycle is several million years, that is, in human spaces of times, phosphate transported into the oceans can be considered as lost for agricultural use. Besides the inorganic phosphorus cycle, there are two organic cycles attached describing phosphorus as part of the food chain. One of the cycles takes place on land (soil–plants– humans/animals–organic waste–soil) and the other in water. The cycle time of these cycles is between a few weeks and up to 1 year (Emsley, 1980, 2001; Bennett and Carpenter, 2002). These originally natural closed cycles are interrupted when phosphorus compounds in animal and human excrements are not used in fertilization. Then, phosphate contained in wastewater is partly transported to the oceans via the discharge systems, partly fixed in sewage sludge, which is deposited in landfill sites or incinerated; in the latter case, phosphorus contained in the ash is deposited in landfill sites or in subterranean storage. The procedure can be similar with organic fertilizers (solid and liquid manure) from intensive stock rearing. The deficit is balanced by chemical fertilizers, that is, the mining of phosphate-rich deposits in the earth’s crust. In Figure 2, the geological and biological cycles are illustrated, including changes due to human impact. The quality of rock phosphate not only depends on its phosphate concentration, but also on its concentration of harmful substances, cadmium and uranium in particular (UBA, 2001; Kratz, 2004). In order to restrict cadmium concentrations in processed ores and, where required, in mineral fertilizers, one has to expect increasing costs for the processing of rock phosphate (ATV-DVWK, 2003) in the future.
What is the required energy input for the recovery process? Is this energy demand less than the realized savings by taking into account storage, transport, and the fertilizing techniques themselves?
•
What about the storage suitability of nitrogen? As fertilizers can only be used during vegetation periods, storage ability, space requirement, product stability, and its manageability are decisive cost factors.
• •
How is the quality of the fertilizer, that is, concentration, contamination, and plant availability? Is this quality constant and guaranteed? What about manageability? Is it feasible to use common equipment for spreading? Is the handling hygienically acceptable?
Last but not least, when discussing this matter one has to add certain constraints, as only a small percentage of the nitrogen employed in fertilizer production can be brought back to agricultural use, even with 100% recovery of the nitrogen existent in the wastewater. Only a fraction of the produced nitrogen reaches the consumer, is taken up with food, excreted, and reaches the wastewater. In vegetarian food, approximately 14% of the nitrogen applied as fertilizer is contained in the food, with meat, only 4% reaches the consumer, as shown in Figure 1 (Galloway et al., 2003 cited in Kroiss, 2006). Correspondingly low is the potential rate of nitrogen fertilizer recovered from wastewater in the fertilizer cycle. According to other authors, the percentage of nitrogen applied as fertilizer that finally reaches the consumer via the food chain and may then potentially be recovered from wastewater, ranges from 14% to 20% (Maurer et al., 2003; Zessner et al., 2010). Even considering these higher numbers, it is clear that nitrogen recovery cannot close the loop of the anthropogenic nitrogen cycle; however, it can make a small but important contribution in reducing the chemical fertilizer consumption. Similar circumstances apply to phosphorus recovery from wastewater, with the important difference that phosphorus is a limited, irreplaceable element. Therefore, its reclamation should be realized with higher priority.
N fertilizer produced 100
−6
N fertilizer applied 94
−47
N in crop
N in feed
N in store
N consumed
47
31
7
4
−16
−24
−3
Figure 1 Fate of nitrogen produced by the Haber–Bosch process in the course of meat production (Galloway et al., 2003, cited in Kroiss, 2006).
Wastewater as a Source of Energy, Nutrients, and Service Water
Human, animal Detergent Industry
Sewer system Organic waste
Plants
343
WWTP
Fertilization Agriculture Mining Erosion
Weathering
Rivers, oceans
Phosphate rock
Cycle times Sedimentation
Tectonic uplift
1−5 years 106−109 years
Sediments
??? years
Figure 2 Geological (inorganic) and organic (land) phosphorus cycles (Bennett and Carpenter, 2002 (cited in Pinnekamp, 2002) modified including human impacts, phosphorus cycle in water not included).
The price increase for phosphorus which has been observed in recent years, however, is attributed to speculations rather than being a consequence of increasing treatment costs. Rising energy costs also play a role, either directly using thermal processes or indirectly using the acid process with phosphate rock digestion, as the price for the required sulfuric acid also strongly depends on energy prices. However, in contrast to nitrogen, the availability of phosphorus is limited. Based on the assessment of the future consumption of phosphorus fertilizers, the availability of presently payable natural phosphate deposits is forecasted to be approximately 60–240 years (Steen, 1998, cf. IFA, 1998). Possible impacts on these estimations are the population development and thus the consumption of fertilizers as well as the activation of those natural phosphate deposits, which are not payable under present technical and economic views. Regional phosphate deposits vary distinctively. Currently, approximately two-thirds of the phosphate rocks are mined in USA, Morocco, and China. In addition, considerable amounts are mined in Russia, Brazil, Israel, Jordan, South Africa, and Tunisia. With regard to phosphorus recovery, there are also questions as to where recovery takes place (within the wastewater treatment process), what are the efforts, and in which form phosphorus is reclaimed (see Section 4.12.5).
4.12.2.3 Water Water resources are limited worldwide and with the still growing demand present a globally increasing problem. Thereby, the population of threshold and developing countries in arid and semi-arid regions as well as densely populated regions and megacities are affected in particular. Besides climatic conditions and a generally uneven distribution of water resources, population growth, increasing per capita water
consumption, conflicts of use, and increasing urbanization are the main causes for the growing (regional) shortage of freshwater. Often, a rather unsustainable handling of existing water resources and the contamination of surface and groundwater aggravate the situation. According to the forecast of the World Water Development Report (UNESCO, 2006) and based on the worst-case scenario, approximately 7 billion people in 60 countries will be confronted with water shortage until the middle of the current century if the present consumption habits do not change. The best-case scenario predicts that still at least 2 billion people in 48 countries will suffer from water shortage. In addition, experts from the Intergovernmental Panel on Climate Change (IPCC, 2007) forecast a further increase of global water shortage due to the effects of global climate change. Water is subject to regionally uneven distribution. In order to describe the (regional) water availability, different indices are used which on the one hand reveal tendencies, on the other hand should be carefully interpreted as to their absolute quantity and should be critically challenged (Box 1). One such parameter is the specific per capita availability of renewable freshwater, a number which depicts the calculatory potential and not the used water volume. The water intensity use index (also called water stress index) expresses the ratio of the mean total annual water withdrawal to the total renewable freshwater resources (Jime´nez and Asano, 2008). Table 2 gives threshold values for the two parameters used to characterize water stress situations. At a global level, water availability for 2006 was 8462 m3 (C a)1, but at a regional level it varies from as little as 1380 m3 (C a)1 in the Middle East and North Africa to almost 53 300 m3 (C a)1 in Oceania. These figures do not reflect the situation of individual countries or within a country. A list of the countries that are water-scarce according to the
344
Box 1
Wastewater as a Source of Energy, Nutrients, and Service Water
Can statistics lie?
Does Germany have a water stress problem similar to, for example, Spain, as the water stress indexes in Table 4 indicate? Everybody who knows these countries is surprised by this result. On the one hand, there is Germany’s green landscape and forests throughout the whole summer, with enough rain to almost abandon irrigation, and there is the dry landscape in Spain, on the other hand, where millions of Germans spend their holidays each year because of the nice, rain-free weather and where intensive agriculture is not thinkable without irrigation. Why does this subjective difference not match the statistical data? One key might be that the withdrawal data do not distinguish between consumptive and nonconsumptive uses. Water extracted for consumptive uses – especially for irrigation, where it is evaporated by plants – is no longer available for other uses, and as a consequence puts higher pressure on water resources than nonconsumptive uses such as cooling in power plants, where most of the water is returned to the water body with almost no quality deterioration and can be used again. The same is true for most of the municipal and industrial waters, which do not disappear by use but are returned to the rivers as (treated) wastewater and might be used again downstream. In addition to the fact that the water, after nonconsumptive use, is still available and, as a result, puts less pressure on the water resources, it has to be considered that in the published statistical data this amount of water is counted as leaving the countries by rivers and, as a consequence, lowers the calculated water resources and increases the water stress index. As this example shows, the method for calculating water stress indexes is questionable and needs at least careful interpretation. The numbers for Germany clearly illustrate this fact. Using the data of the FAO – the data of other sources, for example, German Federeal Statstical Office (FSO, 2009) are up to 15% lower – the water stress index for Germany can be calculated as 31% by dividing the withdrawal of 47 050 million m3 a1 by the natural renewable water resources of 154 000 million m3 a1. The statistical yearbook of Germany states that about 26 000 million m3 a1 of the withdrawn water is cooling water (FSO, 2009) which is mainly returned immediately to the same water body where it has been taken from. Subtracting the amount of cooling water, the water stress index would drop from 31% to (47 – 26)/ 154 ¼ 14%. In addition, 98% of the roughly 18 000 million m3 a1 of all industrial and municipal wastewaters are treated adequately and returned to the surface water. Together with the cooling water, these used waters amount to 44 000 out of 47 000 million m3 a1 which are still available after they have been used. Statistically they might be counted twice, once as withdrawal and a second time as leaving the country by rivers (Cornel and Meda, 2008a).
Table 2
Threshold values of two parameters used to characterize water stress situations Situation
Influence on water reuse
Based on per capita availability of renewable freshwater in m3(C a) 1 Water stress o1700 The region begins to experience water stress and the economy or human health may be harmed. Chronic water o1000 The region experiences frequent water scarcity and scarcity water supply problems, both short- and long-term. Absolute water o500 The region completes its water supply by desalting stress seawater, overexploiting aquifersa or performing unplanned water reuse. Minimum survival o100 Water supply for domestic and commercial uses is level compromised, since the total availability is not enough to fulfill demand for all uses (municipal, agricultural, and industrial). Based on water intensity use index (WIUI) or water stress index (WSI) Water stress 420% The region is experiencing severe water supply problems that are addressed by reusing water (planned or not), overexploiting aquifers (by 2–30 times), or desalting seawater.
Under these circumstances, developing a water reuse program is recommended. Reuse activities have to be put in place. Urgent planned water reuse measures need to be implemented. Under such circumstances the current economic development model is unsustainable.
Integral water management programs including planned water reuse and recycling are vital to the economy.
a
Although groundwater overexploitation has been observed even in countries with water availability over 4000 m3 (C a)1. From Jime´nez and Asano T (eds.) (2008) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA.
water availability per capita index and the WIUI index are presented in Tables 3 and 4, respectively. Note that the list of countries change according to the index used. No country from Oceania, North America, or South America is listed, although it is well known that for some of them, at a regional level, water problems do exist, for example, in Australia, United States (Florida, California, etc.), Great Britain (London), and Mexico City. Almost independent from a country’s water availability, water supply of megacities poses a special challenge as its
water demand by far exceeds the local supply. Long transport distances and/or overexploitation of groundwater reserves with, in many cases dramatic, ecological and economical consequences are common, as examples from all continents would verify. Intra-urban water reuse of adequately treated wastewater or wastewater side streams is a sustainable measure of integrated water resource management by which the demand for potable water can be reduced by up to 50%. In many countries, water reuse is already an indispensable necessity and common practice in water management. In
Wastewater as a Source of Energy, Nutrients, and Service Water Table 3
345
List of water-stressed countries according to the water availability per capita index
Minimum survival level
Absolute water stress
Chronic water scarcity
Water stress
o 100 m3 (C a)1
100–500 m3 (C a)1
500–1000 m3 (C a)1
1000–1700 m3 (C a)1
Libya Jordan Bahrain Yemen Israel Algeria Oman Tunisia
Egypt Morocco Cyprus
Lebanon Syria
Asia (excluding Middle East) Maldives
Singapore
-
Pakistan Korea, Republic India
Central America and Caribbean Bahamas
Barbados
S.Kitts and Nevis Antigua and Barbuda
Haiti
Europe -
Malta
-
Denmark Czech Republic Poland
Sub-Saharan Africa -
Djibouti
Cape Verde Kenya Burkina Faso
Rwanda South Africa Malawi Eritrea Comoros Zimbabwe Ethiopia Burundi Lesotho
Middle East and North Africa Kuwait Gaza Strip United Arab Emirates Qatar Saudi Arabia
From Jime´nez and Asano (eds.) (2008). Water Reuse: An International Survey of Current Practice, Issues and Needs London: IWA
future, it will play an essential part in sustainable water resource management and will be one of the greatest challenges of the twenty-first century. In terms of reuse, treated wastewater meeting the respective demands according to its designated use has to be considered as valuable, usable, and locally available water resource. Thereby, water reuse has a share in reducing the discrepancy between continuously increasing water consumption and limited water resources (DWA, 2008). Depending on the respective boundary conditions, there are various reasons for water reuse (EPA, 2004): 1. Local or regional water shortage often occurs in arid and semi-arid regions, and also in metropolitan areas and megacities, where – nearly independent of the annual rainfall – the local demand is by far higher than the local or regional availability. Against this background, intraurban multiple use, such as greywater treatment and reuse as service water or use of treated wastewater for irrigation of parks, sports fields, and cemeteries is one possibility to reduce the specific freshwater consumption to such a level as needed for cooking, drinking, and personal hygiene. Another possibility is to use service water in those cases
2.
3.
4.
5.
where freshwater quality is not essential, thus reducing the wastewater load on the receiving waters at the same time. Water scarcity and droughts occur in arid and semi-arid regions in particular. Here, economic use of water is a must. This means, multiple water use with subsequent reduced quality standards and with – according to the reuse purpose – adjusted interim treatments. Protection of water resources, that is, within the frame of Integrated Water Resource Management Concepts freshwater extraction which does not exceed the water renewing rate. In many cases, water reuse and use of service water can represent an alternative to freshwater extraction. Economic factors can be a driving force for intensified water recycling, in particular in industrial water management, and also in hotel resorts and irrigation. In many countries, where fees have to be paid for drinking water supply as well as wastewater disposal, water reuse pays off twice. The possibility to use the water ingredients, for example, as fertilizer or the option of heat recovery from water can be of additional interest. Energy saving and minimization of the emission of greenhouse gases are new drivers, as local water treatment
346 Table 4
Wastewater as a Source of Energy, Nutrients, and Service Water List of water-stressed countries according to the water intensity use index
1000–3000%
500–1000%
100–500%
50–100%
20–50%
Saudi Arabia Libya Qatar
Yemen Oman Israel Jordan Iraq Syria
Tunisia Algeria Iran
Morocco Afghanistan Lebanon Cyprus
Asia (excluding Middle East) Turkmenistan
-
Uzbekistan Azerbaijan Pakistan
Bangladesh India Japan
Kazahkstan Armenia Korea, Republic Sri Lanka China Thailand Singapore
Central America and Caribbean -
-
Barbados
-
Cuba
Europe -
-
Malta Hungary Moldova, Rep
Belgium Netherlands Romania Ukraine
Bulgaria Spain Germany Poland Italy Denmark France Portugal
Sub-Saharan Africa -
-
Mauritania Sudan
Niger Somalia
Mauritius South Africa Swaziland Zimbabwe Eritrea
Middle East and North Africa United Arab Emirates
From Jime´nez B and Asano T (eds.) (2008). Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA.
and reuse can be by far more efficient in terms of energy savings than long-distance transport of freshwater, its potential intensive treatment, and subsequent wastewater treatment. 6. Political reasons, such as the independency of water supply from neighboring countries, which can be a strong motivation for multiple water use. Besides the above mentioned reasons, water shortage can result from extensive, often subsidized water use for agricultural irrigation. In combination with the export of agricultural products from water-scarce countries (e.g., Israel and Spain), large amounts of water are indirectly exported as so-called virtual water aggravating water shortage. All these reasons lead to the increasing importance of wastewater as resource whose use is in ideal accordance with the idea of sustainability. Water reuse should therefore be an indispensable part of the respective Integrated Water Resource Management system. Figure 3 shows the per capita reuse rate in m3 (C a)1 and the percentage of reuse water in relation to the total water extraction (data from Earthtrends (2009)). The illustration
clearly shows that already today water reuse is given high priority in most of the world’s arid and semi-arid countries.
4.12.3 Origin and Amounts of Resources 4.12.3.1 Energy In order to minimize the consumption of energy and the emission of greenhouse gases, one has to consider a large number of individual segments of water supply and disposal as far as the detailed discussion of individual techniques. It starts with water supply and disposal, as pointed out in Section 4.12.2. In Figure 4, the values as listed in Table 1 (cf. Section 4.12.2) are presented graphically, supplemented by data for the energy intensity of recycled water treatment and distribution from the same report. It has to be mentioned again that the shown data might be higher than average since they are specific for the situation in California, USA, a region with high overall energy consumption and locally severe water scarcity. Nevertheless, they give a good impression of the relative impact of the different segments of the water supply and disposal chain on the total energy intensity.
Wastewater as a Source of Energy, Nutrients, and Service Water
347
70 Reuse per capita
Reuse/extraction
m3 (c.a)−1, (%)
60 50 40 30 20 10 Singapore
Malta
Jordan
Tunisia
Chile
Syria
Bahrain
Saudi Arabia
Cyprus
UAE
Mexico
Kuwait
Israel
Qatar
0
Figure 3 Reuse rate per capita in m3 (C a)1 and percentage of reuse water in relation to the total water extraction. Data from Earthtrends (2009) Water Resources and Freshwater Ecosystems. http://earthtrends.wri.org/searchable_db/index.php?theme ¼ 2.
Source
Water supply and conveyance 0−1.06 kWh m3
Water treatment
Water distribution
0.026−4.23 kWh m3
0.18−0.32 kWh m3
End use
Recycled water treatment
Recycled water distribution
Agricultural Residential Commercial Industrial
0.11–0.32 kWh m3 Wastewater discharge 0−0.11 kWh m3
Wastewater treatment
Wastewater collection
0.29−1.22 kWh m3
Source
Total water use cycle energy intensity (without end use energy demand) 0.53−5.3 kWh m3
Figure 4 Water use cycle and energy intensities for California. Numbers according to California Energy Commission (2005), p. 144. Adapted from Reiter P (2008) Reducing the Water Utility’s Footprint Through Utility Sponsored End-Use Efficiency. Speech at the IWA World Water Congress Vienna 2008, 8–12 September 2008.
For the further treatment of treated wastewater to be used as service water and its distribution, the energy report quotes 0.1–0.3 kW h m3. This is a fraction of the energy intensities for water supply, treatment, and distribution, and points to the fact that water reuse not only preserves water resources, but may also be reasonable under the aspects of energy consumption and the emission of greenhouse gases. Although the range of the values is rather large and will vary according to the boundary conditions, particularly in the segments supply and conveyance and treatment, the following facts can be deduced:
•
Transport of water and wastewater as well as water distribution and wastewater collection present a significant
• •
factor within the energy balance. This favors small-scale treatment and reuse of recycled water. The use of recycled water from treated wastewater is much more energy-efficient than desalination of brackish water or saltwater (0.11–0.32 kW h m3 compared to 2–4 kW h m3). The partial replacement of potable water by reclaimed water for nonpotable use does not only conserve water resources but can also be a significant contribution to the reduction of the overall energy consumption.
This demonstrates the close link between water consumption, water reuse, and energy. However, as boundary conditions and specific energy consumption can vary significantly, there cannot be general recommendations for water and energy
348
Wastewater as a Source of Energy, Nutrients, and Service Water
resource management, and thus for water reuse; all the more, as the energy demand for the generation of reuse water strongly depends on the raw water quality and the treated wastewater side stream, respectively, as well as the required standards for the reuse water. Bieker et al. (2009) show that for the complete treatment of greywater 0.6–1.2 kW h m3 can be estimated for reaching nonpotable reuse water quality for intra-urban use (disinfection not included). The lower value of 0.6 kW h m3 when using the conventional activated sludge process or biological aerated filters, the upper value of 1.2 kW h m3 when using membrane bioreactors for biological treatment. For disinfection, for example, with ozonation, UV irradiation or membrane filtration, additional 0.035–0.4 kW h m3 must be calculated (Haberkern et al., 2008). In case an additional desalination step is used downstream, for example, in order to use the recycled water for groundwater replenishment or for direct or indirect potable reuse, the energy demand has to be assessed at the top end. Keller (2008) states an energy demand of 0.9–1.2 kW h m3 alone for the further treatment of already biologically treated wastewater to become so-called purified recycled water, via microfiltration, reverse osmosis, UV/H2O2 disinfection, and chlorination. The reuse water was treated to meet the highest standard of drinking quality through a seven-barrier treatment system (Keller, 2008) (see also Western Corridor, 2009). The operator of the new Goreangab water reclamation plant in Windhoek, Namibia, which, since 2002, daily produces approximately 21000 m3 drinking water from a mixture of secondary effluent and reservoir water, quotes an energy
demand of 1.34 kW h m3. This amount includes the whole multibarrier process chain, consisting of powdered activated carbon, preozonation, coagulation, flocculation, dissolved air flotation, dual media filtration, ozonation, biological activated carbon, two-stage granular activated carbon, ultrafiltration, chlorination, and stabilization including the raw water and high-lift pumps. Of course, these figures only reflect the energy for operation and cannot replace thorough life cycle assessments. The examples point out the quality-dependent range of the energy demand and place special emphasis on the question of required quality. Thus, consequent energy management also means to supply the quality required for the respective use, that is, fit for purpose, water for garden irrigation, toilet flushing, or street cleaning does not necessarily be of potable water quality. The above-mentioned data for the energy consumption per m3 water are important parameters which can be compared transnationally only in combination with the per capita water use. This mainly applies to concentration-dependent energy demand values, such as wastewater treatment. Figure 5 illustrates the large range of the municipal water withdrawal per capita, based on data published by the Food and Agriculture Organization (FAO, 2009). (The municipal water withdrawal considers the quantity of water withdrawn primarily for the direct use by the population and is usually computed as the total water withdrawn by the public distribution network. It might include public services, private gardening, small enterprises, and that part of the industries, which is connected to the municipal network and includes
700 FAO 2009
600
Eurostat 2009
l (C·d)−1
500
400
300
200
100
Bahrain Qatar USA Australia Kuwait United Arab Emirates Italy Japan Sweden Spain Greece Belgium-Luxembourg Norway France Israel Portugal Czech Republic Malta Austria Saudi Arabia Switzerland Denmark Finland Poland Jordan India Germany Namibia Singapore United Kingdom China Netherlands Eritrea Central African Republic Cambodia Chad
0
Figure 5 Municipal water withdrawal per capita and day according to FAO (2009) and Eurostat (2009).
Wastewater as a Source of Energy, Nutrients, and Service Water
water losses (FAO, 2009).) One can clearly see that water withdrawal does not depend on water availability alone. All countries with municipal water withdrawal above 400 l (C d)1 can be called rich, even though some of them belong to the arid or semi-arid countries. The data depict a comparison of the countries; however, it should be pointed out that for some countries clearly divergent data are reported by different sources. As to these variations, data from Eurostat, the database of the European Commission, for the ‘‘total water abstraction by public water supply’’, which in some cases deviate significantly, have been added to the figure. Since the water withdrawal for the municipal network may contain transport losses, supply for industry and commerce, public buildings, etc., data for ‘‘water consumption for private households’’ for some European countries according to Eurostat (2009) are additionally shown in Table 5. In comparison to the municipal water withdrawal, the per capita water use in
Table 5
According to FSO (2009). According to Asano (2007). n.a.: not available. b
households is considerably lower. For example, in Germany, the municipal water withdrawal is 174 l (C d)1, whereof 122 l (C d)1 are used in private households (FSO, 2009), and Asano (2007) quotes a specific water consumption for the USA of 274 l (C d)1 against a water withdrawal of 573 l (C d)1. Specific withdrawal factors range from 12 to 14 l (C d)1 in Chad, Cambodia, and the Central African Republic, to 100– 300 l (C d)1 in many European countries. Up to 600 l (C d)1 are used in Qatar and Bahrain, two of the driest countries in the world, where a large percentage of water is generated by energy-intensive seawater desalination. Thus, the energy demand for the processing of drinking water for Qatar and Bahrain is estimated as 600 l (C d)1 4 kW h1 m3 365 d a1/1000 l m3 ¼ 876 kW h (C a)1. This value compares to 27.5 kW h (C a)1 in Germany (193 l (C d)1 0.39 kW h m3 365 d a1/1000 l m3 for drinking water supply, conveyance, and treatment).
Municipal water withdrawal per capita in some countries according to different sources
Chad Cambodia Central African Republic Eritrea Netherlands China United Kingdom Singapore Namibia India Jordan Poland Finland Germany Denmark Switzerland Saudi Arabia Austria Malta Czech Republic Portugal France Israel Belgium-Luxembourg Norway Greece Spain Sweden Japan Italy United Arab Emirates Kuwait Australia USA Qatar Bahrain a
349
Municipal water withdrawal FAO (2009) (l (C d)1)
Municipal water withdrawal Eurostat (2009) (l (C d)1)
Water consumption for private households Eurostat (2009) (l (C d)1)
12 12.4 13.7 18.1 83.3 87.7 95.6 101 103 132 139 150 179 193 209 232 241 248 277 282 285 288 288 301 301 315 318 334 373 381 397 449 493 573 581 660
n.a. n.a. n.a. n.a. 215 n.a. 339 n.a. n.a. n.a. n.a. 150 211 174a 214 360 n.a. n.a. 94 187 253 255 n.a. 188 487 207 357 268 n.a. n.a. n.a. n.a. n.a. n.a. n.a. n.a.
n.a. n.a. n.a. n.a. 122 n.a. n.a. n.a. n.a. n.a. n.a. 86 n.a. 122a 123 229 n.a. 121 75 91 151 n.a. n.a. 104 213 97 164 144 n.a. 203 n.a. n.a. n.a. 274b n.a. n.a.
350
Wastewater as a Source of Energy, Nutrients, and Service Water
Considerable differences in the per capita energy consumption arise from wastewater treatment as well. For example, when taking the per capita consumption of 274 l (C d)1 for USA and 122 l (C d)1 for Germany as a basis, annual wastewater volumes of 100 and 44.5 m3 (C a)1 result, respectively. Combined with the specific energy consumption of 0.3–1.22 kW h m3 (see Table 1), the energy needed for wastewater treatment for USA is estimated to be 30–122 kW h (C a)1. Thereby, wastewater treatment plants with nutrient elimination will rather be in the upper level. In Germany, with an average water consumption of 122 l (C d)1, the average annual energy demand for wastewater treatment (without sewer system) is 34.9 kW h (C a)1 (0.44 kW h m3) with a range of approximately 25 kW h (C a)1 to more than 80 kW h (C a)1 (Keicher et al., 2008). Such numbers should only be compared with great caution because of the specific boundary conditions, such as separate or combined sewer system, vacuum, pressurized or gravity flow sewer, topography, effluent quality standards, treatment level of the wastewater treatment plant, sludge treatment, off gas treatment, etc. All the same, it is obvious that the specific daily water and wastewater amount is decisive for the energy consumption and only the comparison of per capita values is appropriate and assessable. Energy and water are linked inseparably. Reducing the water consumption generally leads to a reduction of the energy consumption. However, here as well, there are exceptions, and one should examine and evaluate every single case. The second aspect when considering energy and water is: How much energy does wastewater contain? Here, different forms of energy have to be taken into account. It has to be distinguished between 1. the potential energy in dependence of the geodetic height which can be of importance at least in high-rise buildings or in topographies with large differences in altitude; 2. the thermal energy which in particular was input for hot water generation at the consumer’s end; and 3. the chemically bound energy which is mainly stored in organic water ingredients.
•
15 K1.16 W h (l K)1 ¼ 254 kW h (C a)1. This is by far higher than the potential energy estimated above. In other words, the amount of energy which is released to the environment when 1 l of water cools by 1 K corresponds to the calculated potential energy of 1 l of water retained at an altitude of 426 m. The chemically bound energy can be defined by the COD. The maximum potential amount of methane per kilogram COD can be calculated stoichiometrically. CH4 þ 2O2 ¼ CO2 þ 2H2O, that is, the COD per mol methane is 64 g O2. With a molar volume under standard conditions (0 1C, 1 atm) of 22.41 l, the CH4 equivalent of COD converted under anaerobic conditions is 22.41/64 ¼ 0.35 l methane g1 COD ¼ 350 l methane kg1 COD. Considering the energy potential of methane, 802 kJ mol1, the result is 802 kJ mol1 15.625 mol methane/kg COD ¼ 12.53 MJ kg1 COD, that is, 3.48 kW h kg1 COD. Based on a daily COD load per capita of 110–120 g, the maximum theoretical energy content is 139–152 kW h (C a)1, in case the entire COD could be transferred to methane and could be utilized.
Table 6 summarizes the estimates, independent from the fact that the total energy content does not allow any conclusion on the reclaimable and technically usable percentage. The comparison shows that the major part of the energy contained in wastewater is stored as thermal energy and is reclaimable as close to the source as possible, whereas the percentage of chemically bound energy is smaller; however, it can be transported in the wastewater via the sewer system almost without losses. The recovery of potential energy from wastewater seems promising only in high-rise buildings or respective topographies. Under normal conditions, the potential energy is much smaller than thermal or chemically bound energy by several orders of magnitudes. Annotation: It might be promising, with high-rise buildings or hillside locations, to treat wastewater from upper floors or higher altitudes, respectively, to be used as service water in the lower floors/altitudes and thereby minimize costs for pumping freshwater. Whereby one can assume that the energy needed for pumping, in consideration of losses in the pipes
The following estimation will show the order of magnitudes:
•
•
When neglecting frictional losses etc., the potential energy is directly proportional to the height. Considering a medium water consumption of approximately 122 l (C d)1 as in Germany, the energy content of wastewater (Epot ¼ m g h) in a height of 50 m is 122 kg (C d)1 9.81 ms2 50 m ¼ 59 841 kg m2 s2 ¼ 59 841 Ws ¼ 6.1 kW h (C a)1. The thermal energy stored in wastewater mainly results from hot water generation and therefore basically affects greywater, that is, wastewater from showers, baths, washing machines, and possibly the kitchen. The maximum recoverable thermal energy is calculated via the specific heat capacity of water, the temperature gradient, and the water quantity according to the equation Etherm ¼ cp DT m. The specific heat capacity of water is 4.18 kJ (kg K)1. This means, the amount of heat stored in water per liter and kelvin is 4.18 kJ ¼ 1.16 W h. Considering, for example, a greywater volume of 40 l (C d)1 and an available DT of 15 1C, the thermal energy amounts to 40 l (C d)1
Table 6 Estimated energy content of wastewater as potential, thermal, and chemically bound energy Calculated energy kW h (C a)1
Potential energy (122 l (C d)1, 50 m height) Thermal energy 40 l (C d)1, DT ¼ 15 1C (or 120 l (C d)1 with DT ¼ 5 1C) Chemically bound energy 110–120 g COD (C d)1 a
6.1
Energy equivalent for driving a car for how many meters/day?a 9
254
1000
139–152
523–571
7 l (100 km)1 (34 mpg); 10 kW h l1 fuel.
Wastewater as a Source of Energy, Nutrients, and Service Water
and the efficiency of pumps and motors, is more than twice as much as the potential energy contained in the pumped water.
4.12.3.2 Nutrients In Table 7, an overview is given on the yearly loads of nutrients and the wastewater streams they mainly occur in, using the example of Germany (Otterpohl, 2002). As can be seen from the data in the table, nitrogen mainly occurs in urine as urea, in a daily quantity of 11–13 g per capita. With conventional wastewater disposal and treatment using the activated sludge process, nitrogen is physiologically bound in the activated sludge to approximately 20–30%. Depending on the process, the remaining part stays in the wastewater as ammonium ion, is nitrified to nitrate with an energy input of 2.15 kW h kg1 N ( ¼ B6 kW h (C a)1), or Table 7 Annual loads of N, P, K, and COD and their distribution among the wastewater side streams Greywater
Urine
Feces
Volume Compound
m3 (C a)1 kg (C a)1
25–100
B 0.5
B 0.05
N P K COD
B4–5 0.75 1.8 30
B3% B10% B34% B41%
B87% B50% B54% B12%
B10% B40% B12% B47%
From Otterpohl R (2002) Options for alternative types of sewerage and treatment systems directed to improvement of the overall performance. Water Science and Technology 45(3): 149–158.
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nitrified/denitrified which requires approximately 1.5– 2.9 kW h kg1 N ( ¼ B3.5–6.9 kW h1 (C a)1, see Box 2). A utilization of nitrogen as fertilizer might be realized by the use of nitrogen-containing reclaimed water for agricultural irrigation and/or by using sewage sludge as biosolids on cropland. With the latter, only that share of nitrogen, which is bound in biomass, can be utilized. In Germany, with typical wastewater treatment processes including nitrification/denitrification with nutrient elimination, only a small part of the nitrogen is returned with the socalled biosolids to the nutrient cycle. Dockhorn calculated 48 000 Mg-N a1 out of 507 000 Mg-N a1 in the influent of the wastewater treatment plant, that is, a percentage of just below 10% is made use of (Dockhorn, 2007: 14). Phosphorus. Based on an average phosphorus load of 1.8 g P (C d)1 (ATV-DVWK A 131, 2000) in the raw wastewater, with German boundary conditions and a per capita wastewater flow of 200 l (C d)1 the influent concentration is around 9 mg l1. An average of approximately 11% of the incoming phosphorus load is removed with the primary sludge during primary settlement (ATV-DVWK A 131, 2000). In biological wastewater treatment, approximately 28% of the incoming phosphorus load is incorporated into the biomass and removed with the surplus sludge, even without specific phosphorus removal processes. Based on the permitted discharge concentrations of 1 or 2 mg l1, respectively, approximately another 50% of the incoming phosphorus load has to be removed specifically, either by biological or by chemical–physical P removal processes or their combination. In summary, this means approximately 90% of the incoming phosphorus load is incorporated into the sewage sludge. In Figure 6, the phosphorus balance for a
Box 2 Energy consumption for nitrogen removal The energy demand for nitrification/nitrogen elimination is estimated as follows. For the oxidation of 1 g NH4–N, stoichiometrically 4.57 g oxygen is necessary. Since the nitrifying bacteria consume some nitrogen for biomass production, the net oxygen consumption for the elimination of 1 g NH4–N via nitrification is 4.3 g oxygen. Assuming an oxygen transfer efficiency of 2 kg O2/kW h, the resulting energy input for aeration is 2.15 kW h kg1 N solely for nitrification (6.0 kW h (C a)1, assuming an annual nitrified quantity of approximately 2.77 kg N (C a)1). In case nitrate is successively denitrified, a fraction of the oxygen input (stoichiometrically 2.9 g O2/g N) can be used (recovered) for the oxidation of a part of the wastewater’s organics. Assuming that 85% of the nitrate is denitrified – the rest of the nitrate is lost via the effluent – the specific oxygen demand amounts to 1.83 g O2 g1 N and the specific energy demand for aeration to 0.92 kW h kg1 N. For total nitrogen elimination (additional denitrification), the energy demand for recirculation and agitation of the anoxic reactor volume has to be added. For recirculation 0.49 kW h (C a)1 is estimated with the following assumptions: specific wastewater quantity: 200 l(C d)1; recirculation factor: 300%; delivery head of recirculation pump: 0.50 m; efficiency of recirculation pump: 60%. For agitation 0.84 kW h (C a)1 is estimated with the following assumptions. Specific volume of anoxic tank: 50 l/C; power density for agitation: 2 W m3. The total additional energy demand of 1.33 kW h (C a)1 referred to as the annual denitrified quantity of approximately 2.36 kg N (C a)1 results in 0.57 kW h kg1 N. Finally, the theoretical energy demand for nitrogen elimination results in 1.48 kW h kg1 N. Another estimation of the energy consumption for nitrogen elimination can be derived from real energy consumption data of wastewater treatment plants. Roth (1998) reports an energy consumption for aeration on wastewater treatment plants with nitrogen elimination of 14.2 kW h (C a)1. The fraction of aeration energy due to nitrogen elimination is estimated at 3.12 kW h (C a)1 (22%) considering the above-mentioned consumption for nitrification, the recovery from denitrification and the consumption for oxidation of carbon compounds (specific organic load: 16.4 kg BOD5 (C a)1, specific oxygen demand: 1.1 kg O2kg1 BOD5). For the energy consumption for recirculation and agitation of the anoxic volume Roth (1998) quotes to be 2.01 respectively 1.75 kW h (C a)1. The total energy consumption results in 6.88 kW h (C a)1. Referred to the annual quantity of nitrified/denitrified nitrogen of approximately 2.36 kg N (C a)1, the specific energy consumption results in 2.91 kW h kg1 N. It has to be remarked that the calculated energy requirements for nitrogen removal only apply to the denitrified nitrogen and not to the total quantity of nitrogen removed. Nitrogen removal by sludge production can be classified as a side effect of carbon removal. In fact, the energy demand to remove nitrogen contained in sludge needs to be calculated separately, depending on the specific sludge stabilization process, for example, sludge digestion with return of the supernatant to the biological treatment or separate treatment with for example, nitration/denitration, annamox, or deammonification.
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Wastewater as a Source of Energy, Nutrients, and Service Water
Influent
Effluent
EBPR and precipitation PS 1.8 g (C.d)−1 100%
0.2 g (C.d)−1 11%
SS 0.5 g (C.d)−1 28%
0.9 g (C.d)−1 0.2 g (C.d)−1 50 % 11%
Approximately 90% in sludge Figure 6 Phosphorus balance for a typical municipal wastewater treatment plant in Germany with biological phosphorus removal and/or precipitation. PS, primary sludge; SS, surplus sludge; EBPR, enhanced biological phosphorus removal.
typical German municipal wastewater treatment plant with phosphorus removal is illustrated schematically.
4.12.3.3 Water Water demand. How much water do we need/consume?
• • •
•
3
1
1–3 m (C a) is needed for drinking and cooking alone. Approximately 50 m3 (C a)1 is consumed in European private households (equates to approximately 140 l (C d)1). Approximately 230 m3 (C a)1 is the specific consumption of European private households, public services, and industrial activities (not including energy generation; for comparison: USA: 266 m3 (C a)1; Africa: 25 m3 (C a)1 (Zehnder, 2003)). 41700 m3 (C a)1 is the total consumption, including food production.
As shown in Figure 7, by far the largest part is used for food production. One kilogram of bread needs approximately 2 kg of wheat (dry weight of total plant) (Allan, 1997; Zehnder, 1997). In order to produce this quantity of plant material, the plants take up at least 1 m3 of water which they mainly release to the atmosphere as transpiration losses. The rule of thumb 1 m3 of water for 1 kg of bread holds true only in case of optimum conditions. In reality, the consumed water quantity is much higher. American farmers, for example, consume approximately 4 m3 of water per kilogram bread equivalent, and in the tropics approximately 5 m3 of water are used per 1 kg of rice, instead of the required 2 m3 (Zehnder, 2001). In order to generate foodstuff which is essential for providing a sufficient nutrition with 2500 kcal d1, presently 500– 1000 m3 (C a)1 with vegetarian diet and 1200–1500 m3 (C a)1 with a diet where 500 kcal out of the 2500 kcal are supplied by meat are needed (Zehnder, 2001, 2003). The higher demand for the nutrition of nonvegetarians derives from an approximately 10 times higher water demand per produced energy unit (kcal) for meat compared with vegetarian food. Even with optimal irrigation, that is, minimizing water losses, one has to calculate 250 m3 (C a)1 for vegetarians and 680 m3 (C a)1 for nonvegetarians (Zehnder, 2001, 2003).
1−3 m3 (C.a)−1 drinking and cooking
ca 50 m3 (C.a)−1 for private households
ca 180 m3 (C.a)−1 for industry and public services 800−1400 m3 (C.a)−1 for agriculture for foodstuffs production
Figure 7 Annual water demand (Europe) (Cornel and Meda, 2008a).
Against this background, it seems that the recovery of water from private households can only make a small contribution toward the reduction of water shortage. However, looking at the drinking water management of large cities, one gets a completely different impression. Beijing, London, Los Angeles, Mexico City, Singapore, Tehran, and Tokyo, cities of different development status and different locality, all have in common that the local drinking water demand by far exceeds the locally available resources. Drinking water supply can only be assured with large efforts and occasionally with severe impacts on the environment. Excessive exploitation of existing water resources, lowering of the groundwater level, energyintensive and costly transport of water over many hundreds of miles, and energy-intensive desalination of seawater are only few of the consequences of the current water supply of the megacities’ population. Moreover, most of the water is only used for the transport of pollutants. The demand of high-quality, potable water could be significantly reduced by reusing reclaimed water. One possibility would be intra-urban use of treated greywater for toilet flushing and other uses where potable water quality is not required as will be shown in Section 4.12.6.1.2. Another possibility is the direct or indirect reuse of reclaimed water as potable water, as practiced, for example, in Singapore, Windhoek, Australia, and California. Highly treated reclaimed water is introduced either directly into the potable water supply up- or downstream of the water treatment plant or into a raw water supply such as potable water storage reservoirs or groundwater aquifers (Asano, 2007: 6). Accordingly, the quality requirements are higher than with nonpotable use, and the required energy for treatment is higher as well. On the other hand, mass restrictions,
Wastewater as a Source of Energy, Nutrients, and Service Water
unavoidable with greywater use, fall away, and a dual supply system, coupled with the potential risk of cross-connection (as exists in buildings with parallel use of potable and service water) is avoided. ‘‘Water reuse is particularly attractive in the situation where available water supply is already overcommitted and cannot meet expanding water demands in a growing community. Increasingly, society no longer has the luxury of using water only once’’ (Asano, 2007). One might add that water reuse saves not only valuable water resources, but might also reduce the energy demand of the water cycle and contributes to reduce the emission of greenhouse gases. However, in places where water is sufficiently available and can be supplied with low energy input, water recycling might be relevant only in case a significant cost reduction can be achieved. In such cases, water supply does not reduce water availability.
Heating circuit
353
User
35 °C
50 °C Condenser Expansion valve
Compressor
Heat pump
Evaporator 12 °C
6 °C Sewer c. 15 °C
4.12.4 Energy 4.12.4.1 Caloric Heat One possibility for energy recovery from wastewater is heat recovery. Domestic wastewater presents a permanently available, year-round heat source with comparably high temperatures, in particular with separate sewer systems and low contents of sewer infiltration water. The heat contained in wastewater can be utilized via heat exchangers and heat pumps. Heat exchangers installed in the sewer transfer the heat from the wastewater to a heat exchanger fluid. Subsequently, in a second heat exchanger, the so-called heat pump evaporator, the heat is transferred to a working fluid with a lower boiling point. As a result of the energy input, the refrigerant evaporates and is then compressed by using electric energy. In doing so, the temperature is increased to a usable level. In a third heat exchanger, the so-called condenser, the vapor releases its heat to the heating circuit. Thereby, the pressurized refrigerant liquefies again. After the expansion and cooling of the refrigerant within the expansion valve of the heat pump, the refrigerant cycle starts anew (Figure 8). With such installations in the sewer system, the cooling normally amounts to 2–3 1C. For example, with 45 m3 (C a)1 and a heat capacity of 1.163 W h (l K)1 approximately 105–157 kW h (C a)1 is produced. To assess the reasonability of this method of heat recovery, it is important to consider the ratio between the useable heat capacity and the added electric power. This characteristic is called coefficient of performance (COP) and is calculated via the heat pump efficiency Zhp and – derived from the second law of thermodynamics – the maximum COP (reciprocal value of the Carnot efficiency):
COP ¼ Zhp ðThot =ðThot Tcold ÞÞ When taking into account friction, pressure losses, temperature gradients during heat transfers, and losses during compression, one can assume a heat pump efficiency factor of approximately 0.45–0.5. In Table 8, COP values for cold water temperatures of 12 and 25 1C, respectively, and two hot water temperatures, that is, 40 1C for a low-temperature heating
Heat exchanger Figure 8 Schematic diagram of a heat pump.
Table 8 Coefficient of performance for heat pumps under different operating conditions Tcold / Thot
12 1 C
25 1 C
40 1C 65 1C
5.6 3.2
10.4 4.2
circuit and 65 1C for a hot water boiler (nominal temperature 460 1C, because of Legionella) are compared to each other. Assuming an efficiency factor of 50%, COP values of 5.6 and 3.2, respectively, for cold water temperatures of 12 1C, and 10.4 and 4.2, respectively, for cold water temperatures of 25 1C are the result. A COP of 3.2, for instance, indicates that 3.2 times more heat is made available than electric energy has to be provided for the heat pump. This is a significant advantage compared to electric warm water heating. However, taking into account the power plant efficiency in power generation and transmission losses in the power grid, assuming a total efficiency factor of approximately 30–35% for power generation and transport/distribution, just the same amount of heat energy is released as in case of direct local use of primary energy for heating (3.2 0.3 ¼ 98%). When also considering that due to fluctuations in temperature, biofilm formation on heat exchangers, etc., the actual efficiency (annual average value) is lower than the COP value indicates, the potentials and limits of heat recovery from cold wastewater become apparent. However, the table also shows that the COP is significantly higher with wastewater (greywater) of 25 1C. This means, in order to achieve optimum energy recovery, the process has to take place as close as possible to the origin of the hot water, for
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Wastewater as a Source of Energy, Nutrients, and Service Water
example, via heat recovery from greywater close to its source. Therefore, efficient heat recovery also asks for preferably decentralized use of the generated heat, thus minimizing energy losses from the generated hot water on its way to the consumer. In Section 4.12.7, the potential of heat recovery is illustrated with the example of a semicentralized supply and disposal center. Via heat exchanger and heat pump, the caloric heat of the (treated) greywater, which is available as service water for toilet flushing, is utilized. This method pays off twice. Heat recovery is efficient as the temperature of greywater is comparably high, and an additional positive effect is that with cooling the service water, the potential of microbial recontamination is reduced significantly. If required, heat pumps can also be used for supplying cooling energy instead of heat. In this case, heat is added to the wastewater instead of taken from it. However, with equal temperature gradients, heat pumps are more efficient for heating than for cooling. The reason is that the waste heat from the compressor can be used in the heat mode but not in the cooling mode.
4.12.4.2 Degradable Organic Constituents As described in Section 4.12.3.1, the chemically bound energy is directly proportional to the COD. Stoichiometrically, 1 kg COD equals 350 l methane (under standard conditions), which corresponds to a net calorific value of 3.48 kW h. Accordingly, 0.39–0.42 kW h are chemically bound in the daily wastewater load of approximately 110–120 g COD/C, amounting to 139–152 kW h annually. In Germany, the specific energy consumption of wastewater treatment plants using the activated sludge process with nutrient removal and sludge dewatering is – depending on the plant size – between 25 and 35 kW h (C a)1 for plants 4100 000 PE (population equivalent) and 55–80 kW h (C a)1 for plants o5000 PE (Keicher et al., 2008). The manual Energie in Kla¨ranlagen (energy in wastewater treatment plants) quotes a guideline value of 23–30 kW h (C a)1 for the energy consumption of wastewater treatment plants with C and N removal and a plant size 430 000 PE and an optimum value of 20–26 kW h (C a)1 (MURL NRW, 1999). Comparing the estimated energy content of wastewater, that is, 139–152 kW h (C a)1, with the specific energy
Table 9
consumption, that is, 20–30 kW h (C a)1, the assumption seems likely that wastewater treatment plants do not need to be energy consumers, but rather net producers of renewable energy, even with aerobic activated sludge plants; as, for instance, Shizas and Bagley (2004) conclude. Looking closer, one will note that though the total energy content is an interesting upper limit value, it does not reveal the percentage that can be transferred into electricity or usable heat. These values depend on the requirements for the discharge quality as well as on the treatment technology itself. The wide range of standards and the large number of different processes make comparative evaluations difficult. In the following assessment, only wastewater treatment plants with C and N and P removal – as it is the standard treatment in the European Community for sensitive areas – are compared. In Western Europe, USA, and Japan, the activated sludge process is the most commonly used technique for municipal wastewater treatment. Thereby, the organic carbon is oxidized to CO2 and bound in the produced biomass by approximately 50% each. Nitrogen, bound as ammonium, is first oxidized to nitrate and then reduced to N2 (nitrification/denitrification). Parts of the nitrogen as well as phosphorus are bound in biomass. The remaining phosphate is removed either biologically or via precipitation/flocculation. In the review below, it is presumed that primary sludge as well as excess sludge is stabilized anaerobically and that the produced biogas is used to produce energy in a block-type thermal power station or micro-turbine. What is the energy balance for such a standard configuration? The hydraulic retention time (HRT) in the primary sedimentation as well as the sludge retention time (SRT) as a function of temperature in the activated sludge process are the fundamental parameters for oxygen demand and thus energy consumption in the activated sludge process on the one hand and the amount of biogas and thus the producible electricity on the other hand. In Table 9, the energy demand for oxygen supply in the biological treatment and the electric energy generation from biogas for a model WWTP with 100 000 PE and for a design temperature of 12 1C are estimated, with varying SRT and HRT for primary sedimentation. In Figure 9, the values are depicted graphically. As expected, from an energy point of view, it is favorable to remove as much biomass as possible with the
Energy demand for oxygen supply and electric energy generation from biogas for different SRT and HRT for primary sedimentation
HRT primary sedimentation in h
SRT in d
Biogas l (C d)1
O2 demand g (C d)1
Energy demand for O2-supplya kW h (C a)1
Net calorific valueb (total in biogas) kW h (C a)1
Electric energy from biogasc kW h (C a)1
0 1 2 1 1
13 13 13 4 25
10.2 18.2 21.0 20.4 17.2
86 69 63 40 74
15.6 12.5 11.6 7.3 13.5
24.4 43.6 50.3 48.8 41.1
8.1 14.4 16.6 16.1 13.6
a
Oxygen transfer efficiency under process conditions: 2 kg O2 kWh1. Methane content in biogas: 66%. c Z ¼ 0.32 (32% electricity, 68% losses and heat). b
Oxygen demand g (C·d)−1
Oxygen demand g (C·d)−1
Biogas production l (C·d)−1
Biogas production l (C·d)−1
100
100
80
80
g (C·d)−1, l (C·d)−1
g (C·d)−1, l (C·d)−1
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60 40 20
355
60 40 20 0
0 Primary Primary Without sedimentation sedimentation primary 2h 1h sedimentation
SRT=4 d*
SRT=13 d**
SRT=25 d**
Primary sedimentation with HRT = 1h, *: Onlycarbonaceous elimination **: Nitrificationand denitrification
Nitrification and denitrification, SRT = 13 d
Figure 9 Influence of SRT and HRT in primary sedimentation on oxygen demand and biogas production for a model WWTP with 100 000 PE and for a design temperature of 12 1C. Modified from MURL NRW (1999) Handbuch Energie in Kla¨ranlagen. Du¨sseldorf (Germany): Ministerium fu¨r Umwelt, Raumordnung und Landwirtschaft des Landes Nord-Rhein Westfalen.
primary sludge and treat it anaerobically rather than oxidize the biomass with the need of oxygen input. Comparing the values of the last column of Table 9 with the energy demand for aeration of the aerobic tank in the activated sludge process which is approximately 28–30 kW h (C a)1 (Bo¨cker and Dichtl, 2001), it becomes clear that it is not possible to run an energy-autarkic operation in these plants with existent technologies and efficiency rates. However, a series of measures for minimizing the energy demand can be derived from these data, which have to be proved case by case:
•
• • •
•
Maximizing of the solids removal prior to the activated sludge process - increase of the biogas production - minimizing/reduction of the oxygen demand (annotation: when using machines, for example, micro-sieves, their energy demand has to be considered). Shortening of the SRT, for example, by two-step process management - maximizing the sludge production, minimizing the oxygen demand. Implementation of energetically more efficient processes, for example, trickling filters, rotating biological contactor, biodisk, etc., where possible. Improvement of the energy efficiency rate for converting biogas into electricity. The electrical efficiency is currently 26% for block-type thermal power station (Schro¨der and Schrenk, 2008), and could be increased by using modern equipment with efficiency rates up to 39% at the optimum operating point, or, in the future, using fuel cells with expected efficiency rates up to 50% (Schro¨der, 2007). Increase of the biogas volume via enzymatic additives or via disintegration; it has to be proved on an individual basis that the potential increase of the energy output is not lower than the invested energy.
•
•
•
Increase of the biogas volume via cofermentation of organic waste, for example, the addition of kitchen, market, or restaurant wastes, contents of grease separators, etc. Strictly speaking, this is not the way to energy autarky of wastewater treatment plants, as additional external energy carriers are cotreated which are not part of the wastewater cycle (cf. Section 4.12.7). Reduction of the nitrogen amount. The energy demand for nitrification/denitrification can be assumed to be 1.5–2.9 kW h kg1 Nremoved, as shown in Box 2. With 11 g N (C d)1 and considering that the nitrogen bound in biomass is not nitrified/denitrified, for N removal alone approximately 3.5 to 6.9 kW h (C a)1 are needed. Local N elimination, for example, by not introducing urine to the wastewater, would result in significant energy saving. However, these energy savings have to be compared to the additional energy demand resulting from storage, disposal, treatment, transport, and use. This also applies to the direct use of urine as fertilizer, as costs for spreading and transport might be relevant due to low N concentrations in the fertilizer. Implementation of alternative N removal processes with lower energy demand, such as nitration/denitration, annamox, or deammonification. Using these methods to treat process water can reduce the internal return load and thereby contribute to energy savings.
Keeping in mind that the main task of a wastewater treatment plant is to treat wastewater, one has to make sure in each single case that measures toward increasing the biogas volume or saving energy do not deteriorate the discharge quality. For example, increased removal of organic substrate during primary sedimentation can lead to substrate shortage during denitrification; cofermentation can lead to an increase in the return load of the wastewater treatment plant and disintegration to an increase in hardly degradable COD, etc.
Wastewater as a Source of Energy, Nutrients, and Service Water
Furthermore, with all measures one should consider the efforts needed and prove whether there are other possibilities which can be realized with equal efforts and yet higher energy savings. The energy consumption of German wastewater treatment plants is estimated to be 4.2–4.4 109 kW h a1 (Schro¨der and Schrenk, 2008), while the total energy demand (2005) is 3.955 1012 kW h a1 (BMWi/BMU, 2006: 8). This means that the energy demand of wastewater treatment plants constitutes only 0.12% of the total annual energy demand in Germany (electricity, oil, gas, etc.), around 1% of the used electric power of 611 109 kW h a1 (BMWi/BMU, 2006: 51) and equals approximately 1/5 of the annual consumption of electricity which is caused by the standby mode of electrical equipment (22 109 kW h a1 in 2004 (UBA, 2008)). On the other hand, wastewater treatment plants are often the largest energy consumers of the municipalities. Energy savings in wastewater treatment plants should not be questioned by any means, as only via the sum of many small individual measures the overall goal of a general reduction of the energy consumption can be achieved. However, one should always be aware of dimensions and significance. Besides the aerobic process, there are anaerobic processes with biogas utilization. Anaerobic ponds and other anaerobic treatment methods, converting organic load to methane gas energy efficiently but emitting methane directly to the atmosphere, are not dealt with in this context, as the effect of methane as greenhouse gas is 25 times higher compared to CO2 and, therefore, in no way, can be called sustainable (IPCC, 2007: 141). To reach the same discharge quality with anaerobic treatment as in the discussed activated sludge process, the design has to consist of an anaerobic plant which can convert approximately 35–45% of the COD to methane (at the common wastewater temperature of maximal 15–25 1C, depending also on residence time and plant design (Urban, 2009)), and subsequent aerobic treatment to remove the remaining carbon compounds and for nitrification/denitrification. Due to the lower sludge production when using the anaerobic process compared to the aerobic biological treatment, the fixation of nitrogen and phosphorus is lower as well. When considering an anaerobic conversion of 40% of the daily COD load of 120 g COD (C d)1, which can be achieved only at a water temperature 420–25 1C as might be typical for example, in Brazil, South Africa, Thailand, and parts of China and USA (Ruhr-Universita¨t Bochum, 2005), a rough estimation of the potential methane production results in 48 g COD (C d)1 350 l CH4 kg1 COD ¼ 16.8 l (C d)1 methane, that is, approximately 25.5 l (C d)1 biogas. This is only slightly more than that produced in anaerobic sludge stabilization (see Table 9). However, this theoretical quantity cannot be achieved in practice, as on the one hand, sulfates in wastewater impair the production of methane and, on the other hand, part of the produced methane remains dissolved in the water phase and is stripped to the air until aerobic conditions are reached. This is doubly unfavorable, as this gas cannot be used for the conversion to electricity and a large part of the achieved CO2 savings of the anaerobic process is compensated again due to the higher greenhouse gas equivalent of methane.
COD of dissolved methane at 0.66 bar partial pressure Dissolved methane Dissolved methane at 0.66 bar partial pressure 100 80 mg l−1
356
60 40 20 0 0
5
10
15
20
25
30
35
40
Temperature (°C) Figure 10 Dissolved methane vs. temperature and COD of dissolved methane.
How much methane is dissolved? The solubility of methane in water depends on the partial pressure of methane in the gas phase and on the water temperature. Figure 10 shows the progression for the temperature range 10–35 1C and a partial pressure of 0.66 bar as can be assumed for biogas. The methane concentrations are 13–21 mg l1, that is, 52– 85 mg l1 COD is dissolved as methane. In other words, considering a wastewater with, for example, 400 mg l1 COD of which approximately 40% are reduced to methane anaerobically (Urban, 2009) at 20–25 1C, approximately 40% (64 mg l1 out of 160 mg l1) of the produced methane remains dissolved in water, cannot be used as biogas but is stripped in the subsequent aerobic plant and emitted to the atmosphere. Against this background, the usable methane volume of 16.8 l (C d)1 in the selected example is reduced to 10.1 l (C d)1 respectively 3679 l (C a)1, corresponding to a calorific value of 36.6 kW h (C a)1 and – under the assumption of a conversion efficiency rate of 32% – an electricity yield of 11.8 kW h (C a)1. This amount is by far not sufficient to aerobically decompose the residual carbon content (which still amounts to 60% of the raw wastewater load) and provide enough energy for nutrient removal. Here, one has to consider that the amount of nitrogen to be nitrified/denitrified and the phosphate load to be precipitated are higher in relation to the organic load than in raw wastewater, as – due to the lower amount of sludge in the anaerobic pre treatment – only 10% of the nutrients are bound in the biomass. Table 10 gives an example of COD removal, methane production, and electricity yield for anaerobic wastewater treatment for different temperatures, based on data from Urban (2009). However, the increased greenhouse gas emission caused by methane emitted to the atmosphere remains the actual obstacle and needs new technical solutions. Cakir and Stenstrom (2005) compare aerobic wastewater treatment, including sludge digestion, with anaerobic wastewater treatment at 20 1C. By including the energy balance and greenhouse gas emissions of CO2 and CH4, only with concentrations above 300–700 mg l1 BODu (corresponding to approximately 400–950 mg l1 COD, with BODu: ultimate
Wastewater as a Source of Energy, Nutrients, and Service Water
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Table 10
COD removal, methane production, and electricity yield for anaerobic wastewater treatment for different temperatures
T 1C
Dissolved methane mg l1
COD of dissolved methane mg l1
COD conversion a %
COD removed b mg l1
Produced methane c (total) l (C d)1
Produced methane (dissolved) l (C d)1
Produced methane (usable) l (C d)1
Electricity yield kW h (C a)1
15 20 25 30
19 17 15 14
76 68 61 56
27 35 44 56
109 139 177 226
11.4 14.6 18.6 23.7
7.9 7.1 6.4 5.9
3.5 7.4 12.1 17.8
4.0 8.7 14.1 20.7
a
Temperature dependency according to Urban (2009). COD of organics converted in methane (both dissolved and in biogas), based on COD concentration of 400 mg l1. c Based on 120 g COD (C d)1. b
biochemical oxygen demand ¼ 1.46 BOD5 (von Sperling and Chernicharo, 2005) and BOD5/COD ¼ 0.5, depending on the sludge residence time), anaerobic wastewater treatment emits less greenhouse gases than aerobic treatment. From their investigations, the authors conclude ‘‘a technology to recover dissolved methane would make anaerobic treatment favorable at nearly all influent strengths.’’ The topic of nutrient removal was not further discussed in the publication. Consequentially, only with concentrated wastewater with temperatures above 20 1C and efficient methane recovery and utilization, anaerobic pretreatment can be a satisfactory alternative to greenhouse gas emissions and energy consumption. Again, in the individual case and particularly considering the requirements for nutrient elimination, anaerobic/aerobic process combinations and energy-efficient aerobic technologies in combination with sludge digesters have to be compared. Only with favorable conditions, that is, high concentrations and small water volumes, an energy-autarkic operation combined with high discharge quality can be realized. Via codigestion of organic (kitchen) waste, this goal can be achieved even in plants with nutrient elimination.
4.12.5 Nutrients In wastewater, the nutrients nitrogen and phosphorus exist in the dissolved form, nitrogen mostly as ammonium, and to a small percentage in organic nitrogen compounds (e.g., proteins and urea), while phosphorus mainly exists as inorganic phosphate and to a small percentage as organically bound phosphorus. The main source of phosphorus and nitrogen are human excrements. The per capita loads are approximately 11–13 g N (C d)1 and 1.8–2 g P (C d)1, respectively. This equals a ratio of 6:1 per weight, and a molar ratio of B13:1 N:P. Nitrogen and phosphorus compounds in domestic wastewater are in excess of what is required for the growth of microorganisms in wastewater treatment plants. With common aerobic wastewater treatment and depending on the plant’s configuration and the sludge retention time, approximately 20–30% of the nitrogen and 30–40% of the phosphorus are bound in the excess sludge. In plants with enhanced biological phosphorus removal and/or phosphorus elimination by chemical precipitation, up to 95% of the phosphorus are bound in the sludge.
In the simplest case, the utilization of nitrogen and phosphorus as fertilizer is carried by using (treated) wastewater for agricultural irrigation. Thereby, nutrient concentration in irrigation water might be too high and need to be controlled in order to avoid over-fertilization of the soils, especially when reuse water is the sole water resource for irrigation. Cornel and Meda (2008a) show that with usual European nutrient concentrations of approximately 55 mg N l1 and approximately 7 mg P l1 in the discharge of wastewater treatment plants without nutrient elimination, the amount of wastewater used for the irrigation should be limited. Taking for example the cultivation of wheat, the total amount of water required by the plants is between 6000 and 10 000 m3 ha1. In the case of irrigation with treated wastewater, the amount of irrigation water has to be limited to 1000–3800 m3 ha1, in order to prevent the nutrient input exceeding the required amount of 60–210 kg N ha1 (Cornel and Meda, 2008a). Depending on the nutrient concentration in the wastewater and the evaporation rate, such estimations undergo considerable variations. If no dilution water such as rainwater, surface water, or others is available, a partial nutrient elimination might be necessary, which can be effectively realized, for example, by complete nitrification or nitrification/denitrification of a partial wastewater flow and subsequent blending with the remaining flow. Basically, the utilization of nutrients with irrigation water is limited to irrigation periods. The storage of nutrientrich water is problematic, because of the risk of heavy algae growth in open storage or due to clogging and quality problems in the case of storage in aquifers. Disinfection of ammonium-rich wastewater causes problems as well. This may lead to different treatment objectives during the year. While nutrients may remain in the treated water during irrigation periods (treatment limited to carbon removal), the water must be treated for storage or discharge during nonirrigation periods; thus, nutrient removal may be required in order to avoid harmful effects on the water reservoir, the underground storage or on the receiving surface water body (Cornel and Weber, 2004). WWTPs applying different operation modes during different seasons can fulfill these varying requirements. A process concept for wastewater treatment with variable operation modes for the seasonal production of nutrientrich irrigation water and nutrient-poor discharge water is proposed, for example, by BMBF (2009) and Meda and Cornel (2010).
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Wastewater as a Source of Energy, Nutrients, and Service Water
Additionally, in many cases wastewater availability and agricultural lands – especially in fast-growing urban areas – are wide apart from each other. This means that investment and operation costs as well as the energy input for transport will increase. In many cases, agricultural use of sewage sludge is practiced by which those nutrients bound in the sludge can be recycled. However, due to the potential contamination with heavy metals, trace organic constituents, etc., agricultural sewage sludge application is discussed controversially and is on the decline, at least in Western and Central Europe. In Germany, for example, the application of sewage sludge in agriculture decreased from 42% in 1995 to 31% in 2003 (Meda et al., 2007). Besides, controversial views exist about the plant availability of chemically precipitated phosphorus, particularly when iron is used for chemical P removal in the treatment plant. In this case, the phosphorus is of low plant availability and may even reduce the plant availability of already-existing phosphorus (Ro¨mer and Samie, 2002a, 2002b).
4.12.5.1 Nitrogen Recovery In his keynote speech at the Nutrient Recovery Conference in May 2009 in Vancouver, Dr. James Barnard stated ‘‘Nitrogen can be recovered from wastewater, but the cost of recovery far exceeds that of fixing nitrogen from the atmosphere’’ (Barnard, 2009). >Attempts were made to strip and recover ammonia at elevated pH values from the effluent and recover as ammonium sulfate, but the method was not economically viable (Barnard, 2009). There are also research reports about stripping ammonia from the so-called process water, that is, wastewater side streams with high concentrations of ammonium (41000 mg l1), as they occur during the dewatering of digested sludge. By stripping at high pH values, ammonia is transferred to the gaseous phase, and subsequently transformed to ammonium salts or ammonia water, that is, aqueous solution with an ammonia content of approximately 25% per weight (Kollbach and Gro¨mping, 1996). However, these processes rather serve the reduction of the ammonium return load than nitrogen recovery. Moreover, they are energy-intensive and need large amounts of chemicals. In practice, only single cases have been realized. Combined precipitation, together with phosphate as struvite (magnesium ammonium phosphate (MAP)) from such concentrated process waters seems more promising. Under appropriate conditions, phosphorus and nitrogen can be removed by adding magnesium salts and reclaimed as valuable products (see Section 4.12.5.2). However, here as well only a small percentage of nitrogen is removed as struvite since the molar ratio of N:P is 13:1 in wastewater (as mentioned above) compared to 1:1 in struvite (unless additional phosphorus is added in the required molar ratio.) Another method for control and recovery of nutrients is urine separation. Urine contains 70–90% of the nitrogen contained in wastewater and approximately 50% of the phosphorus besides some 50% of potassium (see Table 7, Section 4.12.3.2), whereas its volume is less than 1%. Urinediverting toilets can be either water flushed or dry, depending on economic and cultural boundary conditions. They separate
urine from feces, the latter being then collected separately. Urine is almost pathogen-free. During storage, the pH value increases due to hydrolysis, which again contributes to the disinfection of the product. By separating urine, wastewater treatment plants are relieved, and with completely separating urine, nitrification/denitrification steps can be omitted. However, practical experience shows that in reality only a fraction of the expected urine is collected separately, thus reducing the recovery rate considerably. Larsen and Lienert (2007) report a collection rate of 60–75% (demonstration project in Switzerland), and Genath (2009) only 30–40% (demonstration project in Berlin, Germany). Tilley et al. (2009) report a mean collection rate of 30% with a range of 10–75% for a community-based project on urine separation and nutrient recovery in Nepal. The use of urine as fertilizer is particularly attractive in those countries where distances between urine source and its place of application are short. This mainly applies to rural areas and those places where the cost for fertilizers is unaffordable for many people. In case of increased transport distances, upgrading the nutrient concentration should be considered, as the nitrogen concentration in urine is only approximately 1% per weight compared to 40% per weight in commercial fertilizers. Maurer et al. (2003) have calculated that the 10-fold upgrading of the urine concentration via vaporization is more energy-efficient (approximately 5 kW h (kg N)1) than producing ammonium fertilizers via the Haber–Bosch process. Generating MAP (struvite) from urine, however, seems to be more energy-intensive with regard to ammonium alone, but by including phosphorus, the energy balance becomes favorable again (Maurer et al., 2003). Yet, one has to consider that the stoichiometric imbalance between N and P in urine is even more distinctive than in wastewater. Last but not least, one of the main unknowns in conjunction with the use of urine as fertilizer is the fate of pharmaceuticals and endocrine disruptors. Ultimately, by assessing energy and cost balances in the individual case, while taking into account urine storage, transport, distance to the user, and techniques of fertilizer spreading, decision supports toward the usefulness of separate urine collection can be developed. Besides ecological and economical questions, the acceptance of the toilet users and farmers as well as hygienic manageability is decisive.
4.12.5.2 Phosphorus Phosphorus recovery from wastewater has enjoyed great interest over the last 20 years. Thereby, the focus of research and development of new technologies is, on the one hand, on precipitation/crystallization from the aqueous phase, whereby concentrated process water streams are favorable, and, on the other hand, on the recovery from sewage sludge and sludge ashes. As described in Section 4.12.3.2, with the latter, recovery potentials are particularly high.
4.12.5.2.1 Phosphorus recovery during wastewater treatment The implementation of phosphorus recovery during wastewater treatment allows for the separation of already dissolved
Wastewater as a Source of Energy, Nutrients, and Service Water
phosphorus, applying relatively basic technologies. Thereby, phosphorus-rich side streams or process water with phosphorus concentrations 450 mg l1 are economically feasible. One big advantage of phosphorus recovery during wastewater treatment is the possibility of combining it with phosphorus removal. Investigations of recent years showed that phosphorus recovery is particularly successful in combination with biological phosphorus removal in side streams (supernatant liquor of the anaerobic stabilization) or from process water during sludge treatment. The phosphorus-rich water is fed into a precipitation/crystallization tank, where phosphorus is removed as calcium phosphate or MAP (struvite) by adding calcium or magnesium salts and, where need be, seed crystals (cf. Figure 11).
4.12.5.2.2 Phosphorus recovery from sewage sludge – wet chemical technology The wet chemical treatment of sewage sludge involves that in a first step the phosphorus bound in the sewage sludge is dissolved by adding acid or base, in combination with temperature if necessary. Thereby, in most cases (heavy) metals are redissolved as well. After removal of the insoluble compounds, phosphorus can be separated from the phosphorusrich water, for example, via precipitation, ion exchange,
nanofiltration, or reactive liquid–liquid extraction (cf. Figure 12). The same technologies can be applied to recover phosphorus from sewage sludge ash. The advantage here is that by disintegrating the organic matter – including all organic pollutants – there is an enrichment of phosphorus and consequently higher phosphorus concentrations in the liquid phase occur. In contrast to sewage sludge, solid–liquid separation after alkaline or acidic treatment is significantly easier to realize due to the exclusively inorganic formation of the sewage sludge ash (Schaum, 2007).
4.12.5.2.3 Phosphorus recovery from sewage sludge – thermochemical technologies Through specific thermochemical treatment of sewage sludge ash, it is possible to remove heavy metals and, at the same time, improve the plant availability of phosphorus (cf. Rhenania process). Based on the thermochemical approach, ashes are exposed – under suitable conditions – to chlorine-containing substances, potassium chloride or magnesium chloride, and treated thermally. With temperatures 41000 1C, a large percentage of the heavy metals is turned into heavy metal chlorides which vaporize, thus removing them from the ashes (Kley et al., 2005). The heavy metals are captured at flue gas treatment.
2+ 2+ Ca , Mg , seed crystals Process water rich in phosohorus Treated water depleted in phosphorous
Calcium phosphate MAP − struvite Figure 11 Phosphorus recovery from the liquid phase during wastewater treatment (Cornel and Schaum, 2009).
Release of phosphorous/metals
Acid base Sludge ash Energy Residues
(a)
Release of phosphorous/metals
359
Precipitation/ crystallization liquid−liquid extraction ion-exchange nanofiltration
(b)
Residues
Calcium phosphate MAP − struvite
Separation of phosphorous
Figure 12 Phosphorus recovery from sewage sludge and sewage sludge ash – wet chemical technologies (Cornel and Schaum, 2009).
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Wastewater as a Source of Energy, Nutrients, and Service Water
4.12.5.2.4 Products from phosphorus recovery processes With few exceptions, most of the processes include phosphorus separation by precipitation/crystallization of calcium phosphate or MAP (struvite). Calcium phosphate, for example, hydroxyl apatite, is a product directly comparable to rock phosphate. Thereby, one has to keep in mind that in practice the kinetics of calcium phosphate precipitation plays a major role than thermodynamic equilibrium considerations. Thus, in most cases, spontaneous precipitation of calcium phosphate from the solution does not occur at all or only at very high oversaturation. However, the separation of calcium phosphate can be achieved by adding seed crystals, such as sand (Giesen et al., 2005) or calcium silicate hydrate (Berg, 2005), which are able to initiate the precipitation/crystallization process of calcium phosphate. To produce MAP, it is necessary to provide a stoichiometric ratio of magnesium, ammonium, and phosphate of 1:1:1. Filtrates from sludge dewatering are particularly suitable for MAP precipitation as only magnesium has to be added, cf. ATV-DVWK (2000, 2005). Due to thermodynamics, the separation of calcium phosphate and MAP only takes place in alkaline pH medium (pH value approximately 8–10; cf. Stumm and Morgan, 1996 and Wu and Bishop, 2004). Due to the ammonium contents, the utilization of MAP in the phosphate industry is limited. However, direct use as fertilizer seems to be possible. Laboratory-scale tests showed that the uptake of phosphorus from MAP on acidic and neutral soils is comparable to the uptake of triple superphosphate (Richards and Johnston, 2001; Ro¨mer, 2006).
4.12.5.2.5 Exemplary applications of phosphorus recovery Wastewater: crystallization of calcium phosphate – Crystalactors (The Netherlands). DHV Water (The Netherlands) developed a crystallization process for the recovery of phosphorus. Thereby, a so-called Crystalactors is used, a cylindrical fluidized-bed reactor with, for example, sand as seed material. By adding calcium, phosphorus crystallizes on the seed material (quartz sand) in a fluidized bed at pH values of approximately 9, thus forming calcium phosphate. Due to the crystallization, the pellets grow, and separation is possible (Giesen et al., 2005). In 1993, the process was realized in side streams of the wastewater treatment plants Geestmerambacht (230 000 PE) and Heemstede (35 000 PE), Germany, in combination with biological phosphorus removal. With a concentration of 60– 80 mg l1 the phosphorus-rich supernatant liquor from the stripping tank of the biological phosphorus removal unit is fed into the Crystalactors. Here, 70–80% of phosphorus are eliminated and separated. In order to minimize precipitation of calcium carbonate, the pH value is decreased in the inflow area, thus inducing the stripping of carbon dioxide. The separated phosphorus-rich pellets are utilized as a substitute for rock phosphate in the phosphate industry (Giesen et al., 2005). Wastewater: crystallization of calcium phosphate – P-RoC (Germany). Via the application of suitable seed crystals, such as calcium silicate hydrate, a by-product from the building materials industry, a process was developed in the Forschungszentrum Karlsruhe (Germany) which allows
the separation of phosphorus without the dosage of further chemicals. Phosphorus-rich water is fed into a crystallization reactor. By adding seed crystals, calcium phosphate is formed which can then be separated. The phosphorus removal rate is approximately 80%, almost independent of the organic constituents of the water. The phosphorus-rich product can be used in agriculture as well as in the phosphate industry. The process was investigated in pilot-plant scale. Recently, it was shown that the calcium silicate hydrate can also be applied directly to sewage sludge for phosphorus removal and recovery (Petzet and Cornel, 2009). In this process, the calcium silicate hydrate is added to the digester where it removes and recovers phosphorus released during anaerobic stabilization, thus providing a solution for operational problems related to the enhanced biological process removal (EBPR) process such as struvite scaling and high P return loads to the head of the treatment plant. Process water from sewage sludge treatment: crystallization of MAP (Japan, Canada, Germany). Repeatedly, there have been reports about incrustations of pipes following the digestion step, in particular in those wastewater treatment plants with biological phosphorus removal. Due to the formation of ammonium during digestion in combination with dissolved phosphorus and magnesium, slight changes in the pH value can induce spontaneous precipitation of MAP, which can lead to the incrustation of pipes (cf. Heinzmann and Engel, 2005). In this context, processes have been developed, for example, in Japan, Canada, and Germany, which focus on the formation and separation of MAP. In the case of the PHOSNIX process developed by Unitika (Japan), phosphorus- and ammonium-rich process water is fed into a fluidized-bed reactor. The pH value is adjusted to approximately 8.5–9 by adding sodium hydroxide solution, and by adding magnesium MAP crystals are formed which can then be separated. By applying this process, phosphorus removal rates of approximately 90% can be achieved. The generated product can be used in agriculture. Since 1987, an industrial-scale plant is operating in Japan (Ueno, 2004). Similar technologies have been investigated in Canada, OSTARA process (Prasad et al., 2007) operated in industrial scale since 2007, and in Germany, PRISA process – which is so far not realized in industrial scale (Pinnekamp and Montag, 2005). Digested sludge: wet chemical re-dissolution and crystallization of MAP – Seaborne process (Germany). In winter 2006, the Seaborne process was put into operation at the wastewater treatment plant Gifhorn (Germany), a municipal treatment plant with approximately 50 000 PE. Following anaerobic stabilization, sulfuric acid is added to acidify the digested sludge achieving a pH value of approximately 3. In order to improve dewaterability, hydrogen peroxide is added and the sludge is dewatered. The dewatered sludge is thermally recycled via a mono-sewage sludge incineration plant. Heavy metals are precipitated by adding sodium sulfide and separated with a belt filter press. Subsequently, magnesium hydroxide is added and the pH value is increased by adding sodium hydroxide solution. This procedure results in the precipitation of MAP, which can be separated by centrifuges and used in nutrient recycle. The residual water passes an ammonium/ammonia stripping with subsequent acidic wash,
Wastewater as a Source of Energy, Nutrients, and Service Water
thus producing diammonium sulfate. The water from which the nutrients have been removed is fed into the inflow of the wastewater treatment plant (Mu¨ller et al., 2005; Wittig, 2007). Sewage sludge ash: wet chemical redissolution and sequential precipitation – SEPHOS process (Germany). In the case of the sequential precipitation of phosphorous (SEPHOS) process, the first step is the elution of the sewage sludge ash with sulfuric acid. After removing undissolved residuals, the pH value in the filtrate is increased stepwise, whereas at pHo3.5 aluminum phosphates precipitate. The heavy metals such as copper and zinc remain dissolved and precipitate at pH values 43.5. Aluminum phosphate, poor of heavy metals, can be used in the electrothermal phosphate industry. By an alkaline treatment of aluminum phosphate (advanced SEPHOS process) phosphorus as well as aluminum is dissolved. By adding calcium, precipitation of calcium phosphate can be achieved. Aluminum stays in solution and can be recycled as coagulant. Respective investigations have been carried out at the Institute IWAR of Technische Universita¨t Darmstadt (Schaum, 2007). Sewage sludge ash: thermochemical treatment (Germany). Based on the thermochemical approach, ashes are exposed – under suitable conditions – to chlorine-containing substances, potassium chloride or magnesium chloride, and treated thermally. With temperatures 41000 1C, a large percentage of the heavy metals are turned into heavy metal chlorides which vaporize, thus removing them from the ashes (Kley et al., 2005; Prinzhorn 2005). The thermochemical treatment of the sewage sludge ash/chloride mixtures is performed in quasiclosed systems, for example, rotary furnaces. The chlorides are discharged via the gas phase with subsequent precipitation during flue gas cleaning. By applying the mentioned chlorides, potassium and/or magnesium phosphates are formed which can then be used in agriculture. By subsequent specific dosage of nitrogen and/or potassium – following the removal of heavy metals – various multinutrient fertilizers can be produced. After pellets are formed, they can be used as granulates (Prinzhorn, 2005). Respective investigations are carried out within the frame of the EU research project – SUSAN – Sustainable and Safe Re-use of Municipal Sewage Sludge for Nutrient Recovery (SUSAN, 2008). The startup of a pilot plant in Leoben (Austria) with a load of 4000 Mg a1 ash was in 2008. Probably the most cost-efficient method is the direct use of sewage sludge ashes as substitute for phosphate rock. One precondition for electrochemical phosphorus recovery is that iron concentrations in the ash are low, as this leads to the formation of low-grade iron phosphate within the process reducing the phosphate yield. Thus, only ashes from sewage sludge treatment plants with biological phosphorus elimination or precipitation with aluminum salts can be used. Furthermore, concentrations of copper, zinc, and other heavy metals should be as low as possible. Currently, the use of sewage sludge ashes as phosphate rock substitute in the fertilizer industry is investigated in various institutions. With 7–8%, the phosphorus concentrations are somewhat lower when compared to 12–15% in phosphate rock. Concentrations of several heavy metals, especially copper and zinc, are higher, while cadmium and uranium concentrations are considerably lower in the ashes.
361
Considering German boundary conditions, annual costs of phosphorus recovery are estimated at 2–5 h per capita. These costs equal approximately 2–4% of the specific annual costs of approximately 124 h (C a)1 for wastewater treatment and disposal (BGW/ATV-DVWK, 2003).
4.12.6 Water Reuse Wastewater can be a valuable resource that contains the resource water in concentrations of more than 99.5%. After adequate treatment, that is, adapted to its subsequent application, water can be a valuable product to be reused. Thus, water reuse is an essential component of integrated water resource management, not only in arid and in water-deficient areas, but increasingly also in most of the densely populated urban areas, where water demand and supply diverge widely, at least regionally. Water reuse opens up new water resources and reduces the demand for potable water. Potential regional or local lacks of water can be closed. In addition, water reuse reduces the discharge of (treated) wastewater into water bodies. In regions where water supply may be energy-intensive and costly due to extensive transportation and/or pumping, water reuse can be an alternative with lower energy consumption and lower costs than using freshwater. Moreover, finally yet importantly, valuable freshwater resources, such as high-quality groundwater, can be preserved via the alternative use of reclaimed water. Today, the reuse of wastewater for agricultural irrigation is practiced in almost all arid regions. Particularly in threshold and developing countries of Latin America, Asia, and Africa, raw wastewater or insufficiently treated wastewater is used directly in crop irrigation. Often, this is done deliberately, in order to use the nutrients N and P as well as the organic load for forming humus, but often also without being aware of the health risks for farmers, farm laborers, and consumers. This is also true for the indirect and unplanned use of wastewater, where untreated wastewater is discharged into rivers and downstream river water is used for irrigation. ‘‘This major health concern makes it imperative to governments and the global community to implement proper reuse planning and practices, emphasizing public health and environmental protection, during this era of rapid development of wastewater collection and treatment’’ (EPA, 2004). An additional challenge is water supply and disposal in large cities. According to UN-HABITAT (2006) until 2050 approximately 75% of the world’s population will live in cities, 20% in urban areas and cities of 1–5 million inhabitants. The majority of approximately 80% will be living in threshold and developing countries, over half the world’s urban population in Asia. ‘‘According to the conclusions of various water reuse surveys (Lazarova et al., 2001; Mantovani et al., 2001), the best water reuse projects, in terms of economic viability and public acceptance, are those that substitute reclaimed water in lieu of potable water for use in irrigation, environmental restoration, cleaning, toilet flushing, and industrial uses’’ (EPA, 2004). Intra-urban reuse of water for utilizations, which do not require drinking water quality, offers a high potential to save valuable water resources and reduce (waste-)water discharge.
362
Wastewater as a Source of Energy, Nutrients, and Service Water
The freshwater consumption can be reduced by more than 30– 40% when reclaimed water is used for toilet flushing, gardening, irrigation, etc. (Bieker et al., 2009). However, intra-urban water reuse fosters the transition from conventional centralized to nodal, semicentralized supply and treatment systems, with short distances from the firsthand user to the treatment units and back to the secondhand reusers. This will minimize the energy for transport and treatment and offers the chance to recover heat from wastewater, especially from greywater. Again, to avoid any health risks, proper planning and professional operation of water reclamation plants have to be guaranteed. Intra-urban reuse and agricultural reuse are not mutually exclusive; on the contrary, they can be combined, as intraurban water reuse is a so-called nonconsumptive use. Nonconsumptive uses are all kinds of uses by which water is not physically lost but undergoes just a change in its quality, for example, in power plant cooling, where most of the water is returned to the water body with almost no quality deterioration and can be used again. The same is true for most of the municipal and industrial waters that do not disappear by use but are returned to the rivers as (treated) wastewater and might be used again downstream. Agricultural irrigation, on the other hand, is a consumptive use because water is evaporated by the plants and is no longer available for other uses. Therefore, consumptive uses put higher pressure on water resources than nonconsumptive uses and should be considered as the last step of a reuse chain.
4.12.6.1 Reuse Options The fields of application for water reuse are manifold. According to Asano (2007, p. 24), the following categories of water reuse applications for reclaimed water originating from treated municipal wastewater can be established:
• •
• • • •
Agricultural irrigation: Crop irrigation and commercial nurseries. Nonpotable intra-urban uses: Toilet flushing, landscape irrigation like in parks, golf courses, greenbelts, residentials, cemeteries, freeway medians, school yards, fire protection, and air conditioning. Industrial recycling and reuse: Cooling water, boiler feed, and process water. Recreational/environmental uses: Lakes and ponds, streamflow augmentation, fisheries, and snowmaking. Groundwater recharge: Groundwater replenishment, saltwater intrusion control, and subsidence control. Potable reuse: Blending in water supply reservoirs, blending in groundwater, and direct pipe-to-pipe water supply.
Corresponding to the manifold applications, the requirements on the water quality as well as on the treatment technology strongly vary. In the following, reuse for agricultural irrigation, nonpotable intra-urban reuse, industrial recycling, and groundwater recharge are briefly discussed. For detailed reviews, the publications of Asano (2007) and Jime´nez and Asano (2008) should be referred to.
4.12.6.1.1 Agricultural reuse The reuse of treated wastewater for agricultural irrigation presents by far the largest potential. Worldwide, more than 70% of the used freshwater is for agricultural irrigation (United Nations, 2003). While in Central Europe, the water demand for agricultural production is currently provided by sufficient rainfall during vegetation periods, in many countries of Latin and South America, Africa, and Asia between 70% and more than 90% of the annual water demand is needed for agricultural irrigation. Figure 13 shows the percentage of water withdrawals for agricultural use in different regions. Due to its high water demand, agriculture is also by far the largest reuser of water. In most arid regions, wastewater utilization is common practice and a necessity. In Figure 14, those countries with the largest surface areas under irrigation – according to their own statements – with untreated and/or treated wastewater are listed. Although the data are afflicted with great uncertainties and the degree of wastewater treatment will vary enormously, nonetheless, the diagram shows the importance of water reuse for agricultural irrigation. ‘‘It is estimated that at least 20 000 000 ha in 50 countries are irrigated with polluted water (United Nations, 2003), either directly or indirectly, and that 10% of the world’s population consume crops produced with wastewater (Smit and Nasr, 1992). The relative importance of this practice varies by country; in Hanoi, Vietnam, for instance, up to 80% of the vegetables consumed are produced with wastewater (Ensink et al., 2004)’’ (Jime´nez and Asano, 2008: 21). Farmers use wastewater for irrigation, as it is constantly available and, in addition, contains nutrients and humus formers (IMWI, 2003; Jime´nez and Gardun˜o, 2001). Only few are aware of the health risks arising from handling nondisinfected reuse water or are able to foresee the consequences of irrigation, such as salinization of soils or the risk of groundwater pollution. One phenomenon, which is often left unnoticed, is the socalled urban agriculture. What is meant is the cultivation of small parcels of land (0.5–2 ha) in urban and peri-urban areas for producing fruit trees, fodder, flowers, and vegetables. Here, often, untreated wastewater is used for irrigation. It is estimated that several million farmers are practicing urban agriculture and up to 50% of the vegetables offered for sale are being produced this way (Cornish and Lawrence, 2001; IMWI, 2003; Jime´nez and Asano, 2008). Despite the mentioned large numbers, the percentage of reuse water is only approximately 1% of the total water demand for agricultural uses. This leads to the question: What contribution can the reuse of treated wastewater make toward the reduction of water shortage? The proportions make clear that the reuse of municipal wastewater can only contribute modestly within the total water balance and is far from sufficient as a sole source for agricultural production. Even if one does not take into account that availability and demand do not necessarily coincide and under the assumption that the total amount of household wastewater can be used for agricultural irrigation, with 50 m3 a1 from private households and a typical irrigation efficiency rate for sprinkler irrigation of 65%, it is possible to irrigate an area of approximately 20–80 m2 sufficiently, depending on the type of plant (Table 11). This is a
Wastewater as a Source of Energy, Nutrients, and Service Water
Agriculture
Domestic
363
Industrial
100% 80% 60% 40% 20%
Low-income countries
Middle-income countries
High-income countries
Developing countries
Developed countries
Sub-Saharan Africa
South America
Oceania
North America
Middle East and North Africa
Europe
Central America and Caribbean
Asia (excluding Middle East)
World
0%
Figure 13 Water withdrawals by sector for 2006 (Jime´nez and Asano, 2008).
rather small contribution compared to the area of several thousand square meters that is needed for food production per person. When specific yields are taken as a basis, the product quantities that can be generated by using only municipal water for irrigation can be estimated (Table 12). This illustration, as well, shows that these amounts are comparably small and are by far not sufficient to cope with the annual demand on food and energy per person. Although agricultural reuse of wastewater cannot solve the problem of global water shortage and thereby cannot guarantee safe food supply on its own, it presents a certain contribution to sustainability. As discussed earlier, agricultural reuse is not the only reuse option: it can be seen as the last link of consumptive use within a chain of nonconsumptive uses with decreasing quality demands (multiple reuses).
4.12.6.1.2 Intra-urban reuse as service water The objective of intra-urban reuse of water (after respective adequate treatment) is to preserve local/regional drinking water resources. Interim storage is hardly necessary, as water availability and demand occur almost synchronically, delayed to each other only by a few hours. However, due to dual piping for drinking and service water, etc., technical efforts are higher. Intra-urban water reuse facilitates the reduction of the specific drinking water consumption (directly and year-round) to the quantity needed for cooking, drinking, and personal hygiene. Figure 15 shows typical water uses, exemplarily for the city of Qingdao, China, with an average water consumption of 109 l (C d)1 (BMBF, 2006) and for USA with a specific water consumption of 274 l (C d)1 (Asano, 2007: 11).
Although the specific water consumption differs, in both cases the quantity of slightly polluted greywater from showers and washing machines is sufficient to cover the demand of water needed for toilet flushing. It is comparably easy to treat greywater, as it mainly contains only organic impurities, washing agents, and personal care products. Due to mostly low concentrations of N and P in greywater (cf. Table 7 in Section 4.12.3.2), nutrient removal is expected to be dispensable; however, disinfection is usually required in order to reach quality standards like 0.2 mg l1 residual chlorine content at the extraction point (GB/T 18920-2002). Just by using adequately treated greywater, the demand of potable water could already be reduced by 30%. At the same time, the amount of wastewater to be treated would decrease by the same amount (cf. Section 4.12.7). Since treated greywater is used for toilet flushing and subsequently treated as blackwater, blackwater treatment represents the last barrier prior to discharge into the environment. Persistent compounds, such as fragrances and other compounds in personal care products that may only be partially removed during greywater treatment, are of no concern in toilet flushing and can be further eliminated during blackwater treatment. Direct reuse of reclaimed water in households fosters nearby treatment on various accounts. On the one hand, in case treatment is carried close to the greywater’s origin, its high temperature can be used for heat recovery. On the other hand, collection and distribution pipes are short and costs for pumping and losses during transport are low. One of the most important questions is the consumers’ acceptance of reclaimed water. Social-empirical studies have revealed that the more is known about the origin (and whether it is from the own surroundings), the higher is the acceptance (Jeffrey and Jefferson, 2004). Particularly with decentralized, quarter-, or district-based,
364
Wastewater as a Source of Energy, Nutrients, and Service Water Ha irrigated with untreated wastewater (China out of scale) 0
20 000
40 000
60 000
80 000
100 000 120 000
140 000
160 000
180 000
200 000
1 300 000
a
China Mexico India Chilec,a Syna Pakistan Colombia Argentin SAa Ghana Vietnam Peru Turkey Morocco Egyptc Kuwaitb Sudan Tunesia Nepalc Bolivia Ha irrigated with treated wastewater Chile Mexico Israelc,a Egypta Cyprus Italya Argentin Australia UAEc USA Jordan Turkeya Syriab Tunesia Kuwaitc Omanc France Libyac S. Arabia Germany
Note : Information may vary from source. Some countries report agricultural wastewater use without mentioning the amount of hectares involved. aData are confusing. bNo data available, although the practice is reported. cSurface might be greater. Figure 14 The 20 countries reporting the largest surface areas under irrigation with treated and untreated wastewater. From Jime´nez (2006) and (Jime´nez and Asano, 2008).
(grey-)water treatment, heavily polluted wastewater from industry and commerce can be excluded from reuse. Certainly, in case the complete wastewater stream is treated adequately and used as reclaimed water, considerably more water can be reused. This is mostly done in central treatment plants. Then, intra-urban reuse is reasonable in rather central units, such as for fire protection through reclaimed water fire hydrants, in industry, in commercial uses such as vehicle washing facilities, laundry facilities, street cleaning, or for irrigating public parks and recreation centers, athletic fields, highway medians, and shoulders, landscaped areas, and golf courses. As illustrated above, intra-urban water reuse not only preserves valuable resources, but often it is also more energy-efficient and more cost-effective. In particular, this is true in case:
• •
freshwater has to be transported over long distances, inferior raw water quality requires high efforts in drinking water treatment,
• •
seawater has to be processed for drinking water use, and stringent surface water discharge requirements are given.
Water reclamation facilities must provide the required treatment to meet appropriate water quality standards for the intended use. Usually, the discharge of wastewater treatment plants with nutrient elimination is filtrated in order to remove residual solids and is disinfected. Most standards require a defined residual chlorine concentration at the extraction point. Because urban reuse usually involves irrigation of properties with unrestricted public access or other types of reuse where human exposure to the reclaimed water is likely, reclaimed water must be of a higher quality than may be necessary for other reuse applications (see also Section 4.12.6.2).
4.12.6.1.3 Industrial reuse Industrial reuse and recycling is mainly driven by economic forces. Thereby, the main focus is on internal water recycling.
Wastewater as a Source of Energy, Nutrients, and Service Water Table 11 Irrigable area for different agricultural crops, with 50 m3 a1 of water and an irrigation efficiency rate of 65% Agricultural crop
Beans Cabbage Rice Sorghum Wheat Tomatoes Peanuts Corn/maize Cotton Sunflowers Lemons Bananas
Demand on irrigation water per vegetation period (l m2)
Irrigable area (m2)
Min
Max
Medium
300 380 350 450 450 500 500 500 700 800 900 1200
500 500 700 650 650 700 800 800 1300 1200 1200 2200
89 77 72 63 63 57 55 55 37 35 33 21
Data for demand on irrigation water from Lazarova V and Bahri A (2005) Water Reuse for Irrigation – Agriculture, Landscapes and Turf Grass. Boca Raton, FL: CRC Press.
Table 12 Theoretically producible quantity of foods in case 50 m3 a1 of water are used with an assumed irrigation efficiency rate of 65% (specific yield according to Katalyse, (2008), specific irrigation demand according to Zehnder (2003)) Product
Producible quantity (kg)
Sorghum Corn/maize Wheat Clover (Trifolium) Tomatoes Cucumbers Oranges Sunflowers Cotton Bread
133 45–95 45–65 72 52 47 17 4.9 0.01 22–33
Electric energya
166 kW h electric energy (Benergy equivalent of 17 l petrol/diesel)
a
Estimation based on Rosenwinkel (2006).
Qingdao, PR China
Cooking/dish washing 25
Cleaning Drinking 3 7
Toilets flushing 33
365
USA Showers + baths 33
Faucets 42 Dish washers 4
Washing machines 8
Other domestic 5
Toilets flushing 76
Showers + baths 52
Washing machines 57
Figure 15 Typical water uses, left for the city of Qingdao, PR China, right for the USA (values in l (C d)1) (BMBF, 2006; Asano, 2007).
The reuse of water from municipalities and other providers is less attractive, as it produces dependencies, requires quality controls and negotiations with external contractors (Jime´nez and Asano, 2008). Utilizing reuse water for cooling is probably the most common reuse, as, among others, large quantities of water are needed, and quality requirements are comparably low. Usually, filtration and softening to avoid clogging and scaling are sufficient. Power plants are large consumers of reclaimed water (Jime´nez and Asano, 2008). Water recycling and reuse in industry in developed countries is an established and well-developed practice. This can be pointed out with the example of Germany. The volume of water utilized in the German manufacturing industry amounts to roughly 30 200 million m3 a1 (thereof 22 400 million m3 a1 for cooling), the freshwater supply to 6200 million m3 a1. The difference represents the volume of recycled water and amounts to 24 000 million m3 a1, a number which exceeds the total amount of municipal wastewater (9695 million m3 a1) by a factor of 2.4 (quantities for the year 1998, according to the Federal Statistical Office of Germany (FSO) (FSO, 2001; Cornel and Meda, 2008b). The use factor for water including cooling water – which is defined
as the quotient of utilized to provided water – varies between 1.3 in the textile industry up to 21.5 in the vehicle industry. Use factors indicate the extent of water reuse among different industry branches. Table 13 gives use factors among different industry branches. The values are increasing, for example, in the food industry from 3.5 in 1980 to 4.2 in 1998. They might vary in different countries and even in different regions as the driving force for water reuse in industry is, quite often, economics. Thus, reuse depends on water prices and wastewater fees on the one hand and on water treatment costs for adequate standards for internal reuse purposes on the other hand.
4.12.6.1.4 Groundwater recharge One has to differentiate between the so-called unintended groundwater recharge through intensive irrigation and the purposeful recharge with reclaimed water for reaching one of the followings goals:
• •
to prevent saltwater intrusion into coastal aquifers, to control or prevent ground subsidence,
366 Table 13
Wastewater as a Source of Energy, Nutrients, and Service Water Industrial water quantities and use factors for Germany
Industry
Manufacturing industry Food Metal Vehicle Textile Paper Chemical Power plants (public) Municipal wastewater
Wastewater
Water supply
Total million m3 a1
Thereof cooling water million m3 a1
Provided million m3 a1
Utilized million m3 a1
Thereof for cooling million m3 a1
Use factor
6008 363 822 86 175 547 3455 25 984 9695
4243 162 667 45 131 264 2639 25 842 -
6207 416 873 93 183 610 3422 26 559 -
30 226 1728 6018 1989 242 3485 11 836 67 734 -
22 486 834 4925 1092 172 816 10 594 57 457 -
4.9 4.2 6.9 21.5 1.3 5.7 3.6 2.6 -
From Cornel P and Meda A (2008b) Water reuse in Central Europe: The current situation. In: Jime´nez B and Asano T (eds.) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA.
• •
to ensure seasonal storage of potable or nonpotable water, and posttreatment for future applications.
Infiltration and percolation of reclaimed water use the natural purification processes within the soil, that is, filtration, adsorption, ion exchange, precipitation, biological degradation, etc., thus leading to additional cleanup and equalizing the water quality. In addition, the soil functions as a safety barrier. At the same time, by passing through the soil and being mixed with real groundwater, the water loses its identity as reclaimed water. Hereby, the acceptance is increased. Typical retention times in the so-called Soil Aquifer Treatment (SAT) are 20–50 days. The share of infiltrated water in the extracted and used water generally is between 40% and more than 90%. Detailed information on hydrology and the degradation of pollutants during SAT, riverbank filtration, and dune filtration can be taken from the literature (Fox et al., 2001; Drewes et al., 2001; Brauch and Schmidt, 2009; Oaksford, 1985). In case seasonal storage, for example, of irrigation water, is the main objective of infiltrating reclaimed water into the aquifer, prevention of evaporation losses, salinization, and algae growth, as they occur in surface reservoirs, are predominant. Storage processes require profound knowledge of the hydrological and geological conditions. The retention time can be much higher than with SAT. Normally, adequate pretreatment including nutrient elimination is required in order to protect aquifer and groundwater.
•
•
4.12.6.2 Fit for Purpose, Quality Requirements One precondition for unobjectionable and successful water reuse in the long term is the consistent compliance with defined quality standards. Here, various aspects have to be taken into account:
•
Hygiene/protection of human health. In order to protect human health, limit values for pathogens (bacteria, viruses, and helminth eggs) and toxic substances have to be defined and met. When defining these limit values, the type of
•
•
contact between humans and wastewater plays a decisive role. References can be taken from the guidelines (WHO, 2006). Soil protection/plant protection. In order to protect soils, first of all it is important to restrict the concentration of salts and heavy metals. A well-known example is the potential of soil salinization due to irrigation with sodium-containing wastewater. This reduces the water conductivity of the soil and therefore leads to harvest depression. Decisive factors for defining the required water quality are type of soil (texture, grain size, permeability, chemism, and specific salt content), climate (aridity, precipitation quantity and distribution, humidity, and wind), type of plant (nutrient balance in the soil), and irrigation technique (sprinkler irrigation or subsurface irrigation, irrigation quantity, and frequency). Besides the total salt content, the concentration of several ions toxic to plants, such as boron, chloride, and sodium, matters. These ions are taken up via the root system or via the leaves in case of sprinkler irrigation. Boron has a narrow tolerance range and can be toxic already in concentrations slightly above the essential concentration for plant growth. Boron reaches municipal wastewater in the form of perborate, that is, as bleach in washing agents and disinfectants, whereas NaCl exists as a regeneration salt for water-softening ion exchangers in dishwashers. Groundwater protection. Although professional application of irrigation tries to avoid infiltration into groundwater, in practice it is hardly possible to prevent irrigation water from reaching the aquifer at least to a certain degree. However, in case the aquifer is required as seasonal storage for irrigation water, its protection should be of special interest. Application of efficient irrigation techniques. Irrigation techniques also put demands on wastewater quality. Substances with corrosive impacts and insoluble substances should be eliminated during wastewater treatment, in order to prevent clogging of pipes and damage of equipment. Acceptance. In order to get the acceptance of the public, esthetic aspects also have to be considered. Wastewater for irrigation should therefore be as odorless as possible and colorless.
Wastewater as a Source of Energy, Nutrients, and Service Water
•
Storage during nongrowing seasons. In case wastewater for irrigation has to be stored during nongrowing seasons, specific quality standards for storage have to be met, for example, standards for nutrient contents in order to prevent excessive algae growth in case of aboveground storage or to minimize nutrient input into the groundwater in case of subsurface storage.
The US Environmental Protection Agency (EPA, 2004) has listed standards for wastewater quality for defined wastewater ingredients and their impact on wastewater reuse systems, see Table 14.
4.12.6.3 Treatment Options and Energy Requirements Treatment options conditions:
• • • • • •
depend
on
numerous
• •
Optional treatment steps for generating reclaimed water also include
• • • • • •
filtration to remove residual suspended solids; microfiltration and ultrafiltration to remove colloidal solids; nanofiltration, reverse osmosis, and electrodialysis to remove dissolved solids; if applicable or necessary, advanced oxidation, carbon adsorption, and ion exchange to remove trace constituents; disinfection with chlorine, UV, ozone, membranes, or others; and safety chlorination (Table 15).
4.12.6.3.1 Physical and chemical methods
purpose of reuse; quality requirements for use, storage, and transport; legal requirements and quality standards; irrigation technique in case water is used for irrigation; temperature when storing and tendency towards microbial recontamination in intra-urban reuse; and economic options.
preliminary treatment, such as screens and grit chambers; primary treatment, such as clarification and fine screens, sometimes enhanced by chemical treatment; secondary treatment, that is, biological treatment, with or without nitrogen and phosphorus removal and tertiary treatment, also referred to as advanced treatment, is generally defined as anything beyond secondary treatment.
Table 14
It might involve coagulation, flocculation, clarification, filtration, and disinfection.
boundary
Levels of wastewater treatment are generally classified as
• •
367
Starting with municipal raw wastewater, the first treatment objectives are the reduction of the solids content in order to protect the irrigation system and the irrigated soils. Furthermore, high solid contents can reduce the effect of the subsequent disinfection step. Mechanical treatment should therefore be considered as first treatment step, supported by chemical precipitation, if needed. Besides reducing the solid contents up to 90% and the organic load up to 70%, in combination with micro-sieving and/or sedimentation, this so-called enhanced primary treatment can also reduce healthrelevant parameters, such as the concentration of helminth eggs (up to three orders of magnitude), bacteria, and protozoa (up to two orders of magnitude). However, quality standards as listed in Table 14 cannot be met and residual COD concentrations are not low enough for subsequent disinfection. Conclusion. Compared with the use of raw wastewater for agricultural irrigation, the application of physical–chemical
The 20 countries reporting the largest surface areas under irrigation with treated and untreated wastewater
Parameter
Significance for water reuse
Range in secondary effluents
Treatment goal in reclaimed water
Suspended solids
Measure of particles. Can be related to microbial contamination. Can interfere with disinfection. Clogging of irrigation systems. Deposition
5–50 mg l1
o5–30 mg l1
1–30 NTU 10–30 mg l1 50–150 mg l1 5–20 mg l1 o10–107 cfu 100 ml1 o1–106 cfu 100 ml1 o1–10 l1 o1 –100 l1 -
-
o0.1–30 NTU o10–45 mg l1 o20–90 mg l1 o1–10 mg l1 o1–200 cfu 100 ml1 o1–103 cfu 100 ml1 o0.1–5 l1 o1/50 l o0.001 mg-Hg l1 o0.01 mg-Cd l1 o0.1–0.02 mg-Ni l1 4450 mg-TDS l1 0.5–41 mg-Cl l1
10–30 mg N l1
o1–30 mg l1
0.1–30 mg P l1
o1–20 mg l1
Turbidity BOD5 COD TOC Total coliforms Fecal coliforms Helminth eggs Viruses Heavy metals
Inorganics Chlorine residual Nitrogen Phosphorus
Organic substrate for microbial growth. Can favor bacterial regrowth in distribution systems and microbial fouling Measure of risk of infection due to potential presence of pathogens. Can favor biofouling in cooling systems
Specific elements (Cd, Ni, Hg, Zn, etc.) are toxic to plants and maximum concentration limits exist for irrigation High salinity and boron (41 mg l1) are harmful for irrigation To prevent bacterial regrowth. Excessive amount of free chlorine (40.05 mg l1) can damage some sensitive crops Fertilizer for irrigation. Can contribute to algal growth, corrosion (NH4–N), and scale formation (P)
From Jime´nez B and Asano T (eds.) (2008) Water Reuse: An International Survey of Current Practice, Issues and Needs. London: IWA.
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Wastewater as a Source of Energy, Nutrients, and Service Water
Table 15
Treatment levels achievable with some typical treatment trains
Secondary treatment Tertiary treatment
Suspended solids mg l1 Turbidity NTU BOD5 mg l1 COD mg l1 Ntot mg l1 PO4-P mg l1
Activated sludge
None (secondary effluent) 10–30 Granular media filtration o5–10 Filtration þ GACa o5 Coagulation/flocculation o5–7 Coagulation þ filtration o1 Aerated biofilter (BAF) 2–10 Maturation ponds 20–120
5–15 o0.5–5 o0.5–3 o10 o0.5–2 0.5–5 -
15–25 o5–10 o5 o5–10 o5 o5–15 o5–35
40–90 30–70 5–20 30–70 20–40 20–50 40–150
10–50 10–35 10–30 10–30 o5–25 10–30 5–25
6–15 4–12 4–12 o1–5 o1–2 4–12 2–6
Trickling filter
None (secondary effluent) 20–40 Granular media filtration 10–20 Coagulation/flocculation o5–10
5–15 10 0.5–5
15–35 15–35 o5–10
40–100 30–70 30–60
15–60 15–35 10–30
6–15 6–15 4–12
MBR
None (secondary effluent)
o5
5–50
o5–20
o1
o0.1–0.5
o0.1–10
a
Granulated activated carbon. From Lazarova V and Bahri A (2005) Water Reuse for Irrigation Agriculture, Landscapes and Turf Grass. Boca Raton, FL: CRC Press, p. 170.
treatment techniques presents a large improvement, as these methods allow the significant reduction of solids and partial elimination of the organic load and pathogens. In addition, investment costs for these techniques are much lower compared to biological wastewater treatment plants, thus facilitating the implementation in economically weak countries as well. Hence, physical–chemical techniques for wastewater treatment represent a transitional solution for improving the water quality and can therefore be the first step to be realized rapidly in those countries where raw wastewater is currently used for irrigation. However, the disinfection ability of only physically– chemically treated wastewater has to be viewed critically. High residual concentrations of organic substances and turbiditycausing substances put an efficient and ecologically sound disinfection into question. Depending on the disinfection method, there is potential for the production of undesirable by-products.
4.12.6.3.2 Biological treatment The second important treatment objective is the reduction of organic ingredients. Even though this treatment might not be obligatory for agricultural irrigation (organic compounds could even have a positive, improving effect on light sandy soils), it is very important for subsequent treatment, storage, and distribution of the irrigation water. This treatment step minimizes the potential of microbial recontamination and the risk of clogging of the pipes via the production of biofilms and allows for the efficient application of disinfection methods. Other important parameters are the nutrients, as mentioned earlier. For agricultural use there might be an advantage in not eliminating the nutrients but using their fertilizer effect; however, overfertilization and groundwater contamination should be prevented. Nutrient control is essential, in particular with seasonal storage of treated irrigation water in aboveground reservoirs, natural lakes, or aquifers. Quality standards as listed by EPA (see Table 14) show that the elimination of phosphorus is not inevitable; however, limit values of nitrogen can only be met with N elimination or with low-strength wastewater.
Conclusion. The discharge quality of biological wastewater treatment plants normally fulfills the quality standards regarding the solids content, organic ingredients, and nutrients. Deficits might be microbiological requirements, inorganic ingredients such as salts (sodium in particular) and boron as well as micro-pollutants.
4.12.6.3.3 Disinfection Water reuse requires a sufficient water quality adjusted to the intended purpose. In wastewater treatment plants, pathogens in terms of bacteria, viruses, parasites, and helminth eggs occur in concentrations far in excess of WHO guideline values and those quality parameters as listed in Table 14, respectively. In order to protect the health of people who come in direct contact with irrigation water, such as farmers and irrigation technicians, and also uninvolved people in the vicinity of sprinkler irrigation units and consumers who might be exposed to pathogens indirectly via the consumption of field crops, disinfection of irrigation water is good practice nowadays. One may do without, provided that health risks can be excluded by applying other measures, for example, irrigation techniques, off-times for irrigation before the harvest, cleaning measures before consumption, etc. The DWA topics ‘Assessment of process steps for the treatment of wastewater for reuse’ gives an overview on how different process steps affect the reduction of pathogens (DWA, 2008). The data are presented in log scale and they are additive, that is, by combining different process steps the total reduction degree can be estimated. In addition, the WHO guidelines (WHO, 2006) give information on nontechnical measures that reduce the risk of infection via irrigation water even further.
4.12.6.3.4 Other methods In case the salt concentration of the treated water is too high for direct use, desalination steps with reverse osmosis or ion exchanger can be installed downstream. This is necessary in particular in those cases when almost only municipal wastewater is used for irrigation, as practiced for example in Israel. Salts partly derive from drinking water. In households, the increase in salinity results from adding softening agents in
Wastewater as a Source of Energy, Nutrients, and Service Water Table 16 Range of energy demand for some selected biological treatment options according to Lfu (1998) Plant size Small a kW h m3 Aerated wastewater ponds 0.34 Biological contactor 0.23 Trickling filter 0.31 Activated sludge with aerobic sludge 0.61 stabilization Activated sludge with nutrient elimination 0.48
Large b kW h m3 0.29 0.25 0.17 0.312) 0.302)
to be broadly applicable, any alternative management method must be as transparent as practical to users, allowing them to pay a service fee and then flush and forget it, as they do in conventional, centralized systems.’’ Moreover, adds: ‘‘All system components need to be managed to the needs of the technologies employed. Operations and maintenance can not be left to the sole discretion of individual users’’ (Venhuizen, 1997). The following principles define the minimum requirements for a sustainable urban water resource management, at least in densely populated urban areas in which local/ regional water shortage is a considerable challenge for the coming decades:
•
a
Small stands foro1000 PE; large foro5000 PE. b Large for 10 000–100 000 PE.
• washing machines and dishwashers, while there might be an additional increase due to evaporation in open reservoirs. With desalination, one has to take into account that on the one hand approximately 20–30% of the water volume is lost as concentrate, and that on the other hand, the concentrate or regenerate needs further treatment or disposal. In addition, those salts are eliminated which have to be added again later as plant nutrients or trace elements.
4.12.6.3.5 Energy requirements
369
•
•
The level of comfort has to be assured. As for any alternative system, the handling of water reuse systems must be at least as easy and as reliable as in conventional systems and processes. The cost of new systems must not exceed that of existing structures, yet, in the ideal case, should be below the cost of conventional systems. This includes all parts of the system, from the processing of drinking water to the distribution and collection of wastewater and its treatment as well as the treatment and disposal of sewage sludge. Distribution pipes and sewers are important cost factors, regarding investment and operation and maintenance as well as tying up capital for a long time. Accordingly, pipe and sewer lengths should be kept as low as possible in order to minimize capital and operational expenditures. Professional operation is inevitable. Operation as well as maintenance has to be carried out by qualified personnel. Resource conservation is an essential aspect of sustainable water management. Water withdrawal has to be balanced with the natural regeneration rates. To succeed in this point, the percentage of potable water for transporting organic waste, feces, and pollutants has to be minimized respectively, at least in most densely populated areas around the world or just as much in water-scarce regions. Valuable constituents in the wastewater offer an additional approach toward resource conservation. They should be reclaimed, including water for further use. Thereby, the respective hygiene standards for all material streams to be reused or reclaimed have to be provided for.
The energy demand varies widely for the different treatment options. For mechanical treatment DWA (2009) indicates 1– 2 W h m3 for sedimentation (with or without flocculation) and 1–20 W h m3 for sieving (with or without flocculation). Aerobic biological treatment exhibits the highest energy demand. For German conditions, Lfu (1998) indicates 0.17– 0.61 kW h m3, depending on the adopted treatment and the plant size (see Table 16). For disinfection, for example, with ozonation, UV irradiation or membrane filtration, additional 0.035–0.4 kW h m3 must be calculated (Haberkern et al., 2008).
•
4.12.7 Recovery Fosters Decentralization
Against this background, the question arises as to the recommendable size of alternative sanitation systems with emphasis on water reuse and nutrient recovery. The required professional operation, ensuring control and quality standards as well as safeguarding the hygienic safety of drinking water and service water, involves a minimum size for operation units. The minimization of transport costs, the potential of heat recovery from warm wastewater, and the social acceptance of water reuse favor compact-scale decentralized systems. In order to determine the recommendable size of alternative sanitation systems, the fundamental results of the present article are summarized in the following.
As illustrated in the previous chapters, water and nutrients are valuable, yet scarce resources. During the last few years, various approaches toward alternative sanitation concepts have been developed with the aim to reduce water consumption and enable the recovery of nutrients (Cornel, 2007; Larsen and Gujer, 1997; Otterpohl et al., 1997; Venhuizen, 1997; Wilderer and Schreff, 2000; Zeeman et al., 2000). The spectrum ranges from low-tech processes to high-tech solutions with high demands on plant operation and maintenance. Here as well, there is no general solution. While compost toilets with subsequent agricultural application have proved to be practicable in rural areas, there are logistic and hygienic objections against their use in densely populated regions. As Venhuizen said ‘‘Many people have claimed ‘sewerless society’ would minimize problems’’ (with installing sewers). ‘‘Most of them propose composting toilets and other nonstandard plumbing, which would require significant lifestyle changes. But
•
4.12.7.1 Water As described above, water is a scarce resource in most urban areas worldwide (UN Water, 2007; BMZ, 2006). A general reduction of the daily demand can be achieved by applying use-dependent water qualities and domestic water reuse. Intra-
370
Wastewater as a Source of Energy, Nutrients, and Service Water
urban water reuse requires the water to be treated for use as service water, whereby the treatment efforts increase according to the pollution load as well as the needed quality. This means that the lower the water pollution and required water quality, the more cost-efficient the treatment. Greywater, a slightly polluted partial flow deriving from showers, hand wash basins, washing machines, etc. can be treated with relatively small efforts as its temperature is elevated and the nutrient content is low enough not to require nutrient elimination to reach service water quality. Water reuse requires dual piping for potable and nonpotable water as well as dual sewers in case grey- and blackwater (wastewater excluding greywater) are to be separated. In order to minimize capital and operational expenditures, distances have to be kept short. Due to the separate collection of greywater, blackwater accumulates as a separate material flow. High concentrations facilitate the reclamation of nutrients from blackwater and the sewage sludge, generated during the treatment process, respectively. In case service water can be reused in close surroundings of its production, there are also possibilities to reuse treated blackwater locally. In any case, local treatment and discharge into the receiving water bodies are reasonable, in order to keep local water amounts within the natural environment, minimize sewer lengths and related investment and operation costs.
4.12.7.2 Energy Water reuse of partial flows proves efficient also for the energy balance of water systems wherever water has to be transported over long distances or has to be treated energy-intensively because of low qualities. The processing of potable water from seawater via reverse osmosis requires a multiple of the energy needed to process service water from greywater. In figures, around 3–4 kW h m3 are needed compared to approximately 0.5–1 kW h m3 for producing reclaimed water from treated wastewater (Keller, 2008) or for treating greywater. (see Section 4.12.2, respectively Section 4.12.3.1). As described earlier, separating different water flows proves to be reasonable: greywater from showers and washing machines with considerably higher temperatures allows a significantly more efficient utilization of the temperature gradient in generating caloric heat, compared to energy recovery from combined sewer systems on lower temperature levels. However, applications for using the recovered energy are needed, and in order to minimize heat losses, those should be located as close to the place of origin as possible. One positive side effect of energy recovery is the reduced risk of microbial recontamination in the cooled down, reclaimed water that usually has to be stored for a short time to bridge the gap between supply and demand. The lower the storage temperature, the lower is the risk of microbial recontamination in the storage and distribution systems and the lower is the demand of chemicals for disinfection.
4.12.7.3 System Scale The size of a system plays a decisive role regarding costs of sanitation systems. Thereby, sewer and pressurized distribution
grids cause a significant share in the cost of the overall system (Gu¨nthert and Reicherter, 2001). Unaccounted losses of treated high-quality water during distribution in pressure pipes, which can amount to 40–50%, also present a significant cost factor and should therefore be taken into account when assessing the system’s scale. Against the background of the reuse of material flows (after prior treatment) and therefore the need of dual piping and sewer systems, investment and operation costs of the grid may be one of the limiting factors for the system size. Existing concepts such as DeSaR or ecosan follow this approach and focus on decentralized systems. However, in (fast-growing) urban areas with high population densities, the existing small-scale, on-site solutions do not seem to be feasible. Professional operation, stringent and reliable hygiene standards and its professional monitoring, as well as small footprints are indispensable, because of the expected comfort and because of epidemics prevention. When referring to treatment costs, economies of scale for treatment and operation have to be considered. Specific costs of large treatment plants can be reduced considerably by introducing larger-scale systems (Gu¨nthert and Reicherter, 2001; Reicherter, 2003). Regarding intra-urban areas with high population density, from the economic point of view one has to balance largescale plants generating economies of scale (in the plant sector) and small-sized, compact systems with short piping and sewer lengths. This is in accordance with the ecological point of view: optimum resource conservation requires a minimum size of technical plants, yet, at the same time, a compact piping and sewer system in order to minimize the energy input. From the sociocultural point of view, the focus is on hygienic harmlessness and comfort, the latter being, feasible with larger structures at lower costs. Thus, the optimum scale for reclaimed water application infrastructure is beyond the conventional centralized systems with supply and disposal for entire megacities, but rather lies in an effective materials flow management that is able to incorporate regional/local boundary conditions. Regarding economies of scale on the one hand and soft skills of infrastructure systems, such as flexibility, planning safety, and degree of capacity utilization, which all favor rather smaller systems on the other hand, latest research shows that the recommendable size of integrated semicentralized systems for new development areas ranges between 50 000 and 100 000 inhabitants (Bieker et al. (2010); BMBF, 2006). Anyhow, one has to bear in mind that size optima depend on local boundary conditions, treatment techniques, and on the advances in control technology. Taking the above into consideration, it becomes clear that a holistic approach must be chosen to fulfill the requirements of resource savings (ecological aspects), financial interests (economical aspects), and hygiene and safety needs (sociocultural aspects) in terms of sustainable water management. At the same time, the requirements convey that there cannot be a universal solution for everywhere, but the individual regional and locals circumstances and interests (including the financial bearing capacity of the region, the educational status of the people, climatic conditions, traditions, even religious concerns, etc.) need to be considered in order to find an adapted and locally-fitted solution (Wilderer, 2005b).
Wastewater as a Source of Energy, Nutrients, and Service Water
In the next section, a case study for a development area of 20 000 inhabitants in the city of Qingdao, P.R. China, is exemplified.
4.12.7.4 Case Study: Qingdao In 2004, 109 l of potable water per capita and day were used in Qingdao (BMBF, 2006) and as water becomes scarce, in future it will need to be generated from seawater. As depicted in Figure 16, 41 l (C d)1 greywater from showers, baths, and laundry are produced, whereas 33 l (C d)1 are needed for toilet flushing. Thus, the first consideration was to reuse treated greywater for toilet flushing in this completely new planned housing area. This reuse fosters the semicentralized approach and thus a supply and treatment center (STC) for the 20 000 inhabitants in this development area. As solid waste is disposed far outside of the city and not used for energy generation, in a second step the co-processing of the organic solid waste fraction was incorporated in the case study. Figure 16 provides an insight into the material and energy flows of such an integrated (solid waste and water) semicentralized – larger than on-site but smaller than centralized – supply and treatment system. In comparison to the sectored centralized approach, the integrated semicentralized approach can achieve large reduction rates in material and energy flows. Toilet flushing is operated with service water gained from greywater, thus saving 30% of potable water. Higher water reduction rates can be achieved by treating the whole amount of the arising greywater (greywater light plus hand washbasins and kitchens) and locally using it for irrigation of public
371
greens. The flexibility of the semicentralized approach allows an application-optimized operation, also in terms of service water. The treated greywater for nonpotable use in private households has to meet high quality standards. The example of China exposes the needed quality level: according to the Chinese water quality standard for urban miscellaneous water consumption (GB/T 18920-2002), water for toilet flushing has to fulfill the following requirements:
• • • • • •
TDSr1500 mg l1, BOD5r10 mg l1, NH4-Nr10 mg l1, anionic surfactantsr1 mg l1, coliformsr3 l1, and residual chlorineZ0.2 mg l1.
The integration of sewage sludge and waste treatment leads to an increase of the overall system efficiency and a decrease of the amount of residues to be disposed. At the same time, the sludge is stabilized and a solution for the currently tense and severely deficient treatment situation of wastewater sludge (openPR, 2008; Bfai, 2008) is given. The gained biogas from the integrated anaerobic treatment of sludge and waste is energetically sufficient to provide the electric energy demand of the treatment of all considered material flows (greywater, blackwater, and integrated sludge and waste treatment) within a semicentralized STC and even to produce a surplus of electric energy. An energy self-sufficient operation of the integrated STC is in that case possible. In terms of figures, the system demand is 25–50 W h (C d)1 respectively 9–18 kW h (C a)1 for greywater
Water treatment
Service water
76 l (C • d)−1 41 l (C • d)−1 Greywater
Greywater treatment Sludge
250 g (C • d)−1
9a−18b kWhelectr. (C • a)−1 Heat recovery 117−131 kWhcalor. (C • a)−1
Recyclables RDF Residual and biowaste 750 g (C • d)−1
Blackwater
68 l (C • d)−1
Waste and sludge treatment Process water
73 kWhelectr. (C • a)−1 610 g (C • d)−1 residuals
Sludge
Blackwater treatment
20* kWhelectr. (C • a)−1 68 l (C • d)−1(for discharge)
aActivated bMBR:
sludge treatment membrane biological reactor
Figure 16 Material and energy flows in an integrated semicentralized supply and treatment system (scenario greywater light reuse) – the case of Qingdao, P.R. China (calculated with 160 l biogas(C d)1100 l CH4(C d)11 kW htotal energy(C d)1300 W hel.(C d)1).
372
Wastewater as a Source of Energy, Nutrients, and Service Water
treatment, according to the chosen treatment method. Additionally, 55 W h (C d)1 respectively 20 kW h (C a)1 are needed for blackwater treatment. The conversion of biogas into electricity generates approximately 300 W h (C d)1 respectively 110 kWh (C a)1. Approximately 100 Wh (C d)1 respectively 36 kWh (C a)1 are needed for solid waste treatment, so there is a surplus of 200 W h (C d)1 respectively 73 kW h (C a)1. Deducing the system needs for greywater and blackwater treatment, an energy surplus of 95– 120 W h (C d)1 respectively 35–44 kW h (C a)1 is to be reflected in the electric energy budget. Additionally, the caloric heat of the separated greywater can be recovered. Assuming a level of efficiency of heat pumps between 0.45 and 0.5 (see Section 4.12.4.1), between 320 and 360 W h (C d)1 respectively 117 and 131 kW h (C a)1 of caloric heat can be gained from greywater for heating purposes, while assuring a reduction of bacterial regrowth in the service water for intraurban reuse. In addition, the conversion of biogas into electricity generates approximately 700 W h (C d)1 respectively 256 kW h (C a)1 caloric heat, most of which is used for the thermophilic digestion process. It has to be stated that from the energy point of view the reuse of wastewater of any quality is only recommendable, if the energy needed for treatment is lower than the energy demand for the transport of other water sources – as long as those other sources are locally available in sufficient quantities. Concurrently, the energy production from waste and sludge improves the carbon footprint of the STC. The energy is (nearly exclusively) gained from organic material, the wastewater treatment sludge as well as bio-waste and residuals. Using the biogas out of this sludge and waste, not only the energy bill is reduced to a minimum, but also the CO2 balance of the whole system is significantly improved. Ongoing research is going to clarify the carbon dioxide balance in a more detailed manner. Of course, the described scenario is only one out of several. Different boundary conditions cause different solutions, with different techniques at different scales. Nevertheless, water reuse as well as integrated energy recovery and nutrient recovery widen the future prospects of modern, reliable resource-conserving, economical, and sustainable infrastructure solutions.
4.12.8 Summary and Outlook Wastewater is a multisubstance mixture containing urine and feces, a multitude of hygienically questionable germs, potential pathogens and helminth eggs, personal care products, bleaching agents, pharmaceuticals and endocrine-disrupting compounds, and inorganic pollutants such as heavy metals and salts. Yet, wastewater also contains potential nutrients such as nitrogen and phosphorus as well as organic constituents, which are potential sources of energy. The most important resource, however, can be water itself which, in domestic wastewater, accounts for more than 99.5%. Quantitatively, agricultural reuse of adequately treated water represents by far the largest potential for water reuse. In some countries, for example, Israel, more than 75% of municipal wastewater is reused for agricultural purposes.
Depending on local conditions, nutrients contained in the wastewater can also be utilized. Hereby, care is to be recommended to prevent overfertilization. However, one of the real challenges lies in dealing with water supply and demand for irrigation not coinciding in terms of time and/or location, requiring long-distance transport and storage of reclaimed water. Further challenges are the salt content of the wastewater, as salt can accumulate in soil, as well as the hygienic quality of the reclaimed water with respect to farmers, farm laborers, and consumers. With climate changes on the one hand and the increasing cultivation of energy crops on the other hand, one has to expect the demand for irrigation water to increase as well. With the latter, the acceptance of reclaimed water will be high, as these plants are not used for food production. Another large field of application is intra-urban reuse as nonpotable water. Increasing urbanization and ever-growing megacities aggravate the gap between water demand and availability (regional/local). Via intra-urban water reuse, implemented in new housing or development areas, the specific freshwater demand can be reduced by 30% up to 50%. Intra-urban reuse fosters nodal, semicentralized supply and treatment in order to minimize cost- and energy-intensive multiple transport. As intra-urban reuse is mostly nonconsumptive, multiple water use is possible, for example, treated greywater for toilet flushing and, subsequently, treated blackwater for irrigation and nutrient utilization. Reuse of water can contribute to the preservation of valuable freshwater resources. At the same time, water reuse might contribute in saving electric power. In particular, this is the case when freshwater has to be transported over long distances or elaborate treatment is required, for example, processing of potable water from brackish water or seawater. The generation of reclaimed water only requires a fraction of the energy needed for the desalination of seawater. Regarding the generation of energy from wastewater, the largest potential lies in heat recovery, for example, for heating of shower or laundry water, which is the more efficient, the warmer the wastewater is. Caloric heat recovery fosters nearby, decentralized units and is most effective by using relatively warm greywater. Generating electric power from organic water constituents is subject to tight limits. Only in exceptional cases, with low requirements on discharge quality or in combination with cofermentation, it will be possible to run a self-sufficient operation of aerobic wastewater treatment plants with sludge digestion. (Strictly speaking, co-fermentation is an antagonism to self-sufficient wastewater operation, as the organics do not stem from the wastewater itself.) With anaerobic wastewater treatment plants, one has to consider that with temperatures below 23 1C, 40% of the COD are degradable at best. Furthermore, with a temperature range of 15–23 1C and a methane partial pressure of approximately 0.66 bar in the digester, approximately 16–19 mg l1 methane, that is, 64–76 mg l1 COD, remain dissolved in the treated wastewater. With current techniques, this methane cannot be used for energy production; even worse, in the subsequent treatment steps, the methane is stripped and thus increases the greenhouse gas emissions. As the effect of methane as greenhouse gas is 25 times higher when compared to CO2, methane emissions are not to be neglected.
Wastewater as a Source of Energy, Nutrients, and Service Water
Besides potassium and others, wastewater contains the nutrients phosphorus and nitrogen. While nitrogen is available worldwide in sufficient amounts, exploitable phosphorus resources seem to be finite. Regarding nitrogen recovery, the main question is: What is the energy input for the reclamation from wastewater in comparison with the production of nitrogen fertilizers via the Haber–Bosch process? While separate urine collection and its use as fertilizer substitute will be an alternative in small-scale rural areas, questions of logistics, energy-efficiency, hygiene, and acceptance will likely oppose its application in megacities. Moreover, the direct use might be restricted by the contained pharmaceuticals and priority pollutants. On the other hand, in the last few years, numerous processes have been developed by which phosphorus can be recovered from wastewater, sewage sludge, or sewage sludge ash, cost-efficiently and with unobjectionable hygiene quality (Petzet and Cornel, 2010). The phosphorus-containing products can be used either directly as fertilizer substitute, for example, as MAP (struvite), or as substitute for phosphate ore. Considering German boundary conditions, the per capita annual costs are 2–5 h, that is, 2–4% of the annual wastewater fees. Concluding remarks. We need new visions in order to cope with the arising challenges in infrastructure components within the urban century. Urbanization rates, never seen before, lead to new stages of resource scarcity within the cities of today and tomorrow. Conventional solutions offer reliable techniques and systems, but we need to see the bigger picture: Water reclamation will become an integral part of integrated water resource management in most places of the world, especially where no infrastructure has to be planned and built. Water recycling fosters decentralized, nodal structures. Moreover, water supply and treatment will need to be linked increasingly to energy demand and energy recovery. Thus, new technical developments and the enhancement of existing processes are necessary as well as changes in administrative structures, in order to enable people and institutions to take on responsibility for implementation, reliable, and long-lasting functioning, control, and maintenance of (new) infrastructure systems. The abolition of artificial barriers isolating water supply, wastewater treatment, solid waste treatment and energy management, and the involvement of the consumers are further challenges to be met. As we look into the future, we can expect more differentiation and the coexistence of large-scale centralized systems and small- and medium-scale de- and semicentralized solutions, adapted to the respective climate zone, settlement structure, population density, as well as the respective stage of development.
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4.13 Advanced Oxidation Processes M Sievers, CUTEC-Institut GmbH, Clausthal-Zellerfeld, Germany & 2011 Elsevier B.V. All rights reserved.
4.13.1 4.13.2 4.13.2.1 4.13.2.1.1 4.13.2.1.2 4.13.2.2 4.13.2.3 4.13.2.3.1 4.13.2.3.2 4.13.3 4.13.3.1 4.13.3.2 4.13.3.3 4.13.4 4.13.4.1 4.13.4.2 4.13.4.3 4.13.4.4 4.13.5 4.13.5.1 4.13.5.2 4.13.5.3 4.13.5.4 References
Introduction Fundamentals Generation of Free Radicals Homogeneous processes Heterogeneous processes Reaction Mechanisms Reaction Systems Competition kinetics Reaction modeling Guidance for Selecting an AOP Criteria to be Considered Cost-Related Factors of Ozone-Based Processes Cost-Related Factors of UV-Based Processes Description of Processes Ozonation Photo-Chemical Oxidation Fenton and Photo-Fenton Processes Process Combinations Full-Scale Applications Ozone-Based AOPs UV-Oxidation Processes Fenton Process Wet Air Oxidation
4.13.1 Introduction Advanced oxidation is used for various applications in wastewater treatment, water reclamation, indirect potable water reuse, drinking water production, and recently in micro-pollutant control of sewage treatment effluents. Compared to other technologies (e.g., membrane filtration, adsorption, ion exchange, evaporation, and stripping), the organic compounds in water are degraded rather than concentrated or transferred into different phases. Advanced oxidation processes (AOPs) have the ability to generate elevated concentrations of hydroxyl radical dOH, a strong oxidant capable of complete oxidation of most organic compounds into carbon dioxide, water, and mineral acids or salts. The attribution of advanced oxidation to the hydroxyl radical was explicitly mentioned by Glaze et al. (1987) at first. Besides the generation of hydroxyl radicals, many other free radicals are produced, but the hydroxyl radical is the species dominating the pollutant degradation efficiency. The free radical chemistry makes AOPs interesting to the destruction of recalcitrant, anthropogenic and toxic organic water pollutants, bacteria, viruses, and, last but not least, the emerging micropollutants also called as trace pollutants/organics. The advantage of AOPs is the relative high reaction power of hydroxyl radical. In literature, this reaction power is often expressed in terms of electrode potential versus hydrogen
377 377 377 377 385 386 387 387 389 394 394 395 395 396 396 396 398 399 399 400 402 404 404 404
electrode of redox reaction, but this is chemically not correct since the standard redox potential is only related to electron transfer, whereas the OH radical reacts by three different pathways and mostly on hydrogen abstraction pathway in the field of water treatment. As a result of the high reaction power, reactions with dOH radicals are very fast, often close to diffusion-controlled rates, and nonselective (Buxton et al., 1988). Due to the high oxidative and nonselective character of hydroxyl radicals relative to other oxidants, AOPs enable the conversion of nonbiodegradable into biodegradable compounds as well as the generation of undesirable by-products. Therefore, AOPs often need careful control of oxidant dose and/or strategies to avoid or minimize by-product formation. As a consequence, each application needs feasibility studies in laboratory and pilot scale before applying. Due to the large number of process options for AOPs and the limited space for this chapter, only a brief overview about fundamentals, reactions, applications, etc., can be given and some of the processes must have been omitted, unfortunately.
4.13.2 Fundamentals The fundamentals of AOPs include the generation and type of radicals, the fate and type of reactions, the fate of compounds
377
378
Advanced Oxidation Processes
during reactions, etc. Generally, radical reactions include three sections: (1) the starting section by generating radicals, (2) the radical reaction section, and (3) termination section by radical recombination. For generating radicals, there are several different options. Not only dOH radicals are generated, but also other radicals may have sufficient contributions as part of chain reactions depending on conditions of applications and processes. AOPs can be classified as indicated in Table 1 by the way of generation of radicals. Three main classes are given: (1) direct generation of radicals by physically based processes, (2) generation of radicals by the addition of oxidants, and (3) generation of radicals by the use of solid catalysts. Many AOPs include combinations of (1)–(3). Further classification may include the type of energy supply to activate radicalstarting reactions: (1) ultraviolet (UV) light irradiation at different wavelengths, (2) electrochemical power, and (3) temperature. The chemistry of AOPs is very complex due to the many reactions involved. Many details on improving the process efficiency are expected to be still unknown, and an example for this opinion may be the recently quantified contribution of advanced oxidation in ozonation processes during initial phase ozone decomposition, or the progress in development of UV light lamps as well as of new catalysts, for example, for solar radiation approaches.
Table 1
4.13.2.1 Generation of Free Radicals 4.13.2.1.1 Homogeneous processes The direct generation (without addition of oxidants) is as follows: Ionizing irradiation. For ionizing irradiation treatment, which is often called as electron beam (e-beam) process, high-energy electrons are produced and passed through the water. The generation of an electron is similar to a television by emission from a cathode and subsequent electrostatic acceleration. In e-beam processes for environmental applications, the acceleration potential of electron accelerators may range between 1 and 10 MeV (Mincher and Cooper, 2003). By ionizing irradiation, reactive species of both reducing and oxidizing character are generated. The high-energy electron dissipates its energy and generates the following species with indicated G values in terms of mmol J1 in neutral water
Overview and classification of advanced oxidation processes
Processes of free radical generationa
Type of external energy supplyb Without
Solar irradiation
Homogeneous oxidant type addition No addition
Ozone
Combined oxidants Heterogeneous oxidant type addition No addition
Ozone Hydrogen peroxide
UV lamp irradiation
Cavitation
V–UV irradiation
Acoustic cavitation Hydrodynamic cavitation
UV/cavitation O3/UV
Ozonation
Hydrogen peroxide
a
It is not possible to give a comprehensive description of all AOP processes, and therefore, the more commonly investigated and/or applied AOP processes are described in this article while for the others the reader is referred to the literature.
Fenton H2O2/O3
Photo-Fenton
Zero valent Fe Fe–gAl2O3
TiO2/solar
O3/activated carbon het.c Fenton H2O2/activated carbon H2O2/transient metal
O3/TiO2/solar Het. photo-Fenton
O3/cavitation
H2O2/UV Photo-Fenton H2O2/O3/UV
H2O2/cavitation
TiO2/UV
UV/cavitation
Heat/pressure
Electron beam
Pressurized ozonation
H2O2/O3/ cavitation
Cu (Loprox) Cu (wet air oxidation)
UV/cavitation/ ozone Het. photo-Fenton
OMP process
Classification only by radical generation, for pollutant degradation at least one oxidant, e.g., oxygen must be present at least. Electrochemical power is an additional type of external energy supply, but has been omitted due to dependency on electrode material. c Het., heterogeneous. b
Ionizing irradiation
Advanced Oxidation Processes at 107 s after electron injection (Buxton et al., 1988):
H2 O
Irradiation
0.28 • OH + 0.27 e aq + 0.06 • H
ð1Þ
+ 0.07 H 2 O 2 + 0.27 H 3 O + + 0.05 H 2 Mainly, three types of radicals are formed in parallel: (1) OH radical, (2) solvated aqueous electron eaq , and (3) hydrogen atom dH. The dOH radical can be generated directly by decomposing of electronically excited water (reaction (2)) or by producing and subsequent decomposing of a water radical cation (reactions (3) and (4)) (von Sonntag, 2006). By decomposing water radical cation an electron is generated, which becomes a solvated electron eaq , the most powerful reducing agent with reduction potential of 2.77 V (Mincher and Cooper, 2003). It should be noted that Hd is the conjugate acid of hydrated electron eaq with pKa (Hd) ¼ 9.1 (Buxton et al., 1988):
d
H2 Oþ -d OH þ Hþ
ð2Þ
e þ nH2 O-eaq
ð3Þ
H2 O -d OH þd H
ð4Þ
The products of reactions (2)–(4) subsequently act as reactants leading to a summarized reaction (1) at a given time shortly after electron injection. A series of different reactions including their second-order rate constants describing pure water radiolysis is given by Buxton et al. (1988). They also give a direct comparison of second-order rate constants of selected organic chemicals for dOH radical, solvated aqueous electron eaq , and hydrogen atom dH. Some additional important aspects of radiation chemistry, physics of energy absorption, as well as properties of dH and solvated electron can be found briefly in von Sonntag (2006). Some mechanistic aspects of solvated electron reductions (based on eaq generation by Birch reduction) as an alternative for waste remediation are described in Getman and Pittman (2003). It should be noted that radiation technique has also been widely used to clarify basics of free-radical chemistry, for example, of DNA damage (von Sonntag, 2006). Vacuum UV (VUV) irradiation. VUV irradiation is a photochemical process characterized by the low wavelength of emitted UV irradiance below 200 nm, for example, at 185 nm by low-pressure Hg lamps and at 172 nm by Xenon excimer lamps. At these wavelengths, direct photolysis of water takes place:
H2 O þ hnð1727 12 nmÞ-d OH þd H
ð5Þ
The penetration depth of VUV irradiation in water is reduced to a thin layer of maximum 70 mm depth at 172 nm. In the layer, the oxygen is rapidly depleted by peroxyl radical reactions (described later hereinafter) and reactions of Hd with molecular oxygen forming HO2 d radical. Although the quantum yield of water homolysis (reaction (5)) is relatively high at a value of 0.42 for 172 nm (Heit et al., 1998), the process is not effective for high total organic carbon (TOC) concentrated
379
wastewater (Legrini et al., 1993). Increase of process performance is only possible by improved mass transfer of oxygen into the layer. Otherwise this process is limited to the treatment of water and wastewater containing relatively low concentrations of pollutants (Legrini et al., 1993). Cavitation. Hydrodynamic and acoustic cavitations are the principal cavitation types of concern. Hydrodynamic or hydraulic cavitation occurs at lower frequencies below 20 kHz and is produced by pressure variation in a flowing liquid caused by velocity variation. Acoustic cavitation ranges from 20 kHz to 1 MHz, and is a result of pressure variation in a liquid caused by ultrasound. Ultrasound irradiation provides acoustical waves in the irradiated fluid by means of electromechanical transducer. Cavitation is produced in the rarefaction cycle of acoustical wave and the cavitation bubbles disappeared upon the next compression. The phenomenon of interest in sonochemistry is the transient cavitation: once a slow growing bubble is produced, the bubble will become unstable after a number of cycles and its size increases dramatically followed by a fast collapse. During the quasi-adiabatic collapsing phase, the temperature and pressure in the bubbles increase up to several thousand degrees (Flint and Suslick, 1991; Tauber et al., 1999) and several hundreds of bar, respectively. Under such extreme conditions, the water molecules undergo thermal dissociation to yield hydroxyl radicals and hydrogen atom radicals:
H2 OðgÞ-d HðgÞ þd OHðgÞ
ð6Þ
During acoustic cavitation the bubble collapses with higher intensity compared to hydrodynamic cavitation and, therefore, the temperatures and pressures during acoustic cavitation are higher. Once produced radicals, radical recombination reactions occur in both gas and liquid phase and by far the main products of water sonolysis are hydrogen (H2) and hydrogen peroxide (H2O2) (Destaillats et al., 2003). Hydrogen peroxide then can cause many secondary reactions. The reaction of free radicals released by collapsing bubbles with both volatile and nonvolatile water pollutants is therefore always a secondary reaction from the energy balance point of view (Destaillats et al., 2003). This may change for disinfection purposes using the more energy efficient, but also less radicalproducing hydrodynamic cavitation since those systems are widely established. The generation with the addition of oxidants is as follows: The main oxidants available for hydroxyl and other free radical generation without the use of catalysts are hydrogen peroxide and ozone. Also, chlorine and permanganate can be used for generation of free radicals, but these are not treated in this chapter. Some properties of ozone and hydrogen peroxide are summarized in Table 2. Ozone. Ozone cannot be stored because it decomposes to oxygen after generation. As a consequence of this, ozone must be produced onsite by ozone generators, which commonly generate gas streams containing ozone. For application of ozone to water and wastewater treatment gas–liquid contact reactors are necessary; thus, ozonebased AOPs always involve the efficiency of the gas–liquid transfer of ozone, especially due to the relatively low ozone
380
Advanced Oxidation Processes
solubility in water and the deriving mass transfer limitations. The influence of gas–liquid transfer on ozone reaction rates is described later in the text. The ozonation process is a well-known water-treatment process and widely used on a large-scale basis in drinking water treatment due to the combined activity as a disinfectant, strong oxidant, and discoloration agent. However, the process remained as black box for a long time with empirical optimization of process parameters, for example, the ozone dose transferred to water. Recently, ozonation is much better understood and therefore defined as an intrinsic AOP, because free radical reactions are always involved in ozonation of natural water or wastewater. This has been proved by different studies (e.g., Elovitz and von Gunten, 1999; Buffle et al., 2006a, 2006b). During ozonation of natural water or wastewater, pollutants are oxidized via two reaction pathways (Hoigne and
Table 2 Some properties of established oxidants – hydrogen peroxide and ozone Property
Unit
Molar mass Boiling point at 1013 mbar Melting point at 1013 mbar Density at 1013 mbar, 0 1C Max. limit for ambient level Odour threshold UV absorption at 253.7 nm Solubility in water at 0 1C Acidity (pKa)
Oxidants
g mol1 K K kg m3 ppm ppm 1/(M cm) gl1
Ozone value
H2O2 value
48 161.5 80.6 2.14 0.1 0.02 3300 1.05
34 423 273 1.46 Not relevant Not relevant 18.6a Miscible 11.6
a
UV absorption for conjugate base HO2 : 240 1/(M cm).
Bader, 1979; Elovitz and von Gunten, 1999): 1. Direct oxidation of water compounds with ozone. Ozone is a selective oxidant reacting directly with inorganic (e.g., Fe(II), Mn(II), NO2 , HS, As(III)) and organic pollutants (e.g., double bonds, amines, sulfur containing compounds, and activated aromatic rings) often at lower reaction rates compared to dOH radical reaction. 2. Oxidation of water compounds via dOH radical generation. In contrast to direct ozone action, this highly reactive free radical reaction is reacting with almost organic pollutants.
The value of contribution of each pathway depends on several factors such as ozone dose, scavenging capacity of wastewater matrix, pH, etc., and may change during oxidation process. Therefore, the ozonation process has been subdivided into two phases, the (1) initial phase with rapid decomposition of ozone, for example, first 20 s in natural water and (2) the second phase of ozone decrease as shown by Buffle et al. (2006b) (see Figure 1). Recently, three phases have been identified No¨the et al. (2009a) by subdividing the first initial phase into a very fast ozone decomposition phase for first 1 mg of total organic carbon (TOC) corresponding to B0.3-s reaction time. The transformation of O3 into dOH radical may yield 50% (Buffle and von Gunten, 2006) and depends on present reactive moieties of natural organic matter (NOM), for example, carbonates, amines, and phenols. The mechanism of radical generation in natural water as proposed by Buffle and von Gunten (2006) is based on direct superoxide formation by amines and additionally on a direct e-transfer and the oftenreported subsequent ozonide anion radical reaction by phenols (see reactions (7) and (8)). However, it is most important to note that reaction (8) is not correct. The recent knowledge about subsequent reactions is described by reactions (26)–(29), indicating that the yield of dOH radical generation
60
60
Ozone (μm)
60 50 40 30 20 10 0
50 40 0
30
3
6
9
Ozone (μm)
CQFS
40
Ozone (μm)
50
12
Time (s)
20
20
10
10
0
0 0
(a)
30
300
600 Time (s)
900
1200
0 (b)
5
10 15 Time (s)
20
25
Figure 1 Ozone stability in (a) natural water and (b) wastewater. CQFS, continuous quench-flow system. Reprinted from Buffle et al. (2006b) Measurement of the initial phase of ozone decomposition in water and waste water by means of a continuous quench-flow system: Application to disinfection and pharmaceutical oxidation. Water Research 40: 1884–1894, Copyright (2006), with permission from Elsevier.
Advanced Oxidation Processes
381
0.6
kO3 = 1 M–1 s–1 10 M–1 s–1
0.4
River Sihl, 15 °C, pH 8
Fraction P reacting with •OH
0.8
Lake Zürich, 15 °C, pH 8
Porrentruy, 10°C, pH 7.2
1
100 M–1 s–1 1000 M–1 s–1
0.2
0 10–10
10–9
10–8 Rct value
10–7
10–6
Figure 2 Fraction of micro-pollutant P reacting with dOH as a function of Rct. From Elovitz and von Gunten (2006) Hydroxyl radical/ozone ratios duringozonation process. I. The Rct concept. Ozone: Science and Engineering 21(3): 239–260, reprinted by permission of Taylor & Francis Group, http://www.informaworld.com.
is lower:
O3 þ PhO -O3
d
þ PhO
d
ð7Þ
O3 d þ Hþ -d OH þ O2 ðincomplete reaction summary of different reactionsÞ ð8Þ The transient dOH concentrations during initial phase calculated from wastewater ozonation have been found 100 times higher than those occurring in natural water AOP process H2O2/O3 (Buffle et al., 2006c) and therefore, the addition of H2O2 does not seem useful in this initial phase, but maybe later, depending on the water contaminants. The contribution of each pathway to the oxidation of pollutants can be determined by the Rct concept introduced by Elovitz and von Gunten (1999). The idea behind the Rct concept is the use of an dOH-probe. A suitable dOH-probe consists of very low reactivity with O3 and high reactivity with dOH (e.g., p-chlorobenzoic acid (pCBA)); its disappearance during ozonation is an indirect measure of dOH concentration. The Rct value represents the ratio between concentrations [dOH] and [O3] at any time and can be determined in laboratory batch system (Elovitz and von Gunten, 2006) or more accurately over subsecond timescales by continuous quench-flow system (CQFS; Buffle et al., 2006b). The calculated dOH concentration further allows the calculation of O3 and dOH radical exposure to a wastewater by considering the measured ozone decomposition rate during ozonation. With the calculated exposures and the second-order rate constants kO3 and kOH for direct ozone and dOH radical oxidation, respectively, the extent of oxidation of a compound can be predicted for batch reactions as well as for plug-flow reactors by
Z t ½P ¼ exp ½O3 dtðkO3 þ kOH Rct Þ ½P0 0
ð9Þ
with [O3] as time-dependent function (Elovitz and von Gunten, 2006). It is worth noting that Rct changes with reaction time and is only constant within reaction phases, which depend on wastewater characteristics. Therefore, the reaction phases of the related Rct values should be identified before. While second-order rate constants k have been determined for many compounds (see, e.g., database in NIST (2009)), the time-dependent functions of [O3] and [dOH] exposures need to be determined for each wastewater. Assuming the kOH rate constant of 5 109 M1 s1 for dOH radical reactions, the fraction of a compound reacting with d OH radical can be calculated as a function of Rct value for different ozone reaction rate constants kO3. Summarizing these calculations in a figure of Elovitz and von Gunten (1999) (see Figure 2), the importance of dOH radical oxidation during ozonation has been elucidated as significant for all compounds with kO3o104 M1 s1. Hence, wastewater ozonation can generally be categorized as an ozone-based AOP. Influence of gas–liquid transfer on ozone reaction rates. Bubble columns or similar reactors are commonly used for ozone treatment. In these reactors, the overall reaction rate depends on both the transfer rate of ozone from the gas bubbles to the liquid and the rate of reaction of ozone with water pollutants and other species. This process can be described theoretically based on the mathematical model of film theory (Lewis and Whitman, 1924). A model of a gas–liquid reactor considering fast and slow reactions is described, for example, in Benbelkacem et al. (2004), involving ozone as gas component A and nonvolatile pollutant as liquid component B. Concentration profiles at the interface of gas and liquid can be predicted by this model by solving the mass balances within the film using finite difference Runge–Kutta method. A typical gas–liquid concentration profile of liquid film model is illustrated in Figure 3 for a common type of a bubble column.
382
Advanced Oxidation Processes
Cg C
Cg,out
∗
E
1
C ∗1
C 2,bulk
C 2,bulk
D Gas
Liquid film
C1,bulk
C 1,bulk Bulk liquid Cg,in
Figure 3 Sketch of semibatch gas–liquid reactor (left hand) and typical concentration profile predicted by liquid film model (right hand).
Different reaction regimes of ozone reactions have been identified and the Hatta number Ha was defined to describe the ratio between kinetic and diffusion regime (Equation (10)). In addition, the enhancement factor E (Equation (11)) and the depletion factor (Equation (12)) have been introduced to describe the flow rates of gas component at gas– liquid interface:
pffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi k CB;bulk DA Ha ¼ ; kL dCA dx x¼0 E¼ kL ðCA CA;bulk Þ
ð10Þ
DA
dCA dx x¼L D¼ kL ðCA CA;bulk Þ
ð11Þ
DA
ð12Þ
with kL being the mass transfer coefficient (m s1), k the reaction rate constant (m3 mol1 s1), DA the diffusivity of component A (m2 s1), CA the dissolved gas concentration within the film (mol m3), and CA the equilibrium concentration at the gas–liquid interface (mol m3). By calculating E and D the gas flow rate reacting within the film can be calculated as the difference between flow rate entering and leaving the liquid film:
NA;film ¼ ðE DÞ kL S ðCA CA;bulk Þ
ð13Þ
where NA,film is the reaction rate within the liquid film (mol s1) and S the interfacial surface (m2). Results by Benbelkacem et al. (2004) on model calculations for a semi-batch ozonation process taking into account the overall mass balances and second-order reaction type show that four typical kinetic regimes can be observed: (1) the slow kinetic regime (Hao0.02), where the reaction occurs in the bulk liquid and E and D are equal to 1; (2) the specific slow kinetic regime (0.02oHao0.3), where the reaction is negligible within the liquid film and E and D are still B1, this condition is usual to ones for determining mass transfer coefficients; (3) the intermediate kinetic regime (0.3oHao3), where the reaction occurs in both the bulk and film liquid and the mass transfer is accelerated by the chemical reaction (E
and D are not equal to 1); and (4) the fast kinetic regime (Ha 44 3), where the reaction occurs in the liquid film, D is close to zero, while E obtained a high value. Details of concentration profiles and the mass transfer limitations are discussed in Benbelkacem and Debellefontaine (2003) and Benbelkacem et al. (2004). From these discussions, the kinetic regime of batch ozonation processes can be identified by the progress of ozone concentration in off-gas. Ozone/UV process. Free radical generation during ozonation will be accelerated by UV light irradiation. Generally, different ranges of UV light as shown in Figure 4 can be applied: (1) long-wave UV-A, also known as near-UV light radiation, (2) middle-wave UV-B, and (3) short-wave UV-C, also known as far-UV light radiation. For ozone/UV processes, the UV-C range is of main interest due to the high absorbance of ozone in this range. Figure 4 also indicates different typical UV radiation spectra of commonly used low-pressure, low-intensity and mediumpressure, high-intensity mercury lamps together with the relative UV adsorption for DNA. The most effective UV light wavelengths for microbial inactivation are considered to between 255 and 265 nm. The UV light is commonly generated by striking an electric arc between two electrodes in a specially designed lamp, which contains a vapor or gas mixture (e.g., mercury vapor). The emission of UV is based on the energy generated by the excitation of mercury vapor or gas mixture. The initiating reactions of free radical generation are proposed to be a two-step process of light-induced homolysis of ozone:
O3 ðhno 310 nmÞ-O2 þ Oð1 DÞ
ð14Þ
Oð 1 DÞ þ H2 O-d OH þd OH
ð15Þ
Although the UV-C absorbance of ozone is much stronger than that of hydrogen peroxide, the dOH quantum yield is very low in the range of 0.1 (Reisz et al., 2003). A recombination of OH radicals (Equation (16)) in the solvent cage as well as a photolysis of ozone dissolved in water (Equation (17)) leads to production of hydrogen peroxide which has been verified by Peyton and Glaze (1988) as the main primary product of ozone photolysis: d
OH þ d OH-H2 O2
O3 þ H2 O-H2 O2 þ O2
ð16Þ ð17Þ
Advanced Oxidation Processes
Cosmic rays
Gamma rays
X-rays
Ultraviolet
Visible light
383
Radio waves
Infrared
10–3 m
Relative lamp output
100 nm
Short-wave UV (UV-C)
200
Middle-wave Long-wave UV UV (UV-A) (UV-B) 280
400 nm
315
15
0.6
10
0.4
5
0.2
0 220
240
260
280 300 Wavelength (nm)
320
Relative DNA absorbance
Vacuum UV
0 340
Typical low-pressure low-intensity UV lamp Typical medium-pressure high-intensity UV lamp Figure 4 Range of UV radiation and radiation spectra of different UV lamps.
Further OH-radical generation is proposed by photolysis of hydrogen peroxide and by reaction of ozone with HO and HO2 . Hydrogen peroxide (H2O2). Hydrogen peroxide is a commercially available, thermally stable chemical of acceptable on-site storage capability. It is not only an oxidizing species as aforementioned in Table 1, but also acting as a reductant with E1 ¼ 0.7 V. Compared to the use of the other common oxidant, ozone, limitation of solubility in water as well as mass transfer can be neglected and hydrogen peroxide forms a weak acid in water by dissociating to yield the peroxide anion, sometimes called as hydroperoxide:
H2 O2 $ HO2 þ Hþ ðpKa ¼ 11:7Þ
ð18Þ
H2O2/UV process. The photolysis of hydrogen peroxide decomposes the molecule into two dOH radicals by Equation (19), as commonly agreed. UV-C irradiation must be used for photolysis due to the UV absorption of H2O2 in this range:
H2 O2 þ hn-½2d OHcage
ð19Þ
The OH radicals undergo two pathways: (1) recombination in the cage and (2) diffusion out of the cage; therefore, the yield for free OH radicals is 50%. The reaction rate of photolysis depends on pH (see, e.g., Nicole et al., 1990) with increasing rates at higher pH values. A possible explanation of this effect is the higher molar absorption coefficient of peroxide anion at 253.7 nm (Glaze et al., 1987). H2O2/O3 process. This process has been discovered by Staehelin and Hoigne´ (1982), and is sometimes called perozone or peroxone process. The original reaction mechanism concept of Staehelin and Hoigne´ (1982), has been slightly modified by Sein et al. (2007) and recently revised by Mere´nyi et al. (2009). According to Staehelin and Hoigne´ (1982), the reaction mechanism is described by following six reactions ((20)–(25)) expressing that 2 mol OH-radicals may be theoretically generated by 1 mol H2O2 and 2 mol O3. Due to recombination reactions of radicals (e.g., O3 d þ HO2 d ), the yield should always be lower in practice.
H2 O2 $ HO2 þ Hþ HO2 þ O3 -HO2 d þ O3 d ðelectron transferÞ þ HO2 d $ Od 2 þ H pKa ¼ 4:8
ð20Þ ð21Þ ð22Þ
384
Advanced Oxidation Processes O3 d þ Hþ $ HO3 d
ð23Þ
HO3d -dOH þ O2
ð24Þ
HO2 d þ H2 O2 -d OH þ H2 O þ O2
O2d þ O3 -O2 þ O3d
ð25Þ
Iron undergoes a catalytic redox cycle reaction changing its oxidation state between þ II and þ III (reactions (30), (31), (34), and (35)), while dOH scavenging by reaction (33) is expected to be minimal because generation of dOH is catalytic in iron (Pignatello et al., 2006). Although the hydroxyl radical is generated by reaction (30), both reactions (30) and (31) may initiate the sequence of reactions depending on oxidation state of iron, initially added to the process. The reaction with ferrous iron by reaction (30) is several orders of magnitude faster than the reaction with ferric iron by reaction (31). This results in a slow reaction rate limited by reaction (31) after quick and complete oxidation of initially added Fe(II). The reaction rate is thus independent of type of initially added iron. Different speciation of both ferrous and ferric iron as a function of pH must be considered to understand the reaction mechanisms in detail. The ferrous and ferric irons predominantly exist in acidic solutions as hexaquo ion Fe II ðH2 OÞ6 2þ and Fe III ðH2 OÞ6 3þ , respectively, and undergo hydrolysis depending on pH (for ferrous iron: Wells and Salam, 1965, 1968; for Ferric iron: Gallard et al., 1999; summarized in Pignatello et al. (2006)). Due to the complexity of Fe(III) hydrolysis and its high impact on reaction rates, great care is required to obtain well-defined iron salt solutions. To ensure this, some advice for investigators from Pignatello et al. (2006) is given hereinafter: (1) dissolution of ferric salts in neutral water immediately starts hydrolysis and, therefore, concentrated stock solution should be prepared to below 0.1 M and diluted in acidified water Z0.1 M Hþ; (2) the total iron should be kept below 1 104 M in acidic solution less than 102 M Hþ; (3) locally high pH should be prevented for adjusting pH in the acidic range and use of bicarbonate is recommended rather than hydroxide solution; and (4) interferences with colloidal oxides should be avoided and solutions should be used within few hours after preparation. Hydrolyzed species can be detected by turbidity and slight yellow–orange color. The optimum pH values for pollutant oxidation rates have been found at approximately pH 3. This is due to the fact that the rate-limiting step of reaction (33) is due to precipitation of catalyst Fe(III) above pH 3, while the optimum pH of the reaction (32) is around pH 4 with a reaction rate 7 times higher than at pH 3 (Pignatello et al., 2006). Due to the catalytic character of iron reactions, the concentration of iron can be reduced to minimum amounts of ppm of the wastewater volume depending on pollutant type and amount. The peroxide-to-iron molar ratio is usually in the range of 100–1000. It should not be forgotten that not only iron is able to generate radicals. Other transition metals such as copper, nickel, cobalt, and chromium may also undergo the so-called Fenton-type reactions (Goldstein et al., 1993; Koppenol, 1994; von Sonntag, 2006). Photo-assisted Fenton reaction.The Fenton reaction may be improved by additional application of UV, UV/visible light, and near-infrared irradiation. The improvement relates to
Reactions (25) and (26) have been extended by reactions (26)–(28) (Mere´nyi et al., 2009) due to the reinvestigation of theoretical OH radical yield, which is only half of formerly assumed value for ozone (von Sonntag, 2008). Thermo-kinetic and quantum-chemical calculations give rise to the suggestion of short-lived nonradical adduct formation of HO4 by reaction (28) and subsequent dismutation by reaction (29) and HO2 d =O2 d radical reactions (Mere´nyi et al., 2009) taking into account the aspect that the kinetics of free radical chemistry often over-runs thermodynamics:
O3 d $ Od þ O2
ð26Þ
O d þ H2 O$ d OH þ OH
ð27Þ
OH þ O3 -HO4 ðadduct formationÞ
ð28Þ
HO4 $ HO2 d þ Od 2
ð29Þ
Homogeneous Fenton reaction. The Fenton reaction is one of the first intensively studied AOP processes (Fenton, 1894; Haber and Weiss, 1932, 1934; historical review in Koppenol (1993)), which has been already used four decades before to enhance the reaction of hydrogen peroxide with I (see Scho¨nbein, 1857; mentioned in von Sonntag (2006)). The reaction is iron based using different oxidation states of iron. A detailed review and description of fundamental chemistry of Fenton and photo-Fenton was given by Pignatello et al. (2006) and an overview related to solar-based photoFenton can be found in Malato et al. (2009). The classical mechanism for Fenton reaction has been introduced by Haber and Weiss (1932, 1934) and revised by Barb et al. (1949, 1951a, 1951b) and contains a sequence of seven reactions:
FeðIIÞ þ H2 O2 -FeðIIIÞ þ OH þd OH
ð30Þ
FeðIIIÞ þ H2 O2 -FeðIIÞ þ HO2 d þ Hþ
ð31Þ
OH þ H2 O2 -HO2 d þ H2 O
ð32Þ
OH þ FeðIIÞ-FeðIIIÞ þ OH
ð33Þ
FeðIIIÞ þ HO2 d -FeðIIÞ þ O2 Hþ
ð34Þ
FeðIIÞ þ HO2 d þ Hþ -FeðIIIÞ þ H2 O2
ð35Þ
HO2 d þ HO2 d -H2 O2 þ O2
ð36Þ
d
d
In addition, the conjugate base of HOd2 , the superoxide anion O2 d undergoes reactions analogous to reactions (34)– (36). Another reaction for hydroxyl radical generation mentioned in different papers (reaction (37)) can be neglected due to extremely low reaction rate constant (k37 ¼ 3 M1 s1;
Koppenol et al., 1978):
ð37Þ
Advanced Oxidation Processes
faster reactions as well as to higher yield of inorganic products (Pignatello, 1992; Kiwi et al.,1994; Lei et al., 1998; De Laat et al., 1999). Due to the photochemistry of Fe(III) an enhancement of the rate-limiting step of Fenton process (reaction (31)) is achieved. Fe(III) complexes are usually present in water and wastewater and involve ligands of any Lewis base (e.g., OH, H2O, HO2 , Cl, R–COO, RNH2, and R–OH). Due to absorption of photons the complexes undergo a ligand-to-metal charge transfer (LMCT) excitation and a subsequent dissociation reaction to Fe(II) and an oxidized ligand:
Fe III ðLÞn þ hn-ðFe III ðLÞnÞ -Fe II ðLÞn1 þ Lox
ð38Þ
The advantages of photo-assisted Fenton to classical (thermal) Fenton are manifold (Pignatello et al., 2006): (1) the reduced iron can undergo reaction with hydrogen peroxide to yield dOH radical (reaction (30)); (2) the oxidation of ligand may lead to further degradation of pollutants; and (3) Fe(III)-hydroxy complexes present in mildly acidic solutions of pH 3–4 may undergo photo-reduction to form dOH radicals directly (reaction (39)):
FeðOHÞ 2þ -Fe 2þ þ dOH
ð39Þ
The yield of reactions (38) and (39) depends on wavelength and light absorption properties, which are also affected by water contaminants , the so-called inner filter effects as well as of ligand types (Faust and Hoigne, 1978; Benkelberg and Warneck, 1995). It is worth noting that visible light/sunlight has been identified as an appropriate polychromatic radiation source, since it is able to enhance photolysis of ferric iron complexes by overcoming inner filter effects (Gernjak et al., 2003; Oliveros et al., 1997). Also, the increase of temperature up to 50 1C reduces the demand of hydrogen peroxide (Pignatello and Day, 1996) and/or enhances the reaction rate (Sagawe et al., 2001; Go¨b et al., 2001) and may increase up to 5 times (Gernjak et al., 2006), while the optimum has been found at 55 1C due to improved hydrogen peroxide consumption at higher temperatures (Torrades et al., 2003). As a consequence, insulated and hybrid reactors have been developed (Sagawe et al., 2001; Farias et al., 2010) allowing both photochemical and thermal solar irradiation.
4.13.2.1.2 Heterogeneous processes The use of solid catalysts in water and wastewater treatment is based on the science of heterogeneous catalysis involving five reaction steps: (1) diffusion of reactants to the surface of catalyst, (2) adsorption of reactants onto the surface, (3) reaction on the surface, (4) desorption of products off the surface, and (5) diffusion of desorbed products. Noble metals (e.g., Ir, Pd, Pt, Rh, and Ru) and metal oxides of different metals such as Cu, Mn, Co, Cr, V, Ti, Bi, and Zn have been commonly used as heterogeneous catalysts. The catalysts are immobilized on supports, which may be classified by their nature to inorganic and organic supports. The tasks of supports are to: (1) increase the surface area of catalytic material, (2) decrease sintering, and (3) control useful lifetime of catalysts (e.g., Matatov-Meytal and Sheintuch, 1998).
385
Materials such as carbon black, metal oxides (e.g., TiO2 and Al2O3), silica, zeolite, glass and carbon fibers, ceramic materials, pillared clays (Al–Cu and Al–Pt), and many others have been used as supports for catalysts. Heterogeneous catalysis is currently a dynamic field driven by new developments, for example, in catalyst preparation, immobilization techniques, surface modification techniques, and, last but not least, the use and characterization of nanosized particles. However, heterogeneous catalysis in water phase is, to date, of limited application due to unsolved problems in the deactivation of catalysts by poisoning, sintering, or leaching. These problems have still to be overcome to ensure suitable lifetime of catalysts and economic applications. Some processes currently receiving more interest are the heterogeneous photo-Fenton process and the semiconductor TiO2 photo-catalysis due to the possible use of solar radiation, and the electrochemical oxidation due to recent advances in electrode material development, for example, doped diamond electrodes. Semiconductor photo-catalysis. Photo-catalysis is a process that generally may involve photons (by UV or solar radiation) and a catalyst either in liquid (homogeneous photo-catalysis) or in solid phase (heterogeneous photo-catalysis). The most investigated semiconductor photo-catalyst is TiO2 in anatase form. It is an inexpensive mass product, chemically and biologically inert and resistant to chemical and photo-corrosion. Other semiconductors of lower importance are, for example, a-Fe2O3, SrTiO3, WO3, ZnO, and ZnS. The initial step of photo-catalysis is the absorption of photons by the catalyst. UV-irradiation of TiO2 with wavelength o380 nm leads to an excited state, which is commonly explained by the band-gap model illustrated in Figure 5: through absorption of photons the electrons in the valence band of the semiconductor TiO2 are transferred to the valence conduction band, thus creating electron vacancies, also called as electron deficiencies or holes. Electron–hole pairs develop (Equation (40)) and the electron-depleted valence band hole (hþ) has a high reduction potential of 2.9 V versus normal hydrogen electrode (NHE) for oxidizing most of the pollutants present in wastewater:
TiO2 þ hn-TiO2 ðe þ hþ Þ
ð40Þ
An electron transfer to the valence band hole (hþ) either from the adsorbed substrate (Fox et al., 1991) or from the adsorbed solvent molecules (H2O and OH; see Pichat, 1991) has been identified for generating radicals. In water-treatment processes, the adsorbed H2O and OH have probably higher impact on radical generation by Equations (41) and (42), respectively, due to their higher concentration:
TiO2 ðhþ Þ þ H2 Oad -TiO2 þdOHad þ Hþ
ð41Þ
TiO2 ðhþ Þ þ OHad -TiO2 þdOHad
ð42Þ
Molecular oxygen can act as an acceptor of the valence band electron (VBe) released by the excited TiO2 valence band electron hole (hþ) and enables further radical generation by forming the protonated form of superoxide anion
386
Advanced Oxidation Processes Acceptor Reduction
hν
O2
Adsorption
Acceptor
Energy
e–
Conduction band
Conduction band
Eg Eg
e–
•–
O2
Change carriers formation
Recombination
Donor+ Valence band
h+
Valence band
Oxidation
Further degradation P•• H+ +
h+
Donor
Reduction
P
OH•
H2O
Oxidation Adsorption
(Equation (43)) and subsequent dismutation to hydrogen peroxide or peroxide anion. Then, further dOH radicals are generated by dismutation of hydrogen peroxide or by reaction of Equation (44) with hydrogen peroxide as further electron accepting species. Enhancing the trapping of electrons and holes before recombination is one of the keys for optimizing TiO2 photo-catalysis. Electrons can be trapped within 30 ps after excitation and holes within 250 ns (Rothenberger et al., 1985), while interfacial charge transfer takes place in a period of nanoseconds to milliseconds (Martin et al., 1994):
VBe þ O2 -O2 d
ð43Þ
VBe þ H2 O2 -OH þdOH
ð44Þ
The reaction rate of photo-catalysis may be enhanced by the addition of hydrogen peroxide (Ollis et al., 1991; Matthews, 1991), but it is still pH dependent due to changes in adsorption/desorption rates and electron–hole separation efficiencies. The advantage of TiO2 photocatalysis is the potential use of solar radiation due to the spectral absorption characteristic of TiO2 in UV-A regions (see Figure 6). This may receive more interest in future due to more sustainable energy source and possible reduction of carbon dioxide emission. However, photo-catalysis by TiO2/UV is well known as a very slow AOP compared to other AOPs, which is mainly directed to the very low quantum yield (e.g., Ishibashi et al., 2000), even more in the case of solar radiation, where not only the spectral absorption of TiO2 is relatively low but also the energy content of solar radiation is less compared to UV-C radiation. This problem may be rather negligible for the removal of trace pollutants due to very low concentrations resulting in reduced reactor dimensions, which may be economic. On the other hand, the use of other semiconductor materials (e.g., SnO2, ZnO, WO3, GaAs, and GaP) is under investigation due to their higher band-gap wavelength enabling higher absorption of solar radiation and the possible increase of yielding photons. Another option under investigation is the
Arbitrary units
Figure 5 Scheme of band-gap model of semiconductor TiO2 with electron–hole pair. Redrawn from Malato S, Ferna´ndez-Iba´nez P, Maldonado MI, Blanco J, and Gernjak W (2009). Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catalysis Today 147: 1–59, Copyright (2009), with permission from Elsevier.
250
TiO2 Solar spectrum
300
350 400 Wavelength (nm)
450
500
Figure 6 TiO2 absorption spectrum compared with solar radiation spectrum. Redrawn from Malato S, Ferna´ndez-Iba´nez P, Maldonado MI, Blanco J, and Gernjak W (2009) Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catalysis Today 147: 1–59, Copyright (2009), with permission from Elsevier.
doping of semiconductor materials, for example, by Cr, to shift the absorbance to higher wavelengths of visible light, the so-called red-shift. Also, the doping with nonmetallic elements such as sulfur, nitrogen, and carbon could be adopted in order to extend the absorbance wavelength range of TiO2 and enhance photo-catalytic activity (Thompson and Yates, 2008). However, problems of deactivation, leaching, or chemical stability have still to be overcome.
4.13.2.2 Reaction Mechanisms Once generated an dOH radical, it reacts very fast, that is, often close to diffusion-controlled rates (Buxton et al., 1988). The radical reactions can be classified to (1) addition reaction, (2) hydrogen abstraction, and (3) electron transfer. The addition reaction is the preferred pathway, if possible. The yield of different reactions in relation to dOH yield depends on the moieties present in water, for example, (1) hydrogen abstraction may be negligible when C ¼ C and C ¼ N double bonds are present (So¨ylemez and von Sonntag, 1980) and
Advanced Oxidation Processes
(2) electron transfer yield increases, when the potential sites of d OH radical addition are substituted by halogens and may dominate for Br (see hereinafter). Despite the fact that radical reactions described in this chapter are mainly addressed to dOH, not only dOH radicals may be involved in free-radical reactions for water and wastewater treatment. Also, inorganic, carbon-centered, heteroatom-centered, or peroxyl radicals can be involved depending on specific compounds in polluted water. Some additional details about radical formation and compiled reaction rates of these radicals can be found in Buxton et al. (1988) and von Sonntag (2006). However, overall reaction rates of AOPs derived from wastewater and water-treatment investigations are commonly traced back to dOH reaction rates. This should be kept in mind, for example, for possible interpretation of results in wastewater treatment by AOPs. Another fact not treated in this chapter is the possible change of redox property of the radical from reducing species to oxidizing ones (readers are referred to, e.g., von Sonntag, 2006). One of the main difficulties in application of AOP is based on the influence of radical scavengers in the (waste)water matrix. Scavengers could inhibit the pollutant degradation reactions, and the predominant effect of scavengers in a given wastewater depends on all involved moieties of compounds in the water. One important scavenger is the carbonate in its different dissociation forms, mostly existing in natural water and wastewater. Other scavengers many times present in these water types are NOMs often called humic substances, whereas phenol and amine are present as repeating and promoting moieties of NOM. Addition reactions. The hydroxyl radical reacts readily with C ¼ C (reaction (45)), C ¼ N, and S ¼ O (except SO4 2 , reaction (46)) double bonds, and with transition metal ions (reaction (47)) by addition reaction leading to an adduct formation and subsequent adduct decomposition. Due to its electrophilic character, electron-rich positions at carbon atom are preferably attacked, for example, the C5 ¼ C6 double bond is rather attacked at C5 than at C6, due to the electron-richer region at C5 (von Sonntag, 2006): H + OH
H OH
ð45Þ
the H-abstraction chain reaction scheme is often used to explain oxidation of organic compounds, it is important to know that peroxyl radicals are known as weak oxidants due to its low reduction potential and may act as one-electron oxidants toward strong electron donor compounds. They, therefore, do not preferably undergo an auto-oxidation chain reaction at common water-treatment AOP conditions, but the elimination of superoxide radical (reaction (51); see von Sonntag and Schuchmann, 1991; von Sonntag, 2006). The peroxyl radical chemistry is much more complex as peroxyl radicals undergo a number of unimolecular reactions, not only HO2 d =O2 d elimination reaction through the formation of a double bond but also addition to C ¼ C double bond, or O-transfer reactions (for compiled details, see von Sonntag (2006)):
HRH þ dOH-HR d þ H2 O
ð48Þ
HRd þ O2 -HROOd
ð49Þ
HROO d þ HRH-ROOH þ HR d ðnot at common wastewater conditionsÞ
ð50Þ
HROO d -HO2 d þ R
ð51Þ
From the above reactions, it is obvious that the presence of oxygen significantly influences the radical reaction system. If oxygen is absent or not sufficiently present, reaction (49) is of minor importance compared to other reactions and the radical recombination by reaction (54) prevails. On the other hand, a very high amount of molecular oxygen improves the relative proportion of bimolecular recombination (von Sonntag and Schuchmann, 1991). Most of the carbon-centered radicals react fast and irreversible with O2 following reaction (49), while other organic compounds do not so, for example, hexandienyl and thiyl radicals react only reversible with O2 and phenoxyl and other hetero-centered radicals do not react with O2 (von Sonntag, 2006). The rate of reaction cycles depends on the presence of inhibitors/scavengers. One of the well-known scavengers is carbonate in its different dissociated forms. The reactions of carbonate and bicarbonate lead to carbonate radicals, which are mostly known as poor reactants: d
R2 S ¼ O þd OH-R2 SðOd ÞOH-RSðOÞOH þ R d Tl 2þ þ d OH-HOTl 2þ -Tl 2þ þ H2 O
ð46Þ
ð47Þ
Hydrogen abstraction reactions. Based on the bond dissociation energy of HO–H higher than C–H bond, a hydrogen atom can be removed from organic compound HRH (Equation (48)), thus forming a carbon-centered radical dR. Many authors propose a chain reaction to be initiated by reaction of d R with molecular oxygen (Equation (49)) producing a peroxyl radical, which may subsequently react with other organic compounds by reaction (50), and ideally lead in its final step to carbon dioxide, water, and inorganic salts. However, since
387
OH þ HCO3 -H2 O þ CO3 d
ð52Þ
OH þ CO3 2 -OH þ CO3 d
ð53Þ
d
Electron transfer reactions. The addition reaction and the electron transfer are in competition and although the electron transfer is thermodynamically favored, the addition reaction is often preferred, while the direct electron transfer has been rarely observed (von Sonntag, 2006). Direct electron transfer is favored, for example, in reactions with halogenated phenolate ions, because the ortho- and para-position of preferred HOd addition, are blocked by substituents and the electron transfer might become dominant (von Sonntag, 2006). The yield of electron transfer contributing to the total reaction increases with the size of halogen and dominates for Br and I by 73% and 97%, respectively. The other reactions with
388
Advanced Oxidation Processes
preferred electron transfers are reactions with thiol and thiolate (Akhlaq and von Sonntag, 1987). Radical recombination reactions. The radical recombination reactions terminate the chain reactions by yielding combined molecule whereas larger molecules decompose, for example,
HROO d þ HROO d -ROH þ ROH þ O2
ð54Þ
4.13.2.3 Reaction Systems Reaction systems may be described by different reaction pathways and sequences. The complexity of these systems is exemplary, illustrated by H2O2/UV and ozone/UV–ozone/ H2O2 processes. The ozonation process is treated separately to the ozone/UV/H2O2 reaction system due to different phases of ozone decomposition by direct and indirect ozone action and the direct formation of the ozonide radical anion O3 d .
4.13.2.3.1 Competition kinetics The sequence of radical reactions may be most affected by promoting and inhibiting/scavenging compounds depending on the water/wastewater characteristics. Not only the reactions but also the dissociation equilibriums depending on pH and the progress of pH during oxidation must be taken into account. The same goes for heterogeneous catalysis where, for example, the charge of surface and subsequently the adsorption/desorption kinetics are affected by pH value. Besides (1) pH, some common parameters, which may scavenge the removal of target pollutants, are (2) carbonate/ bicarbonate, (3) dissolved oxygen, (4) nitrate ion, and (5) NOM, (6) OH, (7) Fe2þ, (8) primary and secondary alcohols, (9) phosphate ion PO3 4 , and (10) specific chemicals (pBA, tert-butyl alcohol, etc.). 1. The pH value has direct and indirect effects. A direct effect is that the hydrated electron eaq is scavenged by hydrogen ion at pH values below neutral pH to form the hydrogen atom. At pH higher than neutral, the dOH radical may dissociate to produce the hydrogen ion and oxygen anion ( d OH$ O d þ H þ ; pKa ðd OHÞ ¼ 10:8). Indirect effects are, for example, through the carbonate/bicarbonate concentration (see (2)) or the hydroxide anion concentration (see (6)). 2. The total carbonate concentration in water is part of the carbon dioxide buffer system in natural water and most types of wastewater due to dissolution of biologically generated or atmospheric CO2 to produce carbonic acid and the subsequent dissociation of carbonic acid. The scavenging effect of carbonate ion CO3 2 and bicarbonate ion HCO3 is described by reactions (54) and (55), where the carbonate ion is the more important hydroxyl radical scavenger due to 50 times higher reaction rate constant. Considering the carbonate–bicarbonate dissociation equilibrium of pKa value of 11.5, the dOH scavenging prevails at high pH, where most commonly a reduced reaction rate has been found compared to neutral and acidic pH, thus indicating an inhibition of radical reaction in pollutant degradation at higher pH. 3. The role of dissolved molecular oxygen is different for various AOPs. Mostly, it accelerates the reaction rates of
4.
5.
6.
7.
8.
photochemical and Fenton reactions (e.g., Sun and Pignatello, 1993; Kim and Vogelpohl, 1998; Bossmann et al., 2001; Legrini et al., 1993), and , in addition, serves as an oxidant. The oxygen incorporated into organic pollutants may originate partly from dissolved oxygen as shown by isotope-labeled 18O2 (Kunai et al., 1986); subsequently, dissolved oxygen is desirable to promote reduced consumption of hydrogen peroxide. In contrast to this, especially in radiolysis the oxygen will be rapidly reduced by solvated electron eaq and Hd to produce superoxide radical anion O2 d . The second-order reaction rate constants are very high at 1.9 1010 and 2.1 1010 M1 s1. The superoxide radical anion is a reducing species and relatively inert compared to hydrated eaq , and thus the ebeam process efficiency will be reduced by the presence of oxygen (Mincher and Cooper, 2003). In other AOPs the superoxide radical anion plays an important role, especially in ozone-based and ozone-involved AOPs because the relatively low reactive superoxide radical reacts fast with ozone resulting after sequence of fast reaction in generation of further dOH radicals. At high concentration levels, the nitrate ion NO3 can act as an electron eaq scavenger and, for example, for the ebeam process the product after a sequence of reactions is nitrite NO2 . This eaq scavenging may affect the recombination of eaq and dOH radical in radiolysis, thus enhancing the oxidative reactions with dOH (Mincher and Cooper, 2003). However, the nitrite ion generation can be compensated by ozone addition (Mincher and Cooper, 2003). The NOM could initiate or inhibit the target radical reactions depending on the type and quantity of different moieties of NOM. Humic acids as a part of NOM also act as initiator or inhibitor depending on its concentration (Xiong and Graham, 1992). Especially during ozonation, new electron-rich moieties may be produced by direct ozone reaction. These moieties are then involved in the generation of new free radicals No¨the et al. (2009b). Some important moieties of NOM are amines and phenols (Buffle and von Gunten, 2006). The hydroxide anion is often mentioned as a possible initiator for ozone-based systems, but it should be noted that the reaction rate is low (70 M1 s1) and therefore mostly negligible. OH can react directly with ozone forming the anion of hydrogen peroxide, which reacts also with ozone generating an ozonide anion radical O3 d and a superoxide radical HO2 d , both known as chain carriers for hydroxyl radical generation. Dissolved Fe2þ may play a role as initiating species for dOH radical formation following the Fenton reaction. Higher concentrations of Fe2þ could be expected if Fe(II) salts are used as precipitators in pretreatment steps of water and wastewater treatment. Primary and secondary alcohols are often called promoters for ozone-based systems due to the radical chain decomposition cycle by production of the chain carrier superoxide (HO2 d =O2 d ). Methanol (MeOH) is a typical model compound for investigating ozone-based AOPs. Phosphate can act as an efficient inhibitor in concentrations at 50 mM (Staehelin and Hoigne´, 1985). In Fenton
Advanced Oxidation Processes
reactions phosphate and iron(III), in addition, may precipitate forming Fe(III)–phosphate complexes and subsequently inhibiting the Fe(II)–Fe(III) cycle and reducing the dOH radical generation. 9. Some chemicals well known as inhibitors are acetate, tertbutanol (t-BuOH), and p-chlorobenzoic acid (pCBA). The latter one is also known as species with very low direct reaction with ozone and thus useful as an dOH probe for ozone-based reactions. Chloride (Jayson et al., 1973; Pignatello, 1992) and bromide (Zehavi and Rabani, 1972) may also serve as inhibitors at high concentrations due to the generation of weakly reactive radical products. Sulfate ions inhibit the Fenton reaction due to the formation of a mixture of FeSOþ 4 and Fe(SO4)2 complexes (Pignatello et al., 2006; Pignatello, 1992), which cannot be coordinated by hydrogen peroxide (De Laat and Lee, 2005).
389
electron transfer reactions with strong oxidants. This is one of the reasons for the differences in UV/H2O2 systems compared to UV/ozone systems. The path (h) in Figure 7 indicates that the organic radicals may initiate polymerization of unsaturated organic compounds in cases of lack of oxygen. It is worth noting that multi-chromatic or so-called broadband UV lamps are commonly used in H2O2/UV systems; therefore, direct photolysis by UV may also be a significant part of H2O2/UV reaction systems, especially in cases where the transmission of wastewater is relatively high. Ozone/UV and ozone/H2O2 reaction system. The reaction system of ozone-based systems is summarized by Figure 8 considering the H2O2/UV reaction system from Peyton (1990) and Legrini et al. (1993). The hydroxyl radicals can be generated by (a) photolysis of O3, (b) reaction of HO 2 with O3, or (c) photolysis of hydrogen peroxide, which is either a product of photolysis of ozone as indicated by path (d) as well as a product of ozone reaction with unsaturated organic compounds. The sequence of dOH radical reactions is comparable to the peroxyl radical formation of UV/H2O2 process. In these ozone-based processes, the peroxyl radicals may be considered as true propagators of the radical reactions due to the fast reaction of superoxide radical anion with ozone as indicated by path (e). Ozone reaction system. The ozone reaction mechanism in natural water and wastewater systems may be classified in two phases, the (1) initial phase, and the (2) second phase of ozone decrease (Buffle and von Gunten, 2006). During the initial phase, the ozone decomposition is very fast due to involvement of higher amount of free radical generation. The first initial phase can be subdivided further into a very fast and a fast phase No¨the et al. (2009b). The species amine and phenol (Ph), both present as moieties of NOM, accelerate the ozonide radical anion O3 d formation either by direct formation or by direct electron transfer, respectively. The ozonide
H2O2/UV reaction system. The oxidation of organic pollutants by H2O2/UV process may be summarized by reaction system of Peyton (1990) and Legrini et al. (1993) as shown in Figure 5. One started with reaction (17), indicated as reaction path (a) in Figure 5, the hydrogen abstraction by the dOH radical yielding the peroxyl radical is proposed and indicated as path (b) and (c). The peroxide radical may then undergo four possible pathways, (d) heterolysis and forming of organic cation and superoxide radical anion, or (e) 1,3 hydrogen shift and homolysis into hydroxyl radical and carbonyl compounds, or (f) back-reaction to oxygen and HRd, or (g) hydrogen abstraction starting a chain of oxidation reactions. The superoxide radical anion O2 d dominates in neutral solutions by pKa (HO2 d ) ¼ 4.8 (see, e.g., Czapski and Bielski, 1993; Behar et al., 1970) and a solvolysis reaction yielding hydrogen peroxide has been proposed by path (i). It should be noted that the superoxide radical anion is not very reactive (E1 ¼ 0.33 V; compilation of rate constants in Bielski et al. (1985)), except for
H2O O 2– • (i)
RH+
H2O2
RO (e)
(d) (a) RHO2•
O2
H2O HO•
(f)
HRH
h
(c)
RH• (b)
(g)
O2 (h) HRH
RHO2H
Polymer products
Figure 7 The reaction system of H2O2/UV process (Peyton, 1990; Legrini, et al., 1993). Redrawn with permission from Legrini O, Oliveros E, and Braun AM (1993) Photochemical processes for water treatment. Chemical Reviews 93: 671–698, Copyright 1993, American Chemical Society.
390
Advanced Oxidation Processes HO2• HO2– H2O
O2– • (i)
RH+
h
H2O2
RO (e)
H+
(d) (a) RHO2•
O2
(f)
HRH
(c)
RH
h
HO3•
HO•
•
H2O
(b)
O2
O2
(g)
O– •
O3–•
O3
(h) HRH RHO2H
Polymer products
OH –
Figure 8 The reaction system of ozone/UV and ozone/H2O2 process.
RH+
O2– •
HRH
RHO2•
RHO2H
HO2• O3 O2
OH –
RH •
A H2O A
•+
O
O2 O3– • O2
–•
HO •
H2O
H+
HO3•
HRH
O2
Figure 9 The reaction system of ozone in natural water and wastewater.
radical anion O3 d then undergoes reactions to generate dOH radicals (Buffle and von Gunten, 2006). Recent studies No¨the et al. (2009a) indicated that the radical formation does not stop after degradation of involved moieties. It is interesting to note that there is a continuous generation of dOH radicals during ozonation of wastewater effluents due to the formation of new electron-rich moieties by direct reaction with ozone No¨the et al. (2009a). The reaction system is summarized in Figure 9. It indicates the linkage with other AOPs by reaction of ozone with the superoxide radical anion O2 d =HO2 d formed by ozone/H2O2, ozone/UV, and ozone-activated carbon processes. The dOH
radical subsequently generated (reaction (21)) undergoes the same peroxyl radical reaction sequence as described before depending on scavenging capacity of water matrix and pollutants such as NOM and carbonate.
4.13.2.3.2 Reaction modeling Reaction modeling mainly comprises models of chemical reaction, mass transfer, fluid dynamic, and, in addition, fluence rate if photons are involved. Chemical models are restricted to the sequence of reactions considering reaction rate constants, reaction equilibria, etc., with the main objective to allow the determination and evaluation of reaction mechanisms and
Advanced Oxidation Processes
reaction competitions. Mass transfer and fluid dynamic models may be combined with chemical models, but the main objectives of these combined models are (1) the determination and optimization of design of reactor and process and/ or (2) the reduction of the experimental effort for upscaling of reaction system from laboratory to technical scale. Many different models of each of the above-mentioned topics can be found in literature, for example, free-radical reaction models, gas–liquid contact reactors, ozone reaction systems, heterogeneous catalytic systems, and fluence rate models. However, the examples briefly described hereinafter have been restricted to (1) chemical model examples illustrating both the complexity as well as the usefulness, for example, for oxidation of micro-/trace pollutants, and (2) photochemical reaction system to exemplify the relation of different factors on process efficiency which are also valid analogously for photo-catalytic processes. Chemical reaction model. The chemical models for AOPs are based on reaction rates with hydroxyl and other radicals, which are described by second-order rate law corresponding to Equation (55):
Rr ¼ kox;p Cox Cr
ð55Þ
with kox being the reaction rate constant, Cr the concentration of reactant r, and Cox the concentration of oxidant. The oxidant could be ozone or hydroxyl radical. Reported steady-state hydroxyl radical concentrations of H2O2/O3/UV systems are in the range of 1011–109 mol l1 (Glaze et al., 1987). Due to the fast reactions the concentration is expected to be low, but it depends on different factors, for example, concentration of reactants, photon flux, etc. Some reaction rate constants and dissociation equilibria are listed in Tables 3 and 4 for hydroxyl radical reactions as well as for ozone reactions. A large compilation of several hundreds reaction rate constants is listed in Buxton et al. (1988) or can be found most completely at Notre Dame Radiation Laboratory/National Institute of Standards and Technology (NDRL/NIST) Solution Kinetics Database (NIST, 2009). Rates for ozone-based systems for some emerging micro-pollutants can be found in Dodd (2008). Reactions with NOM expressed as dissolved organic carbon (DOC) (Reisz et al., 2003; No¨the et al. (2009a)) have been considered; therefore, the modeling of degradation of several trace organics in natural water as well as in urine was possible. Several approaches for kinetic modeling have been proposed in the literature. In ozonation processes, the ozone decomposition rate is one of the key parameters in plant operating. For natural waters, the ozone decomposition rate in the first initiating decomposition phase substantially depends on hydroxyl radical reactions and needs to be modeled adequately. For example, for better understanding of the reaction system, Buffle et al. (2006b) investigated conceptually the effect of reactive moieties distributions ci on ozone decomposition applying a model of coupled differential equations, which describe the temporal behavior of homogeneous, isothermal chemical reactions. This example illustrates that a more detailed simulation is very useful to understand the operationally observed instantaneous ozone demand based on parallel reactions of different moieties of NOM with
391
different oxidants. More than 1000 rate constants of reactions with ozone and hydroxyl radicals have been extracted from National Standard Database (NIST, 2002) and deployed by a program code called ACUCHEM (Braun et al., 1988), which can handle up to 80 simultaneous reactions and 40 species. Another possibility for kinetic modeling is the use of the chemical kinetics simulator free of charge by downloading from IBM Almaden Research Center website. This simulation tool has been used, for example, to evaluate the contribution of H2O2 formation during wastewater ozonation to yield dOH radical (Noethe et al., 2009). The aforementioned type of modeling refers mainly to mechanistic aspects of reactions, whereas the modeling of elimination rates is more useful from the operators and engineers point of view. Established parameters for possible use in ozonation reaction models are the chemical oxygen demand (COD) or the DOC (Beltra´n et al., 1995, 2001; Steensen, 1998; Wenzel, 1998). Since both the COD and DOC are summarized parameters comprising various compounds of different reaction rates, they therefore do not give suitable simulation results with one reaction rate constant. To overcome this problem, the DOC can be classified into several fictive groups of pollutants with different apparent reaction rate constants for each group of DOC. The advantage of this approach has been confirmed by No¨the et al. (2009a) for simulation of micro-pollutant elimination. The DOC of the wastewater was subdivided into three group components of DOC containing their corresponding concentration ci:
cðMatrixÞ ¼
n X
ci
ð56Þ
i¼1
The ozone decomposition rate then depends on these group components by Equation (57) and the group component elimination rates can be modeled by reaction (58) with the prerequisite that the ozone decomposition rate for each group component is known. Then the concentrations (ci) can be adapted iterative numerically by minimizing the error between simulation and experiment No¨the et al. (2009a):
n d½O3 X ki ½O3 ci ¼ dt i¼1
ð57Þ
dci ¼ ki ½O3 ci dt
ð58Þ
An analogous scheme in grouping the summary parameter TOC can be found, for example, in Zazo et al. (2009) for modeling the elimination rate of one pollutant by Fenton process. They simulated the pollutant degradation by a consecutive reaction scheme. The measured TOC in solution is always a sum of the different TOC fractions consisting of pollutants with concentration TOCA that reacts easily with hydroxyl radical forming a by-product fraction TOCB, that also react itself with dOH but at slower reaction rates forming final fraction TOCC. Fluence rate distribution models. In H2O2/UV and other photochemical reaction systems, the fluence, often called UVdose, and the UV absorbance of wastewater must be considered besides the chemical reactions. Not only AOP
392 Table 3
Advanced Oxidation Processes Reaction rate constants for ozone reactions and OH radical reactions
Reactions
Rate or equilibrium constants: M2 s1, M1 s1, s1, or dimensionless
Refs. and notes
B105 in hydr. urine B103 in ED diluate
Dodd (2008) Dodd (2008)
1.5 109 1.1 108 2.8 106 B5 106 6.0 105
Sehested et al. (1983) Sehested et al. (1984) Staehelin et al. (1982) Dodd (2008) Liu et al. (2001)
B3 108 8.5 106 3.9 108 9 107 1.1 108 2.7 107 2.7 107 7.5 109 1.2 1010 4.5 109 4.8 109 1.6 109 2.8 109 3.01 109 5.6 108 7.3 108 1.0 109 1.1 1010 2.0 106 1.0 106 4.3 109 5.0 105 7.6 109 1.5 108 1.8 108 2.0 108 6.0 107 1.5 1010 2.0 109 5.6 104 1.1 1010 1.1 1010 3.2 108 2.8 109 2.6 109 1.0 107 1.8 108 4.2 109 5.0 106 1.2 1010 1.25 1010 9.22 109 2.46 109
Westerhoff et al. (2007e) Buxton et al. (1986) Buxton et al. (1986) Neta et al. (1978) Sehested et al. (1984) Christensen et al. (1982) Kwon et al. (2009) Christensen et al. (1982) Buxton et al. (1988) Merenyi et al. (1999) Goldstein et al. (1998) Crittenden et al. (2005) Sutherland et al. (2007) Sutherland et al. (2007) Sutherland et al. (2007) Sutherland et al. (2007) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Crittenden et al. (2005) Kwon et al. (2009) Jones et al. (2009) Jones et al. (2009)
Ozone reactions O3 þ DOC-O3 d þ DOCdðþÞ O3 þ DOC-O2 d þ DOCðoxidizedÞ O3 þ DOC-O2(1Dg) þ DOC(oxidized) O3 þ DOC-DOC(peroxide) O3 þ O2 d -O3 d þ O2 d OH þ O3 -HO2 d þ O2 O3 þ HO2 -d OH þ O2 þ O2 d ONOO þ O3 -ðd NO2 þ O2 d þ O2 Þ? NO2 þ O3 -NO3 þ O2 Hydroxyl radical reactions d OH þ DOC-DOCd ( þ H2O, for Habstr.) d OH þ HCO3 -CO3 d þ H2 O d OH þ CO3 2 -CO3d þ OH d OH þ NH3-dNH2 þ H2O d OH þ O3 -HO2 d þ O2 d OH þ H2 O2 -O2 d þ H2 O þ Hþ d OH þ H2 O2 -H2 O þ HO2 d d OH þ HO2 -O2 d þ OH þ Hþ d OH þ OH-Od þ H2O d NO2 þ dOH-ONOOH ONOO þ dOH-dNO þ O2 þ OH MTBE þ dOH TAME þ dOH DIPE þ dOH TBF þ dOH TBA þ dOH Arsenic trioxide þ dOH Bromide ion þ dOH Carbon tetrachloride þ dOH Chlorate ion þ dOH Chloride ion þ dOH Chloroform þ dOH CN þ dOH Dibromochloropropane þ dOH 1,1-Dichloroethane þ dOH 1,2-Dichloroethane þ dOH HCN þ dOH Hydrogen sulfide þ dOH Hypobromous acid þ dOH Hypoiodous acid þ dOH Iodide ion þ dOH Iodine þ dOH Iron þ dOH p-Dioxane þ dOH Tetrachloroethylene þ dOH Tetrachloroethylene þ dOH Tribromomethane þ dOH Tribromoethylene þ dOH Tribromomethane þ dOH Vinyl chloride þ dOH PNDA þ dOH-OH-PNDA Domoic acid þ dOH Kainic acid þ dOH
DOC, dissolved organic carbon; MTBE, methyl tertiary butyl ether; TAME, tertiary amyl methyl ether; DIPE, di-isopropyl ether; TBF, tertiary butyl formate; TBA, tertiary butyl alcohol; HCN, hydrogen cyanide; PNDA, p-nitrosodimethylaniline.
Advanced Oxidation Processes Table 4
393
Reaction rate constants for carbonate radical reactions, superoxide radical reactions, and other radical reactions Rate or equilibrium constants: (M2 s1), (M1 s1), (s1), or dimensionless
Refs. and notes
B3 106
Canonica et al. (2005e)
B3 104 8 105 5.6 107 3.7 106
Dodd (2008) Behar et al. (1970) Behar et al. (1970) Goldstein et al. (1998)
1.5 109 B105
Sehested et al. (1983) Dodd (2008)
9.6 107 6.7 109
Christensen et al. (1988) Huie et al. (1993)
2.8 103 3.6 109 5.2 1010 3.7 104 1.1 105 B2.5 109 B100
Elliot et al. (1989) Buxton et al. (1986) Bu¨hler et al. (1984) Bu¨hler et al. (1984) Bu¨hler et al. (1984) Dodd (2008) Dodd (2008)
1.7 106 3.5 109 3.2 109 B7 105 B3 105
Buxton et al. (1970) Merenyi et al. (1999) Goldstein et al. (2005) Lazlo et al. (1998h) Dodd (2008)
B3 109 B3 109 4.5 109 1.8 109
Dodd (2008) Dodd (2008) Løgager et al. (1993) Løgager et al. (1993)
Other reactions (nitrogen-based) ONOO - d NO þ O2 d ONOO þ CO2 -0:33d NO2 þ 0:33CO3 d
2.0 102 2.9 104
ONOO -NO3 ONOO-dNO2 þ Od ONOOH-dNO2 þ dOH ONOOH-NO3 þ Hþ ONOOH þ H þ -NO3 þ 2Hþ ONOOH þ H2O þ Hþ-HNO2 þ H2O2 þ Hþ HNO2 þ H2O2 þ Hþ-ONOOH þ H2O þ Hþ O2 NOO -d NO2 þ O2 d O2 NOO -NO2 þ O2 O2 NOOH-d NO2 þ HO2 d
B8 106 B106 3.5 101 9.0 101 4.3 1.1 101 9.6 103 1.0 1.4 2.6 102
Goldstein et al. (2005) Goldstein et al. (2005), Lymar et al. (1995) Goldstein et al. (2005) Goldstein et al. (2005) Goldstein et al. (2005) Goldstein et al. (2005) Merenyi et al. (2003) Merenyi et al. (2003) Merenyi et al. (2003) Goldstein et al. (1998) Goldstein et al. (1998) Goldstein et al. (2005)
Reactions
Carbonate radical reactions CO3 d þ DOC-DOCdðþÞ þ CO3 2 CO3 d þ NH3 -d NH2 þ HCO3 CO3 d þ H2 O2 -HO2 d þ HCO3 CO3 d þ HO2 -O2 d þ HCO3 ONOO þ CO3 d -d NO þ O2 þ CO3 2 Superoxide radical reactions O3 þ O2 d -O3 d þ O2 O2 þ DOCðoxidizedÞ -O2 þ DOC O2 d þ HO2 d -HO2 þ O2 d NO þ O2 d -ONOO d
dðÞ
Other radical reactions O3 d -Od þ O2 O2 þ Od -O3 d O3 d þ Hþ -HO3 d HO3 d -O3 d þ Hþ HO3 d -d OH þ O2 DOCd þ O2-DOCOOd DOCOO d -DOCðoxidizedÞ þ HO2 d DOCOO d þ OH -DOCðoxidizedÞ þ O2 d Od þ H2O-dOH þ OH d NO2 þ Od-ONOO d NO þ HO2 d -ONOOH NH2 O2 d -d NO þ H2 O NO2 þ DOC-DOCdðþÞ þ NO2 NO2 þ DOCd-DOCNO2 d NO þ DOCd-DOCNO d NO2 þ O2 d -O2 NOO d NO2 þ HO2 d -O2 NOOH d
d
reactions but also direct photolysis of organic compounds are often involved in theses processes. The fluence or UV-dose is the area-specific total radiant energy of all wavelengths passing from all directions through a cross sectional area in terms of, for example, J m2. UV light lamps have different range of UV emission as indicated by Figure 4. Modeling of fluence rate distribution in UV reactor allows simulation of photolysis-induced AOP kinetics and thus the degradation of photo-reactant component. The fluence rate models must be combined with computational fluid dynamics (CFD) as well as with the aforementioned chemical models and/or models of direct pollutant photolysis to calculate the
overall oxidation rate and photoreactor performance. An example can be found in Sozzi and Taghipour (2007). For fluence rate distribution modeling, the so-called multiple point source summation (MPSS) model (Blatchley, 1997) may be used. In this model the UV lamp is approximated as a series of point source emitting diffuse radiation to calculate the UV irradiance distribution. The fluence rate E is estimated at each point by
Eðr; zÞ ¼
n X i¼1
p li exp s ð r r Þ w l n4pl2i r
ð59Þ
394
Advanced Oxidation Processes
with z and r being the axial and radial distances, P the total lamp output, sw the absorption coefficient of the medium, rl the lamp sleeve radius, and li the distance from a specific point (r, z) to ni point source out of the total of n sources. The rate E(r, z) as a term of the local radiant power E is then useful to calculate the local degradation rate as a function of local energy absorption, quantum yields, oxidant and pollutant concentration, as well as reaction rate constants. The fluence rate model can be improved by considering refraction and reflection at the UV-lamp–water–sleeve interface (air–quartz–water) and the geometry of the lamp by a series of cylindrical elements instead of points (Bolton, 2000). Experimental validation is absolutely essential to consider the effect of interference of UV-absorbance between target pollutant and water matrix as well as its dependence on depletion or degradation. More accurate simulations include the validation of fluence of UV lamp by the use of an actinometer. An actinometer is defined as a chemical/physical system that determines the number of photons in a beam integrally or per unit time. A possibly useful tool for experimental validation may be the procedure of defining shape factors and classifying the wastewater absorbance into several groups of different UV absorbance and TOC degradability (Muret et al., 2000). More difficult than aforementioned modeling of artificial UV reaction systems is the simulation of solar-radiation-based processes since the solar radiation may change every day and even every minute depending on weather and the angle between sunlight source and reactor position. The reader is referred to Farias et al. (2009) and Imoberdorf et al. (2007), in which the authors describe the modeling and absorbance of polychromatic sunlight and the elimination rates of pollutants during different sunlight fluent rates as well as the quantum efficiencies in multi-annular photo-catalytic reactor. It is important to note that they developed a model of reduced parameter to enable less complex simulation supporting the applicability of solar-based reaction systems. It seems to be necessary to operate an actinometer in parallel for evaluation of actual fluence rates enabling suitable plant operation to reach target treatment effluents.
4.13.3 Guidance for Selecting an AOP As previously mentioned, each AOP for a particular application needs feasibility studies in laboratory and pilot scale before application due to the necessity of strategies to minimize by-product formation and/or careful control of oxidant dose. In principle, a preselection of one or more specific AOPs based on exclusion criteria is most useful. For example, costs of alternative treatment processes such as (1) membrane processes including concentrate treatment, (2) activated carbon adsorption including regeneration of adsorbent, (3) evaporation, or (4) combustion, etc., can be used for the estimation of maximum limits for operating costs which can be used for calculating the maximum oxidant dosage, energy demand, chemicals for pH adjustment, and disposal of waste sludge (Fenton reactions) before starting laboratory investigations.
4.13.3.1 Criteria to be Considered Before selecting a specific AOP or a combination of AOPs for a particular application, several factors should be considered such as (a) water quality, (b) the yield of hydroxyl radicals with a given AOP, (c) the amount of radical scavengers, (d) the required energy input of the AOP system, (e) the oxidant residual (if used) and its impact on downstream processes, (f) plant design, and (g) investment and operational costs. The criteria (a) to (e) are preferably evaluated by lab-scale studies whereas criteria (f) and (g) are generally referred to pilot-scale investigations for validation. Suitable consideration of these factors is only possible by adequate knowledge about analytical and process equipment as well as some key data on costs for equipment. Water quality. The tasks associated with AOPs are (1) reduction of toxicity or endocrine disrupting effects by transformation or mineralization of target compounds, (2) partial oxidation of biologically refractory dissolved organic compounds to improve the subsequent biodegradation, (3) color removal, (4) odor removal, and (5) disinfection. Many water-quality parameters such as summarized parameters (COD, DOC, suspended solids, heavy metals, alkalinity, color, colony forming units (CFUs), etc.) are regulated and thus the analytical methods for these parameters are well known and not subject of this chapter. The yield of hydroxyl radical of a given AOP. The yield of dOH radicals depends on wastewater characteristics, process configuration, and type of AOP. The yield is mostly determined indirectly by the evaluation of pollutant degradation rates for different process configurations, because the direct detection of hydroxyl radical is very difficult due to its fast reactive character and the resulting extremely low dOH concentration. A more detailed evaluation of dOH radical contributions to an oxidation process is possible with the use of a suitable dOH probe, which allows the identification of a characteristic product by taking into account the competition kinetics. The use of dOH probes allows the calculation of dOH concentration. For example, pCBA has been identified as a suitable dOH probe for ozonation processes (Elovitz and von Gunten, 1999; No¨the et al. (20009a), elucidating that d OH yield of ozonation is at least 13% in the third phase No¨the et al. (2009a) and much higher in the first initial phase. Using such dOH probes, the comparison of dOH radical yield of different AOPs is possible for each particular application. The presence, type, and quantity of radical scavengers. The competition reactions are strongly affected by scavengers as described earlier. AOPs rely on a suitable ratio of number of d OH radicals generated to the number of radicals consumed in reactions to the number of radicals decomposed by termination reactions. This ratio may lead to several suitable AOP applications even in the presence of inhibiting scavengers due to increased efficiency of target reaction cycles depending on water and wastewater characteristics and the relation of inhibiting to promoting effects of compounds and moieties. The required energy input of the AOP system. The most important energy-consuming factor of different AOPs is the power required, for example, for ozone generation and ozone gas-to-liquid mass transfer at ozone applications or artificial
Advanced Oxidation Processes
UV irradiation at UV-based applications. Other examples are the power needed for cavitation systems, e-beam processes, or electro-chemical reactions. The energy for pumping, mixing, and solid/liquid separation is approximately at the same level compared to that for non-AOP systems and therefore out of the focus hereinafter. One of the keys for the improved applicability of ozone in the last few decades was the enhanced energy efficiency in ozone generation. Several factors such as temperature, materials, purity of liquid oxygen supplied, as well as construction details of ozone generator influence the energy demand for ozone generation which has been reduced to date to approximately 6–10 kWh kg1 ozone generated (Ried et al., 2009). Also, UV systems have been optimized in the last few decades relating hydrodynamics, fluence rate, oxidant dosage, etc. The systems are now much more compact, a result mainly based on improved fluid dynamics and absorbance efficiency supported by CFD simulation combined with fluence rate models. In addition, the lifetime and fluence rate of UV lamps have been increased significantly. Moreover, the fouling/scaling problem has been reduced by the application of cleaning devices. Since the fouling/scaling problem is mainly addressed to the temperature increase at the quartz sleeve surface to water, new lamp developments, for example, in plasma-induced UV lamps, may overcome these problems. Plant design. The design of AOP system must address the minimization of by-product, for example, by careful control of oxidant dose and residual oxidant in AOP effluents considering its impact on downstream processes (e.g., biological systems). The bromate issue is one of the well-known byproduct examples, which needs a careful control of ozone dosage as a prerequisite besides others. The plant design concerns not only the AOP itself but also the process integration and combination with non-AOP systems. As discussed later hereinafter, AOPs mostly produce biodegradable compounds from biologically refractory organics. If AOPs are selected as final treatment step, complete mineralization and negligence of environmental impact through by-product generation should be ensured. However, this may cause increased treatment costs and failure of costefficient application. An often better strategy, and mostly not to be missed for drinking water production, is to integrate AOPs as an essential part of multi-stage water and wastewater-treatment concept
considering the benefits of AOPs compared to other processes. This may cause a reduction of the required oxidant to a level as low as possible for supporting cost-efficient applications. Investment and operational costs. In general, investment and operational costs are very specific for each application depending on various site-specific factors. All the points must be considered for the optimization of the investment and operational costs. Additional very specific cost factors of interest are, for example, (1) the costs for chemicals such as ferric/ ferrous solutions or hydrogen peroxide, (2) the lifetime of UV irradiation lamps, ozone generator, reaction system considering, for example, scaling, fouling, corrosion problems, and catalysts, if used, concerning inactivation, leaching, etc., (3) the available space for treatment system, (4) local energy costs, (5) costs and local presence of system service, (6) operational flexibility, etc. Despite the fact that costs will change every time depending on market structure, some costs should be mentioned hereinafter, because it may be useful for possible evaluation of new applications either by indicating the necessity of cost reductions and process optimization for new applications or by giving a benchmark enabling the development of new AOPs (e.g., solar-based photo-catalytic processes).
4.13.3.2 Cost-Related Factors of Ozone-Based Processes One of the main specific cost factors is the concentration of pollutant in the water to be oxidized and the ratio of mass of oxidant to mass of pollutant both contributing to the specific ozone dosage per volume of wastewater. Figure 10 summarizes different regions of ozone dosages for different applications. The ozone costs are mainly influenced by the energy demand of ozone generator and the costs for oxygen supply for ozone generation costs, which also depend on the size of ozone generator. Taking the previously mentioned 10 kWh kg1 and assuming 0.2 Euro/kWh as a basis for cost calculation, the specific energy costs per cubic meter wastewater allocated to ozone generation will range from 1 to 40 Eurocent m3 wastewater. Additional costs to be considered are the oxygen supply and the energy costs for gas–liquid mass transfer of ozone into water, which depend mainly on the size and configuration of the ystem. A typical oxidant-to-pollutant ratio for COD elimination is 2–4 kgO3 kg1COD. Optimization of the specific ozone dosage, or generally speaking the oxidant dosage, is crucial for the application of
AOX-/COD-elimination Odor control Decoloration Micropollutant control Disinfection 0 1
5
10
25
395
50
100
200
Ozone dosage (gO3 m–3wastewater) Figure 10 Range of ozone dosage and related treatment effects. Courtesy of ITT Wastewater GmbH Wedeco redrawn with minor changes.
396
Advanced Oxidation Processes
cost-effective AOPs. Several strategies have been established to reduce the ozone dosage: (1) partial oxidation to produce biodegradable compounds from recalcitrant organics, (2) pretreatment by removal of radical scavenging compounds, for example, precipitation and particle removal, (3) pretreatment by more cost-effective concentration of recalcitrant pollutants and subsequent treatment of concentrates, for example, the retentate of membranes processes, or the regeneration of adsorbents by AOPs, (4) improving the yield of OH radical generation by optimizing the process combination of oxidants and UV irradiation as well as their sequence of use, and (5) optimization of number and location of oxidant dosing points considering reaction kinetics and by-product control.
4.13.3.3 Cost-Related Factors of UV-Based Processes The cost efficiency of UV-based processes depends mainly on the energy demand of UV system and the hydrogen peroxide dosage. While the hydrogen peroxide dosage still has the analogous cost-related dependencies as mentioned in previous ozone paragraph and also needs to be minimized by optimization tasks, the energy efficiency of UV irradiation mainly depends on the UV light absorption characteristics of the wastewater. If the water matrix absorbs the main fraction of UV irradiance, the efficiency is low and the process may become uneconomic. Other AOPs are then expected to be more cost efficient. To allow comparison of different AOPs, the figure-of-merit concept introduced by Bolton and Bircher (1996) should be useful. A figure-of-merit considers the electrical energy demand per order of magnitude of oxidative degradation EE/O in relation to the volume of wastewater or to the mass of pollutant and can be calculated either from continuously operated or from batch-operated systems by Equations (60) and (61), respectively:
EE=O ¼
V0l
EE=O ¼
Pel ðcontinuous systemsÞ logðc0 =cÞ
ð60Þ
Pel t 1000 ðbatch systemsÞ VR logðc0 =cÞ
ð61Þ
This approach is useful since the depletion of summarized pollutant parameters, for example, characterized by TOC, does not follow the second-order rate law due to the change of nature of TOC during oxidation. It may also be useful for scale-up tasks. While the electrical energy demand for ozone generation and subsequent mass transfer as well as for UV systems could be directly inserted in Equations (60) and (61), the oxidant hydrogen peroxide needs to be considered indirectly by converting the costs of hydrogen peroxide to the energy equivalents at site-specific conditions. It should be noted that the EE/O values may also be calculated for both electrochemical and solar-irradiated systems (Bolton et al., 2001). Typical EE/O values range from 0.1 to 1–5 kWh m3 whereas the lower range is assigned to disinfection issues and the higher values to COD-degradation with AOP H2O2/UV. Table 5 exemplifies some typical EE/O values at different H2O2 dosages, indicating that (1) EE/O is reduced at higher oxidant dosages and (2) efficiency may decrease slightly for
Table 5 Some typical EE/O values for atrazine degradation in lake water by UV/H2O2 process (Kruithoff et al., 2007) H2O2 dosage (g m3)
EE/O pilot scale (bench scale) (kWh m3)
4 8 15 25
1.50 1.30 1.00 0.70
(1.30) (1.15) (0.90) (0.65)
scale-up from bench to pilot scale. Due to the relative low dosage of oxidant in this case study, the total EE/O decreases but at higher dosage values typical for high-loaded industrial wastewater, the EE/O may not decrease due to possible stronger energy-equivalent contribution of H2O2, which is based on the costs of H2O2 and the energy equivalent of the costs. It is also worth noting that the EE/O may increase due to decrease of degradation efficiency after several months of UV lamp operation caused by fouling and/or scaling. An example of this decreasing efficiency was taken from Kruithof et al. (2007) for pesticide degradation in IJssel lake raw water located in The Netherlands, which was implemented as pretreatment step to potable water production (Figure 11). This decrease should be avoided or compensated by automated cleaning and adaptation of power input to UV lamp, respectively. A comparison of three different AOPs (1) H2O2/UV, (2) ozone/UV, and (3) H2O2/O3 by EE/O approach has been provided by Mu¨ller and Jekel (2001) for atrazine degradation. To enable such a comparison, the production costs of oxidant hydrogen peroxide have been converted to energy equivalents based on local energy costs. The calculated EE/O values range from 0.90 to 2.92 kWh m3 for the ozone/UV process, from 1.67 to 2.27 kWh m3 for the UV/H2O2 process, and from 0.102 to 0.135 kWh m3 for the O3/H2O2 process. This indicates that the O3/H2O2 process has been identified as the best option in this case by far. However, it is important to note that photochemical-based AOP often includes an optimization potential with respect to, for example, the strong impact of hydrodynamics or the type and fluence rate of UV lamp. This optimization procedure has not been documented and therefore the comparison just relies only on the used system layout. A strong indication of the existing optimization potential is the EE/O values provided by Mu¨ller and Jekel (2001) for the UV/ozone process, which have been found lower for a two-step process compared to the higher values of the onestep process.
4.13.4 Description of Processes 4.13.4.1 Ozonation A typical ozonation scheme is given in Figure 12. The process consists of (1) an oxygen gas supply, (2) an ozone generator, (3) a gas–liquid contact tank, which, in principle, is the ozone reactor, (4) a degassing tank, with decomposing residual ozone in gas phase, (5) a cooling unit for ozone generator, (6)
Advanced Oxidation Processes
397
100 After start-up After 2000 h Degraddation rate (%)
80
60
40
20
0 Atrazine
Pyrazone
Diuron
Bentazone
Bromacil
Figure 11 Pesticide degradation in pretreated Ijssel Lake Water at the beginning and after 2000 h of plant operation.
5–32 °C
Electrodes
Cooling unit
Fan
Ozone
Ozone destructor
Liquid oxygen tank
400 V FI
FI
Demister (option) Ozone generator Controls + power supply Treated water
Wastewater
Diffusors
Diffusors
Ozone reaction tank
Figure 12 Ozone system layout. Courtesy of ITT Wastewater GmbH Herford (formerly Wedeco) with minor modification.
a power supply, and (7) a control unit. Different online-sensor devices could monitor the ozone feed gas flow and its ozone concentration, the ozone concentration in off-gas and the dissolved concentration O3 after reaction chamber. These allow control of variable ozone dosages depending on water flow rate. The water or wastewater flows through the reaction chamber, where the ozone feed gas can be supplied by different gas–liquid contact systems. Various contact systems can be used, depending on water flow, reaction kinetics, and required ozone dosage: (1) gas diffuser (e.g., a ceramic disk), (2) two-phase side stream, or (3) main stream injector. The reaction chamber may be constructed as a pressurized reaction
tank and/or as a cascaded reaction system comprising baffles for avoiding short-cut flow. Important parameters to be measured or calculated for optimizing purposes are (1) ozone dose as difference of gaseous input and output mass flow, (2) the aqueous ozone concentration if not zero, (3) the ozone exposure, (4) the hydroxyl radical exposure, (5) the pollutant-specific ozone dose, and (6) for batch process, time-dependent oxidation rate of pollutant; for continuous process, pollutant concentration at input and output at different retention times considering hydrodynamic flow regime. If interference effects are expected, for example, by DOC, carbonate, etc., additional analytical measurements are necessary.
398
Advanced Oxidation Processes
In combined ozone/UV processes, the ozone is transferred into the wastewater either before entering UV-system or simultaneously to oxidation at different points of reaction chamber.
The water to be treated enters the UV photo-reactor after injection of suitable hydrogen peroxide dose, which is often well distributed by static mixers. Depending on the reaction kinetics and possible side reactions, additional H2O2 injection points at different reactors within a series of reactors have been considered. In few applications, the pH value has been lowered by an acid to shift the carbonate–bicarbonate equilibrium, and thus reducing the dOH radical scavenging capacity. After oxidative treatment neutralization may be necessary to meet discharge requirements. Photochemical processes rely significantly on (1) emission spectrum and intensity of UV lamp, (2) the absorbance of pollutants, (3) possible UV interferences, and (4) fouling and scaling characteristics of wastewater. Several types of established UV lamps are listed in Table 6 with respect to their typical operating parameters. The low pressure–low intensity (LP–LI) and low pressure–high intensity (LP–HI) lamps are usually applied for disinfection and combined UV/ozone processes, respectively, while medium pressure–high intensity (MP–HI) lamps are most useful for H2O2/UV applications. However, this
4.13.4.2 Photo-Chemical Oxidation A typical photo-oxidation treatment system’s layout based on H2O2/UV process is illustrated in Figure 13. Commonly, a UV-oxidation system comprises (1) a photo-reactor containing one or more UV lamps, (2) a hydrogen peroxide storage tank, (3) a UV lamp control panel, (4) a cleaning unit for the UV-transmissive quartz tubes, (5) an inert gas cooling system for UV lamps (only medium pressure lamps), (6) a power supply, and (7) a control unit with UV-intensity sensor as most important sensor device to control the fluence rate by automated cleaning and/or adapting the UV lamp power. The hydrodynamic characteristics inside the photoreactor is often controlled, for example, by baffles or special guiding plates.
Treated wastewater
Acid
Base
UV lamp
Hydrogen peroxide
Baffle Reactor
Wastewater Static mixer
Static mixer Oxidation unit
Figure 13 UV-oxidation system layout.
Table 6
Characteristics of different UV lamp types
Characteristic
Emission type Peak output wavelength Germicidal output to input Energy output at 254 nm Power consumption Lamp current Lamp voltage Operating temperature Mercury vapor pressure Lamp length Lamp diameter Sleeve life Ballast life Estimated lamp life
Unit
– nm % W W mA V 1C mmHg m mm Year Year h
Type of lamp MP–HI
LP–HI
LP–HI
Polychromatic 200–400 10–15 – 1 000–10 000 Variable Variable 500–900 102–104 0.05–1.95 Variable 1–3 1–3 3000–8000
Monochromatic
Monochromatic 253,7 30–40 25–27 40–100 350–550 220 35–50 0.007 0.75–1.5 15–20 4–6 410 8000–12 000
LP–LI, low pressure–low intensity; LP–HI, low pressure–high intensity; MP–HI, medium pressure–high intensity.
25–35 60–400 200–1200 Variable Variable 60–100 0.01–0.8 Variable Variable 4–6 410 7000–10 000
Advanced Oxidation Processes
position is part of a controversial discussion due to different operational costs for cleaning, cooling, lamp life, etc. Known UV interferences affecting the efficiency of H2O2/ UV process are nitrate410 ppm, nitrite410 ppm, phosphate41%, chloride41%, COD41000 ppm, and ferrous ion (Fe3þ)450 ppm. Scaling may occur at concentrations of Caþ 450 ppm, Fe3þ450 ppm, and Mg2þ41000 ppm.
A final filtration unit to remove fine aggregates may be necessary to meet discharge requirements. To enhance the Fenton reaction, the reaction system may be heated and temperature controlled at approximately 50 1C to take advantage of faster reactions and reduced oxidant dosage (see Pignatello et al., 2006). The Fenton process can easily be upgraded to a photo-Fenton process by the implementation of a second cycle loop with UV irradiation (see photo-Fenton box in Figure 14). This allows reduction of all, the catalyst dose, the amount of precipitates, as well as the sludge to be disposed. On the other hand, optimization due to fouling/scaling risks as well as energy minimization due to additional power that is required are necessary. This has to be proved before application.
4.13.4.3 Fenton and Photo-Fenton Processes The general Fenton process layout consists of (1) storage tanks for acid, base, hydrogen peroxide, and ferrous or ferric salt solutions, (2) first pH control unit to dose acid and control acidic conditions, (3) injection system for ferric or ferrous salt solutions, (4) injection system for hydrogen peroxide, (5) stirred tank reaction system comprising up to three reaction tanks in cascade, (6) second pH control unit to dose a base and control neutral conditions, (7) flocculation system with flocculant dosage, (8) sedimentation tank, (9) water filtration unit, (10) dewatering device for sediment which has been separated from the sedimentation tank, (11) a power supply, and (12) a control unit. Figure 14 illustrates a typical flow scheme of Fenton process. The acid, catalyst(s), and the oxidant hydrogen peroxide are mostly injected into a cycle loop of reaction system. The pH control is easy to operate, whereas the optimal catalyst and hydrogen peroxide dosage as well as the reaction time should be previously determined and validated. Optimal pH is known at approximately pH 3; however, higher values (e.g., up to pH 4) are sometimes applied. After processing Fenton reactions the wastewater is neutralized with caustic, lime slurry or another base, and thus ferric ion complexes precipitate. For improving the settling efficiency of precipitates and subsequent dewatering efficiency, flocculants are often added.
4.13.4.4 Process Combinations AOPs will take their advantage mostly in combination with other processes. One of the best-known examples is the combination of AOPs and biological (BIO) processes. Two approaches have been established. (1) Minimization of costs and optimization of BODproduced-to-CODremoved ratio by consciously reducing the oxidant dosage to enable partly oxidation. This approach leads to a transfer of COD removal from AOP to biological process. Different system layouts have been realized for ozone-based AOPs from sequential process schemes AOP–BIO, BIO–AOP, and BIO–AOP–BIO to continuous cycle loop process between BIO and AOP depending on the concentration level of pollutant. The latter process is interesting from engineering point of view due to higher potential for process optimization and by-product control by varying the cycle flow rate in addition to oxidant dosage. (2) The second approach is a combination of AOPs and
Optional installation for photo-Fenton Catalytic oxidation UV reactors
Post filtration, e.g., fixed bed Effluent
Fe 3+ (catalyst) acid H2O2 (oxidant)
Raw wastewater M
Additives
Chemicals
• Coagulants • Neutralization agents
• Flocculation agents
M
M
Mixing and buffer tank
M
OI
Precipitation and neutralization
Figure 14 General system layout for Fenton reaction treatment.
399
Sludge treatment
Filter press
400
Advanced Oxidation Processes
artificial groundwater recharge for removal of organic micropollutants; the improved removal of both transformed micropollutants as well as NOM offers an advantage for by-product control in drinking water production processes involving chlorine as disinfectant. Another option for a combined process is to concentrate the toxic or recalcitrant organics before applying AOPs. Different concentrating processes exist, for example, membrane processes and adsorption processes. The combined processes can be useful when the alternative processes are easily and cost effectively operated and the concentrates are classified as hazardous waste; or in case of absorption, if a new process of heterogeneous catalytic oxidation may operate well.
4.13.5 Full-Scale Applications Applications described hereinafter cover AOPs for drinking water supply as well as wastewater and waste treatment due to the reason that AOPs are involved in all fields, and the boundaries between drinking water and wastewater treatment are fading. The latter one has gained significant attention by new findings in early 1990s about linkage between the occurrence of endocrine disrupting chemicals (EDCs) in wastewater receiving water bodies and the reproductive impacts on aquatic life. Ever since, many studies have been started with focus on occurrence of trace contaminants in effluents of wastewater treatment plants (WWTPs), surface water (rivers, lakes, etc.), groundwater, and drinking water. New analytical methods for detecting trace contaminants in sub-mg l1 concentration range have led to an increased number of so-called emerging contaminants, micro-pollutants, and EDCs and currently to the status of a very dynamic field in wastewater
Diffusive sources (agriculture atmospheric deposition, transport systems, sewer leakage, etc.)
and water treatment. Nowadays, it has become clear that WWTP effluents form a significant emission source for pharmaceuticals, EDCs, personal care and health products, and other nonbiodegradable hazardous contaminants for regional surface water and raw water. Figure 15 illustrates schematically the link between emission sources and drinking water supply. Emission sources can be classified as diffusive- and point-source emissions. The diffusive-source emissions cannot be controlled by wastewatertreatment options and often cause direct contamination of groundwater. However, controlling the diffusive contamination has been realized in the European Union (EU) for some toxic compounds, for example, by removing the chemicals from the market (some priority substances of EU, see hereinafter), restriction of imports, etc. The point-source emissions are connected to wastewater streams and therefore regularly received by surface water bodies. This emission situation generally allows several treatment options at different locations 1–5 in Figure 15. AOPs may play an important role especially as an essential part of a multi-barrier water-treatment concept for the degradation of toxic, mutagenic, carcinogenic as well as bio- and ozone-refractory organic pollutants. Some of these pollutants are classified as drinking water-relevant compounds indicating that these compounds are passing through natural or closely natural drinking water plants. Tables 7 and 8 summarize the degradability by AOPs of hazardous as well as trace organics in relation to the source character and the possibility for oxidation and detoxification.
4.13.5.1 Ozone-Based AOPs Full-scale ozone-based AOPs for treating process water, wastewater, cooling water, and landfill leachate are implemented
Point sources Domestic WWTP 1
Industrial WWTP 2
Hospitals 3
Surface water body
Water work
Groundwater
4
5
Point of use
5
Point of use
5
Point of use
Drinking water Figure 15 Treatment options for AOPs in context to different pollution sources and linkage between wastewater and drinking water. WWTPs, wastewater treatment plants.
Advanced Oxidation Processes
401
Table 7 List of priority substances of EU framework directive, the main pollution sources, and the degradability of the organic compounds by advanced oxidation processes (AOPs) including ozonation Compound
Class
Main pollution source
Degradability by AOP (only organics)
Alachlor Atrazine Benzene
Herbicide Herbicide Industrial chemical
Agriculture Agriculture Industrial WW, atmospheric
99% 80–99% Almost completely
Lead
Heavy metal
(Inorganic)
C10-13-Chlorinated paraffins Cadmium
Flame retardant Heavy metal
Chlorfenvinphos Trichloromethane Chlorpyrifos Diethylhexyl phthalate (DEHP) 1,2-Dichloroethane Dichloromethane
Insecticide Industrial chemical Insecticide Plastics softener Industrial chemical Halogenated hydrocarbon
Rainwater, sewage, industrial WW Not known Rainwater, agriculture, atmospheric Agiculture Industrial WW Agriculture, atmospheric Atmospheric, rainwater, sewage Industrial WW Industrial WW, atmospheric
Diuron Endosulfan Hexachlorobenzene Hexachlorobutadiene
Herbicide Insecticide Industrial chemical Industrial chemical
Agriculture Agriculture Industrial WW No production and use in EU
Almost completely Up to 97% 20%
Isoproturon Lindane (g-hexachlorocyclo-hexane) Nickel
Herbicide Insecticide Heavy metal
Almost completely Almost completely (Inorganic)
Nonylphenol
Industrial chemical
Agriculture Agriculture (sewage, rainwater) Rainwater, erosion, mining, atmospheric Sewage
Octylphenol Polyaromatic carbons Polybromated diphenylethers
Industrial chemical Industrial chemical Industrial chemical
Automobile traffic, rainwater Atmospheric Diffuse (sewage)
Almost completely Up to almost completely
Pentachlorobenzene Pentachlorophenol Mercury Simazine
Industrial chemical Industrial chemical Heavy metal Herbicide
Sediments (old contaminations) Industry Erosion, drainage, sewage Agriculture
495% (Inorganic) Yes
Hexabutyl distannoxane (organic tin) Trichlorobenzene Trifluralin
Industrial chemical Industrial chemical Herbicide
Water traffic Atmospheric Agriculture
Yes 70–97%
worldwide in chemical, pharmaceutical, textile, pulp and paper, semiconductor, and other industries. In this field, more than 1000 ozonation plants have been installed from 1954 to 1999 according to Bo¨hme (1999), most of them for process water conditioning (660). Within this context, the combined ozone-biology leachate treatment is a typical German exception relying on early strong requirements since 1989 to meet effluent limits below 200 mgCOD l1. However, this is a good example for being a more cost-efficient combined AOPprocess in the 1990s compared to other technologies such as stand-alone AOPs, absorption, or membrane processes. Recently, this approach receives an increasing interest for industrial wastewater treatment as well as for sewage posttreatment. Table 9 summarizes different wastewatertreatment applications based on ozone-biology process combinations as far as known to the author. Sewage may contain high fractions of refractory compounds, for example, originating from industrial sites. In the case of a Danish sewage treatment plant, the main fraction originates from pharmaceutical industry and due to planned
Almost completely (Inorganic) Almost completely Almost completely Almost completely Almost completely Almost completely
Almost completely
extension of its production capacity an adequate upgrading of sewage treatment plant was decided (Ried et al., 2009). The industrial wastewater is characterized as almost biorefractory containing, for example, the pharmaceuticals furosemid, sulfamethizol, and ibuprofen in much higher concentrations compared to the commonly known sub-mg regions of micro-pollutants. Due to the high fraction of biorefractory organics, the COD of the treatment plant effluent needs to be reduced from 120 to below 75 mg l1. The system was designed for an ozone dosage of 180 kg h1 ozone per 1200 m3 h1 wastewater and 15 min of reaction time, and enables the COD reduction of B40%. This was confirmed by effluent concentrations of approximately 65–75 mgCOD l1. It is worth noting that a complete removal of pharmaceuticals was detected at a much lower ozone dosage of 10–30% compared to COD removal design dosage. However, for safe operation issues, the ozonation system has been extended with a hydrogen peroxide unit, thus enabling perozone process. From the operating conditions, the costs were estimated to approximately 0.2 Euro m3
402
Advanced Oxidation Processes
Table 8
List of selected pharmaceuticals, the removal efficiency of sewage plants, and the degradability by AOPs including ozonation
Chemical group
Pharmaceutical compound
Removal by sewage treatment plant (%)
Removal by AOPs (posttreatment) (%)
Antibiotic
Ciprofloxacin Clarithromycin Erythromycin Sulfamethoxazol Trimethoprim
83 66 45 51 20
91 84 84 90 82
Lipid regulator
Benzafibrate Clofibrin acid Fenofibrin acid
68 (n ¼ 14) 17 (n ¼ 7) 46 (n ¼ 4)
83 (n ¼ 2) 71 (n ¼ 3) 63 (n ¼ 1)
Beta-blocker
Atenolol Metoprolol Sotalol
9 (n ¼ 1) 57 (n ¼ 4) No data
65 (n ¼ 2) 87 (n ¼ 2) 97 (n ¼ 2)
Analgesic anti-inflammatory
Acetylsalicylic acid Diclofenac Ibuprofen Paracetamol Phenazone
88 42 83 92 33
Not relevant 92 (n ¼ 9) 92 (n ¼ 2) 91 (n ¼ 1) no ref.
Antiphlogistic
Indometacin Propyphenazone
66 (n ¼ 3) No data
71 (n ¼ 2)
Anti-epileptic
Carbamazepine
12 (n ¼ 14)
96 (n ¼ 3)
Broncholytic
Theophylline
Biodegradable (n ¼ 1)
Not relevant
Hormones
17b-estradiol 17a-ethinylestradiol
84 (n ¼ 11) 74 (n ¼ 10)
85 (n ¼ 3) 85 (n ¼ 6)
Contrast media
Diatrizoate Iomeprol Iopamidol Iopromide
9 (n ¼ 2) 9 (n ¼ 2) 9 (n ¼ 2) 33 (n ¼ 5)
30 66 44 64
Cytostatic
Cyclophosphamide Ifosfamide
17 (n ¼ 1) 3 (n ¼ 3)
No ref. No ref.
(n ¼ 5) (n ¼ 3) (n ¼ 2) (n ¼ 7) (n ¼ 2)
(n ¼ 2) (n ¼ 15) (n ¼ 16) (n ¼ 2) (n ¼ 1)
(n ¼ 1) (n ¼ 2) (n ¼ 2) (n ¼ 8) (n ¼ 3)
(n ¼ 1) (n ¼ 1) (n ¼ 2) (n ¼ 5)
AOPs, advanced oxidation processes; n ¼ number of considered references.
wastewater (Ried A, 2009; ITT Wastewater, personal communication). Another example for an ozone-based AOP concerns the improvement of safety and reliability of drinking water-supply system, which has been recently implemented at first time for contaminated groundwater pretreatment. A process flow scheme is given by Figure 16. A part of the raw water is ozonized while hydrogen peroxide is injected to the other part. Both streams are then mixed together passing a static mixer and subsequently entering the oxidation reactor. Optimum process parameters validated by pilot trials are: 2.5 g m3 ozone dosage, H2O2/O3 ratio equal to 0.7 (g g1), and mean hydraulic retention time of 10 min. The plant was designed to treat 64 000 m3 d1 of groundwater contaminated by approximately 20 mg l1 of hydro-chlorinated carbons in two parallel lines. Due to the increased chloride dioxide consumption in the waterworks after startup of AOP system, the dosage of hydrogen peroxide has been reduced to the ratio of 0.5 (g g1) with only slightly decrease to 75% of pollutants removal rate. The operating costs have been estimated to be below 0.01 Euro m3. Another field of application of ozone-based AOPs is the minimization of excess sludge production of aerobic
treatment systems. While several ozone systems have been implemented in Japan, only few case-study systems are under operation in Europe. Currently, this relies mainly on higher sludge disposal costs in Japan compared to Europe. The established systems ozonize the sludge in a cycle loop to aerobic basin, which is an effective tool to minimize sludge production below the yield of 0.11 kg VSS per kgCOD of raw wastewater input. The specific ozone dosage for B100% sludge reduction ranges between 0.12 (Paul and Debellefontaine, 2007) and 0.18 kgO3/kgTSreduced (Sakai et al., 1997). It is worth noting that the 100% reduction has been reached only by release of some suspended solids via final clarifier as well as by increase on soluble COD in the effluent. Detailed information about process description, efficiency, different applications, costs, etc., can be found in Yasui et al. (1996), Sakai et al. (1997), Paul and Debellefontaine (2007), and Sievers et al. (2004). Although the reduction potential is very high and effective, the main competition processes are not processes combined to aerobic system but rather to anaerobic systems with the aim to improve energy yield from waste sludge. The latter one seems to be more sustainable in a period of increasing discussion about importance of renewable energy. Although there are some investigations about combining
Table 9
List of combined ozone-biology applications
Location
Wastewater origin
Target pollutant(s)
Process integration
Flowrate (m3 h1)
Generator size (kgO3 h1)
Ozone dose (kg m3)
Year
Reference
Lang Papier WWTP Ochtrup
Paper industry Textile industry
COD Color þ AOX þ PVAL
Bio-O3-Bio Bio-O3
580 160
2 50 26
0.172 0.075
1999 1992
Ried et al. (2000) Ried et al. (2007)
WWTP Prato, Italy WWTP Ranica, Italya
Textile þ sewage Textile þ sewage
Color þ surfactants Color þ surfactants
Bio-O3 Bio-O3
5000 2500
4 40 2 28
0.032 0.112
1992 2006
Kaulbach (1993) Ried et al. (2009)
Catania, Italy Tubli TSE, Bahrain
Industry þ sewage Sewage
COD Colony forming units
Bio-O3 Bio-O3
4800 8333
4 15 3 48
0.012 0.017
2002 2002
Ried et al. (2007) Ried et al (2007)
SCA-Laakirchen LLTP*** Hellsiek, Germany LLTP Braunschweig, Germany LLTP Mu¨nster, Germany LLTP Fernthal, Germany LLTP Bornum, Germany WWTP Kalundborg, Denmark Wacker Chemie, Germany DOW Bo¨hlen, Germany BASF Schwarzheide– Germany WWTP Regensdorf, Switzerlanda WWTP Ilkeston, Great Britaina
Paper industry Leachate
COD COD
Bio-O3-Bio BioQuint
1100 15
75 24
0.068 0.533
2005 1997
Liechti (2005) Ried et al (1999)
Leachate
COD
Bio-O3-Bio
25
3 12
1.440
1992
Ried et al. (2007)
Leachate
COD
Bio-O3-Bio
10.5
25
0.952
1994
Ried et al. (2007)
Leachate þ extern
COD
Bio-O3-Bio
6.25
34
1.920
1993
Leachate
COD
BioQuint
6.25
8
1.280
1995
Steegmans et al. (1995) Ried et al. (1999)
Industry þ sewage
COD
Bio-O3-Bio
1000
2 90
0.180
2003
Ried et al. (2003)
Chemical industry
COD
Bio-O3
100
43
0.430
2006
Ried et al. (2007)
Chemical industry Chemical industry
COD Nitroaromates
Bio-O3 Bio-O3-Bio
30 20
20 25
0.666
2004 2000
Ried et al. (2007) Ried et al. (2007)
Sewage
Micropollutants
Bio-O3-sand filter
B400
5
0.006
2007
Ried et al. (2009)
Sewage
Micropollutants
Bio-O3
B40
0.8
0.01
2008
Ried et al. (2009)
a
Case studies LLTP***, landfill leachate treatment plant. COD, chemical oxygen demand; AOX, adsorbable organic halogen compounds; PVAL, polyvinylalcohol.
404
Advanced Oxidation Processes H2O2
CIO2 Static mixer
Raw water
Reactor
Water reservoir
Injector
Pump O3 /O2 Figure 16 Ozone/hydrogen peroxide (perozone) process for raw water pretreatment.
AOPs and anaerobic systems onsite at sewage treatment plants, no onsite application exists to date.
separation. The separated condensate is of high quality appropriate for rinse water input, and the up-concentrated electrolyte is also of high quality suitable for input to Watt’s basic electrolyte tank.
4.13.5.2 UV-Oxidation Processes The H2O2/UV process is a well-established AOP. In 1996, more than 200 UV/oxidation treatment installations were put up for process water, groundwater, and drinking water (see, e.g., AOT Handbook Calgon Corp., 1996). Some of the contaminants besides others to be oxidized were: 1,1-dichloroethane (DCA), 1,1,1-trichloroethane (TCA), hydrazine, Nnitrosodimethylamine (NDMA), trichloroethylene (TCE), perchloroethylene (PCE), methyl ethyl ketone, TNT, phenol, vinyl chloride, benzene, chlorobenzene, toluene, xylene, pentachlorophenol (PCP), nitro-glycerin, total petroleum hydrocarbons, PAHs, etc. During the last decade, this process has received increasing interest in industrial wastewater treatment (e.g., electroplating, circuit printing, chemical, pharmaceutical industry, etc.) for degradation of complexing agents, regeneration of electrolytes of surface plating baths, treatment of rinsing water, and process water recovery. An interesting example and representative for a sustainable and economic process-integrated solution is the regeneration of plating baths by UVoxidation of aged electrolyte solutions. This enables not only (1) the degradation of accumulating residuals for less waste emission, but also (2) an increased product surface quality, (3) reduced metal electrolyte consumption, (4) improved water recovery of rinse water leading to zero water discharge, and (5) the use of reaction heat released by UV-oxidation in an integrated concept of evaporation/condensation. This combined process solution is an excellent example for integrating AOPs in industrial production processes reaching less emission up to zero wastewater discharge as well as additional resource-sparing objectives (Dams et al., 2008). The entire process is illustrated in Figure 17 pointing out (1) a nickel electrolyte bath for surface plating, (2) a rinsing cascade, (3) a waste reception tank for aged nickel electrolyte as well as discharged rinsing water, (4) a UV-reaction system, and (5) a Watt’s basic electrolyte tank. The UV-reaction system is treating the mixture of rinsed wastewater and aged nickel electrolyte in a loop. The TOC concentration decreases by UV-oxidation, and parallel to this the electrolyte is up-concentrated by water evaporation and condensate
4.13.5.3 Fenton Process The classical Fenton process is probably one of the oldest AOPs applied to remove hazardous compounds from wastewater. This process has been widely applied in different fields such as paper, chemical, pharmaceutical, textile, de-inking, TNT-production, and metal industry. It has been usually applied for wastewaters with COD range from 1 to 100 g l1 as a cost-efficient alternative to wet air oxidation (WAO) or incineration. Some authors identified the classical Fenton process as sometimes more cost efficient to the H2O2/UV or other AOPs (see Pignatello et al., 2006); however, this cannot be generalized, because the cost efficiency depends on so many different factors that each application requires a careful engineering and subsequent cost calculation. Although the optimal temperature for Fenton process has been found at approximately 50–60 1C, many applications operate at higher temperatures up to 100–130 1C. In these applications a second catalyst is often mentioned besides Fe, probably Cu.A compilation of fields of Fenton process application can be found in Suty et al. (2004). Recently, different modified Fenton processes received increasing interest in research and development up to pilot scale, namely (1) photo-Fenton, (2) solar-Fenton, (3) electroFenton, and (4) heterogeneous fixed-bed Fenton.
4.13.5.4 Wet Air Oxidation There exist some well-established applications in WAO, for example, (1) Bayer LOPROXs process, (2) WAO ATHOSs, (3) Zimmermann process (ZIMPRO), and (4) VerTech deep shaft. All these processes have similar operating conditions, high pressures and high temperatures – up to subcritical conditions. While the LOPROX process operates at 3–20 bar and 120–200 1C at hydraulic retention times (HRT) below 3 h, the ATHOS process reaches approximately 54 bar and 250 1C. Both processes use homogeneous Cu ion (if needed) as catalyst and oxygen as oxidant. The ZIMPRO process has been the first WAO introduced by Zimmermann (1958) with several installations in USA while the VerTech process is unique using
Advanced Oxidation Processes
405
Plating process
Rinsing process
Product
Product
Nickel electrolyte 1
2 Bath /rinse discharge Renewed rinse water
Heat pump
Cooling tower
UV reactor
4
3
5
Waste reception tank
Watt’s basic electrolyte
Oxidant Figure 17 Process scheme for water, heat, and metal recovery by integrated application of UV-oxidation type Enviolets (Dams et al., 2008).
a 500–1200 m deep-shaft reaction system. The WAOs are often characterized as relative expensive processes; however, the economic application seems to be possible for larger installations between 3000 and 15 000 tons of total solids per year, as the recently increased number of ATHOSs applications indicates. This process enables mineralization of 80–90% of COD and 95% sludge volume reduction.
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Relevant Websites http://www.almaden.ibm.com IBM Almaden Research Center.
4.14 Biological Nutrient Removal GA Ekama, University of Cape Town, Cape Town, South Africa & 2011 Elsevier B.V. All rights reserved.
4.14.1 4.14.2 4.14.3 4.14.4 4.14.4.1 4.14.4.2 4.14.4.2.1 4.14.4.2.2 4.14.4.2.3 4.14.4.3 4.14.4.3.1 4.14.4.3.2 4.14.4.3.3 4.14.4.3.4 4.14.4.3.5 4.14.4.4 4.14.4.4.1 4.14.4.4.2 4.14.4.4.3 4.14.4.4.4 4.14.4.5 4.14.4.6 4.14.4.7 4.14.5 4.14.5.1 4.14.5.1.1 4.14.5.1.2 4.14.5.1.3 4.14.5.2 4.14.5.2.1 4.14.5.2.2 4.14.6 4.14.6.1 4.14.6.2 4.14.6.3 4.14.6.4 4.14.7 4.14.7.1 4.14.7.1.1 4.14.7.1.2 4.14.7.1.3 4.14.7.1.4 4.14.7.1.5 4.14.7.1.6 4.14.7.1.7 4.14.8 4.14.9 4.14.9.1 4.14.9.2 4.14.9.3 4.14.9.3.1 4.14.9.3.2 4.14.9.4 4.14.9.5
Introduction System Configuration and Organism Groups Transformations in the Biological Reactor Wastewater Characterization Introduction Carbonaceous Organic (C) Materials Carbonaceous material (COD) fractions Quantification of COD fractions Analytical formulation for COD Nitrogenous Materials Nitrogenous material fractions Quantification of N fractions (Nti) Analytical formulation Maximum specific growth rate of nitrifiers at 20 1C Typical wastewater TKN characteristics Phosphorous materials Phosphorus fractions Quantification of P fractions Analytical formulation Typical wastewater phosphorus characteristics Inorganic Dissolved, Settleable, and Nonsettleable Solids Other Materials Wastewater Characterization for Plant Wide Modeling Modeling Biological Behavior Biological Growth Behavior Stoichiometry and kinetics Monod growth kinetics for utilization of RBSO Active site surface kinetics for hydrolysis/utilization of BPO Organism Decline Endogenous respiration Death regeneration AS System Constraints Mixing Regimes Solids Retention Time or Sludge Age Nominal Hydraulic Retention Time Connection between Sludge Age and Hydraulic Retention Time Model Development – Completely Mixed Aerobic System Building Up the Model in Stages Unbiodegradable particulate organics (Supi) Unbiodegradable soluble organics (Susi) Biodegradable organics Complete utilization of soluble biodegradable organics The mass balance on oxygen Complete utilization of BPOs Integration of biodegradable and unbiodegradable organics models The COD (or e ) Mass Balance The AS System Steady-State Equations for Real Wastewater Effluent COD Concentration ISS Concentration Process Design Equations For the influent For the system Active Fraction of the Sludge Steady-State Design Chart
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4.14.9.6 4.14.10 4.14.11 4.14.11.1 4.14.11.2 4.14.11.3 4.14.12 4.14.12.1 4.14.12.2 4.14.13 4.14.14 4.14.14.1 4.14.14.2 4.14.14.3 4.14.15 4.14.15.1 4.14.15.1.1 4.14.15.1.2 4.14.15.2 4.14.15.3 4.14.15.3.1 4.14.15.3.2 4.14.15.3.3 4.14.16 4.14.17 4.14.18 4.14.18.1 4.14.18.2 4.14.18.3 4.14.19 4.14.19.1 4.14.20 4.14.20.1 4.14.20.2 4.14.20.3 4.14.20.3.1 4.14.20.4 4.14.20.5 4.14.20.6 4.14.21 4.14.21.1 4.14.21.2 4.14.22 4.14.22.1 4.14.22.2 4.14.22.3 4.14.23 4.14.23.1 4.14.23.2 4.14.24 4.14.24.1 4.14.24.2 4.14.24.3 4.14.24.4 4.14.24.5 4.14.24.5.1 4.14.24.5.2 4.14.25 4.14.25.1 4.14.25.2
The Calculation Procedure Reactor Volume Requirements Determination of Reactor TSS Concentration Reactor Cost SST Cost Total Cost Carbonaceous Oxygen Demand Steady-State (Daily Average) Conditions Daily Cyclic (Dynamic) Conditions Daily Sludge Production System Design and Control System Sludge Mass Control Hydraulic Control of Sludge Age Flow and Load Equalization Tanks Selection of Sludge Age Short Sludge Ages (1–5 days) Conventional plants Aerated lagoons Intermediate Sludge Ages (10–15 Days) Long Sludge Ages (20 Days or More) Aerobic plants Anoxic–aerobic plants Anaerobic–anoxic–aerobic plants Sludge Age – The Dominant Driver for Size Nitrification – Introduction Nitrification Biological Kinetics Growth Growth Behavior Endogenous Respiration Nitrification Process Kinetics Effluent Ammonia Concentration Factors Influencing Nitrification Influent Source Temperature Unaerated Zones Maximum allowable unaerated mass fraction DO Concentration Cyclic Flow and Load pH and Alkalinity Nutrient Requirements for Sludge Production Nitrogen Requirements N (and P) Removal by Sludge Production Nitrification Design Considerations Effluent TKN Nitrification Capacity Mass of Nitrifiers (MXA) and Nitrification Oxygen Demand (FOn) Nitrification Design Example Wastewater Characteristics Nitrification Process Behavior Biological Denitrification Interaction between Nitrification and N Removal Benefits of Denitrification N Removal by Denitrification Denitrification Kinetics Denitrification Systems The Ludzack–Ettinger system The four-stage Bardenpho system Denitrification Kinetics Denitrification Rates Denitrification Potential
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Biological Nutrient Removal 4.14.25.3 4.14.25.4 4.14.26 4.14.26.1 4.14.26.2 4.14.26.3 4.14.26.3.1 4.14.26.3.2 4.14.26.3.3 4.14.26.3.4 4.14.27 4.14.27.1 4.14.27.2 4.14.28 4.14.28.1 4.14.28.2 4.14.28.3 4.14.28.3.1 4.14.28.3.2 4.14.28.3.3 4.14.28.3.4 4.14.28.3.5 4.14.28.3.6 4.14.28.3.7 4.14.28.3.8 4.14.28.3.9 4.14.29 4.14.30 4.14.30.1 4.14.30.2 4.14.30.2.1 4.14.30.3 4.14.30.3.1 4.14.30.3.2 4.14.30.3.3 4.14.30.3.4 4.14.31 4.14.31.1 4.14.31.1.1 4.14.31.1.2 4.14.31.1.3 4.14.31.1.4 4.14.31.1.5 4.14.31.1.6 4.14.31.1.7 4.14.31.2 4.14.31.3 4.14.31.3.1 4.14.31.3.2 4.14.31.3.3 4.14.32 4.14.32.1 4.14.32.2 4.14.32.3 4.14.33 4.14.33.1 4.14.33.2 4.14.33.3 4.14.33.4 4.14.33.5
Minimum Primary Anoxic Sludge Mass Fraction Denitrification – Influence on Reactor Volume and Oxygen Demand Development and Demonstration of Design Procedure Review of Calculations Allocation of Unaerated Sludge Mass Fraction Denitrification Performance of the MLE System Optimum recycle ratio a The balanced MLE system Effect of influent TKN/COD ratio MLE sensivity diagram System Volume and Oxygen Demand System Volume Daily Average Total Oxygen Demand Biological Excess Phosphorus Removal Introduction Principles of BEPR Mechanism of BEPR Background Biological P removal microorganisms Prerequisites Observations Biological P removal mechanism Fermentable COD and slowly biodegradable COD Functions of the anaerobic zone Influence of recycling oxygen and nitrate to the anaerobic reactor Denitrification by PAOs Principles of Maximizing BEPR Model Development for BEPR Early Developments RBO and Anaerobic Mass Fraction NDBEPR system kinetics Enhanced PAO Cultures Enhanced culture development Enhanced culture kinetic model Simplified enhanced culture steady-state model Steady-state mixed culture NDBEPR systems Mixed Culture Steady-State Model Division of Biodegradable Organics between PAOs and OHOs Subdivision of influent RBO Conversion of FBSO Effect of recycling nitrate or oxygen Steady-state FBSO conversion equation Mass of VSS in the NDBEPR system PAO P release P removal VSS and TSS Sludge Masses in the Reactor (System) BEPR System Design Considerations Process volume requirements Nitrogen requirements for sludge production Total oxygen demand Influence of BEPR on the System Influence on VSS, TSS, and Carbonaceous Oxygen Demand VSS Composition P/VSS ratio Factors Influencing Magnitude of BEPR Sludge Age and Anaerobic Mass Fraction Raw and Settled Influent Influence of Influent RBO Fraction Influence of Recycling Nitrate and Oxygen to the Anaerobic Reactor Subdivision of the Anaerobic Reactor into Compartments
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4.14.34 4.14.34.1 4.14.34.2 4.14.34.3 4.14.34.4 4.14.34.5 4.14.35 4.14.36 References
Denitrification in NDBEPR Systems Introduction Experimental Basis for Denitrification Kinetics in NDBEPR Systems Denitrification Potential in NDBEPR Systems Principles of Denitrification Design Procedures for NDBEPR Systems Analysis of Denitrification in NDBEPR Systems Denitrification in the UCT System Conclusion
4.14.1 Introduction To comply with more stringent effluent legislation, the functions of the activated sludge (AS) system have expanded to progressively include the biological removal of carbon (C), nitrogen (N), and phosphorus (P). Not only have these expansions increased the complexity of the system configuration and its operation, but, concomitantly, the number of biological processes influencing the effluent quality and the number of compounds involved in these processes have increased. With such complexities, designs based on experience or semi-empirical methods no longer will give optimal performance; design procedures based on more fundamental behavioral patterns are required. To meet this requirement, over the past two decades various research groups have contributed to developing conceptual and mathematical steadystate design and dynamic kinetic simulation models for the biological nutrient removal (BNR) AS system. These models have progressively included aerobic chemical oxygen demand (COD) (carbon) removal and nitrification (Marais and Ekama, 1976; Dold et al., 1980), anoxic denitrification (van Haandel et al., 1981; WRC, 1984; Dold et al., 1991 [UCTOLD]; Henze et al., 1987 [ASM1]), and anaerobic/anoxic/aerobic biological excess phosphorus removal (BEPR; Wentzel et al., 1990, 1992 [UCTPHO]; Henze et al., 1995 [ASM2]). The models enable system design and operational parameters to be readily identified, provide guidance in selecting values for these parameters, and quantify the expected behavior of the system. For mathematical modeling of wastewater-treatment systems, generally two levels of mathematical models have been developed, steady state and dynamic kinetic simulation. The steady-state models have constant flows and loads and are simple to use. This simplicity makes these models very useful for design. In these models, complete descriptions of system parameters are not required, but rather the models are oriented to determine the important system design parameters from performance criteria using algebraic equations with explicit solutions. The dynamic models are much more complex than the steady-state ones and have varying flows and loads with the result that time is included as a parameter. Although dynamic simulation models are useful for predicting the timedependent system response of an existing or proposed system, they comprise time-based differential equations which require numerical integration with computer software to generate solutions. Also, their complexity demands that many more kinetic and stoichiometric constants need to be supplied and
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all the system design parameters have to be specified. In fact, the steady-state models are very useful for calculating the initial conditions required to build dynamic simulation models (such as reactor volumes, recycle and waste flows, and values for the various concentrations in the reactor(s)) and for cross-checking simulation model outputs. Hence, steady-state models based on the same bioprocess principles but with simplified equations that yield closely similar results are not only a very useful complement to the complex dynamic simulation models but also give insight into the dynamic models. In this chapter, attention is focused on the steadystate models for biological organics (COD, C), N, and P removal, and where relevant their links to the dynamic models are discussed. First, an overview of the fundamental principles and functional organisms on which both the steady-state and dynamic kinetic models are based is presented.
4.14.2 System Configuration and Organism Groups The expansion in function of the AS system to include biological N and P removal has been accomplished by manipulating: (1) the system configuration, through the incorporation of multiple in-series reactors with various interreactor recycles or timed aeration cycles, some aerobic and others not and (2) the wastewater characteristics, through primary sedimentation and acid fermentation of primary sludge. The objective of these manipulations is to create environmental conditions in the AS system that are conducive to the optimal growth and action of organisms that naturally perform the biological reactions necessary to treat the wastewater – aerobic zones/periods for nitrification and organics removal, anoxic zones/periods for denitrification and organics removal, and anaerobic/aerobic sequence of zones/periods with the influent fed to the anaerobic zone/period for biological excess P removal (BEPR). In the highly diverse mixed cultures that develop in these AS systems, for the purpose of design, only the behavior of whole populations or groups of organisms with the same function is considered. The principal organism groups, their functions, and the zones in which these functions are performed are summarized in Table 1. From Table 1, for the design of nitrification–denitrification (ND) BEPR-AS systems, three organism groups and their interactions need to be taken into account (Wentzel et al., 1992): (1) heterotrophic organisms unable to accumulate polyphosphate (polyP), termed ordinary heterotrophic organisms (OHOs); (2) heterotrophic organisms able to
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Table 1 Principal organism groups included in models for activated sludge systems, their functions and the zones in which these functions are performed Organism group
Biological process
Ordinary heterotrophic organisms (OHOs, unable to accumulate polyphosphate)
COD removal and ammonification
Phosphorus accumulating (heterotrophic) organisms (PAOs, accumulate polyphosphate).
Autotrophic nitrifier organisms (ANOs)
Zone Aerobica þ
(organic degradation; release of organic N as ammonia, NH4 ) Denitrification (organic degradation; ammonification; reduction of nitrate nitrite – NO3 -NO2 -N2 ) Fermentation (conversion of FBSO to VFA)
Anoxica
Anaerobica
P release (VFA uptake; PHA storage) P release (VFA uptake; PHA storage) P uptake (PHA degradation; denitrification?) P uptake; P removal (PHA degradation; DO uptake)
Anaerobic Anoxic Anoxic Aerobic
Nitrification (NH4 þ -NO2 -NO3 ; DO uptake)
Aerobic
a
In this chapter, the following working definitions apply: aerobic – presence and/or influx of dissolved oxygen (DO) and nitrate/nitrite; anoxic – absence and zero influx of DO but presence and/or influx of nitrate/nitrite; Anaerobic – absence and/or no influx of dissolved oxygen (DO) and nitrate/nitrite.
Enmeshed with sludge mass
Biodegradable
Transforms to active organisms
Settleable Suspended Precipitable
Enmeshed with sludge mass Transforms to set. solids
Biologically Transferred utilizable to Nonprecip and bio util
Biomass in reactor all settleable none supended
Unbiodegradable
Unbiodegradable
Inorganic mass all settleable none suspended
Biodegradable
Transforms to active organisms Enmeshed with sludge mass Transforms to active organisms
Biodegradable
Organic volatile settleable solids (VSS)
Escapes with effluent
Inorganic set. solids (ISS)
Unbiodegradable
Sludge constituents
Reaction
Total settleable solids (TSS)
Soluble Dissolved Organic
Particulate Suspended Settleable Partic
For the AS system, it is necessary to characterize the wastewater physically (soluble, nonsettleable (colloidal and/or suspended), settleable, organic, and inorganic) and biologically (biodegradable and unbiodegradable). The physical, chemical, and biological transformations of the organic and inorganic wastewater constituents that take place in the biological reactor are outlined in Figure 1. Some of these transformations are important for achieving the required effluent quality, while others are not important for the effluent quality but are important for the system design and operation. In the Figure each of the wastewater organic and inorganic fractions have soluble and particulate fractions, the latter of which subdivides further into suspended (nonsettleable) and settleable ones. Each of the three organic subfractions, in turn, has biodegradable and unbiodegradable constituents. The inorganic particulate subfraction comprises both settleable and suspended (nonsettleable) constituents, while the soluble inorganic subfraction comprises both precipitable and nonprecipitable and biologically utilizable and nonbiologically utilizable constituents. In the biological reactor the biodegradable organics, whether soluble, nonsettleable, or settleable, are transformed to OHOs (XBH), which become part of the organic (volatile) suspended solids (VSSs) in the reactor. When these organisms die, they leave behind unbiodegradable particulate (but not soluble) organics, called endogenous residue, comprising
Inorganic
4.14.3 Transformations in the Biological Reactor
Wastewater constituents
Soluble
accumulate polyP, generically called phosphorus accumulating organisms (PAOs); and (3) autotrophic nitrifier organisms (ANOs) mediating nitrification. This chapter focuses on all three groups which together accomplish carbon (COD), N, and P removal via their normal bioprocesses.
Solids Gas
Escapes as gas
Escapes with effluent
Figure 1 Global transformation reactions of organic and inorganic wastewater constituents from the particulate and soluble forms in the solid and liquid phases to the solid phase as sludge constituents, and gas and liquid phases escaping to the atmosphere and with the effluent, respectively.
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mainly unbiodegradable cell wall material (XEH). This endogenous residue becomes part of the VSS mass in the reactor. If the reactor includes anaerobic zones/period, the readily biodegradable organics are converted to short-chain fatty acids and transformed to PAOs (XBG), which also become part of the VSSs in the reactor. When the PAO organisms die, they also leave behind unbiodegradable particulate (but not soluble) organics, called endogenous residue, comprising mainly unbiodegradable cell-wall material (XEG). The unbiodegradable suspended and settleable organics (XI) from the influent become enmeshed with the OHO and endogenous residue masses. Together, these five constituents (XBH þ XEH þ XBG þ XEG þ XI) form the organic component of the settleable solids that accumulates in the biological reactor (VSS, Xv) of the BNR system. If the system comprises only aerobic or anoxic and aerobic zones/periods, the PAOs will be absent, and the organic VSS (Xv) will comprise only the OHOs (XBH) and their endogenous residue (XEH) and the unbiodegradable particulate organics (UPO, XI) from the influent. The inorganic settleable and suspended constituents, together with the precipitable soluble inorganics from the influent, form the fixed inorganic component of the settleable solids mass (ISS) in the reactor. The sum of the organic (VSS) and inorganic settleable solids is the total suspended solids (TSS, Xt). The biologically utilizable soluble inorganics are absorbed by the biomass and become part of it or are transformed to the gas phase, in which case they escape to the atmosphere. The soluble inorganics (anions and cations) absorbed by the OHO and PAO biomass become part of the biomass intracellular dissolved constituents. When the solids from the reactor (TSS) are dried in the VSS and TSS tests procedure, these intracellular dissolved inorganics precipitate and add to the (fixed) ISS from the influent. This is particularly relevant for BEPR systems which stimulate the growth of PAOs – their polyP content adds up to 8 times more inorganics content than the OHOs. The nonprecipitable and nonbiologically utilizable soluble inorganics escape with the effluent. Because of the efficient bioflocculation capability of the organic AS mass, all the solids material, whether biodegradable or unbiodegradable, organic or inorganic, get enmeshed with and become part of not or the settleable solids. Very little suspended or colliodal (nonsettleable) solids mass is formed in the reactor, but when it does it cannot be retained in the system anyway and escapes with the effluent, unless membranes are used for solid–liquid separation instead of secondary settling tanks.
4.14.4 Wastewater Characterization 4.14.4.1 Introduction The AS system comprises a biological reactor and a secondary settling tank (Figure 2). Irrespective of whether or not biological N and/or P removal are included, many different biological and physical processes take place in the biological reactor, and the physical process sedimentation takes place in the secondary settling tank. These processes form the basis for subdividing the influent wastewater C, N, and P materials into subfractions. On entry of the influent into the biological reactor, the particulate (suspended) materials, which include
Aeration
Influent
Aerobic reactor
Waste flow Secondary settling tank Effluent
Sludge recycle Figure 2 Activated sludge system with biological reactor and secondary settling tank with excess sludge withdrawal direct from the reactor.
both settleable and nonsettleable, organic and inorganic material, are enmeshed (a biologically mediated flocculation) and become part of the AS mixed liquor, which is virtually all settleable (Figure 1). The soluble (dissolved) materials, both organic and inorganic, remain in solution. In the biological reactor, the bacteria present will act on the biologically utilizable material, termed ‘biodegradable’, whether organic or inorganic, soluble or particulate, and transform these to other compounds or products, either gaseous, soluble, or particulate: The gaseous products escape to the atmosphere, the particulate products become (or remain) part of the mixed liquor solids, and the soluble products become (or remain) dissolved in solution. The nonbiologically utilizable organics, termed ‘unbiodegradable’, will not be transformed and will remain in either the soluble or particulate form. Therefore, recognizing that biological processes take place in the reactor, the first major division of the influent is based on whether the material is ‘biodegradable’ or ‘unbiodegradable’. After biological treatment the flow passes from the biological reactor to the solid–liquid separation system, sometimes membrane reactor but usually a secondary settling tank. In the secondary settling tank, the bioflocculated particulate materials making up the mixed liquor (whether organic or inorganic, biodegradable, or unbiodegradable) settle out and are returned to the biological reactor. The particulate components of the mixed liquor entering the settling tank are thus retained in the system. All the soluble components of the mixed liquor (whether organic or inorganic, biodegradable, or unbiodegradable) cannot settle out and escape with the effluent. The settling behavior in the secondary settling tank therefore forms the basis for subdividing the influent unbiodegradable material into subfractions: the influent unbiodegradable material passes unmodified through the biological reactor to the secondary settling tank; ideally, all the particulate (including the enmeshed influent nonsettleable particulates) material settles out in the secondary settling tank and these constituents are therefore termed unbiodegradable particulate; the soluble constituents cannot settle out so that these constituents are termed unbiodegradable soluble. With regard to the influent biodegradable material, because practically all of this material gets biologically transformed to biomass in the biological reactor preceding the secondary settling tank, it cannot be subdivided into subfractions based on its behavior in the secondary settling tank; subdivision of the biodegradable material is based on the rates of transformation/ utilization by the bacteria in the biological reactor. As it
Biological Nutrient Removal
happens, the soluble organic constituents are more easily utilizable than the particulate ones, so that the physical size of the organics also plays an important role in their rate of utilization. For this reason, physical separation tests can be used to assist in the identification of the readily and slowly biodegradable organic material (COD) fractions. From the above, to assess the performance of the AS system, the wastewater constituents need to be characterized: (1) biologically, that is, as biodegradable (biologically utilizable) or unbiodegradable (biologically nonutilizable) material and (2) physically, that is, as soluble or particulate material. Therefore, for steady-sate and dynamic kinetic models based on fundamentals of biological behavior for AS systems, with or without biological N and P removal treating raw or settled wastewater, it is necessary to divide the influent constituents into at least three fractions: (1) biodegradable, (2) unbiodegradable soluble, and (3) unbiodegradable particulate. This general, but not complete, wastewater characterization structure conforms to the biological degradation and physical solid/liquid separation processes that take place in the AS system. When only organic (C) material removal is considered, this structure is applied to the organic (COD or carbonaceous) constituents of the wastewater; with additional nitrification or nitrification and biological N removal, it is also applied to the N constituents; with C, N, and P material removal, it is applied to all three of these groups. The quantity or concentration of each constituent fraction is assessed chemically. In AS models, the COD test forms the basis for specifying the various fractions of organic or carbonaceous (C) material, the total Kjeldahl nitrogen (TKN) and the free and saline ammonia (FSA) tests form the basis for specifying the various nitrogen (N) constituents, and the total phosphorus (TP) and orthophosphate (OP) tests form the basis for specifying the phosphorous (P) constituents of the wastewater. This section provides some detail on the characterization of the COD, N, and P constituents of the wastewater using these tests.
4.14.4.2 Carbonaceous Organic (C) Materials The COD test measures the electron-donating capacity of the organics in the wastewater, which closely approximates the free energy available in the organics (WRC, 1984). For AS system design, it is necessary to quantify, to various degrees, the constituents making up the organic (C) material (measured as COD), as these significantly affect the system response, for example, carbonaceous oxygen demand, sludge production, denitrification, and phosphorus removal. The extent of characterization required for the organic materials depends on the design objectives for the AS system – if N and/or P removal are incorporated, information additional to the three groups of organics is required, which are presented below.
4.14.4.2.1 Carbonaceous material (COD) fractions The first division of the influent COD (Sti) is based on whether the COD fraction undergoes biological degradation or not, that is, into biodegradable COD (Sbi) and unbiodegradable COD (Sui) respectively.
415
Unbiodegradable subfractions. The influent unbiodegradable organics (COD) are subdivided into two fractions, unbiodegradable soluble organic COD (USO, Susi) and UPO COD (Supi). Both fractions are accepted to be unaffected by biological action in the system so that at steady state, the fluxes (mass/d) of these materials that enter the system are equal to the fluxes that exit the system. As both fractions are unbiodegradable, their differentiation is based on their behavior in the secondary settling tank. The USO (Sus) passes out in the secondary settling tank overflow and appears as COD in the effluent. Since the USO (Sus) flows out with the effluent, it has a direct influence on the effluent COD concentration. It can be accepted that for AS systems with sludge ages greater than about 3 days, the effluent soluble COD (say o0.45 mm filtered) (Suse) is closely equal to the influent unbiodegradable soluble COD (Susi) (Ekama et al., 1986). This assumes that (1) the influent (and any generated) soluble biodegradable COD has been completely utilized and (2) negligible unbiodegradable soluble organics are generated during biological treatment in the reactor. Over the many years of research into AS systems, both assumptions have come to be accepted as reasonable and are implicitly incorporated in most models, for example, the steady-state one of Marais and Ekama (1976) and WRC (1984). The latter assumption also has been incorporated in the more complex mixed culture AS simulation models, for example, the IWA-AS model No 1 (ASM1) (Henze et al., 1987), UCTOLD (Dold et al., 1991), UCTPHO (Wentzel et al., 1992), and IWA ASM2 (Henze et al., 1995). A literature review by Dold et al. (1986) demonstrates the validity of this latter assumption. Experimental work by Torrijos et al. (1994), in which the respirometry and utilization of the more exactly defined colloidal (0.2–50 mm) and soluble (o0.1 mm) organic fractions from real wastewater were examined, further validates this assumption, and in addition two other important assumptions included in the kinetic models, that is, (1) that the soluble and colloidal organic fractions are utilized simultaneously but at markedly different rates and (2) that variability in utilization rates of different sized organics within the soluble (o0.1 mm) group is very small and markedly faster than the organics in the colloidal group, so that insofar as kinetics of utilization and hence wastewater characterization are concerned, it is sufficiently accurate to recognize only two biodegradable fractions: a soluble rapidly biodegradable one comprising dissolved organics and a lumped particulate (suspended) slowly biodegradable one comprising both nonsettleable and settleable organics. The UPOs (Sup), such as paper, hair, and other fibrous material, are enmeshed in the biological reactor mixed liquor which settles out in the secondary settling tank. Thus, UPOs (Sup) are retained in the system to accumulate as UPOs (VSS), which essentially are all settleable. At steady state, the flux of Sup entering the system with the influent is equal to the flux of this material, enmeshed with the mixed liquor VSS (and denoted XI), exiting via the sludge waste stream (Figure 2). From a mass balance, the mass of unbiodegradable organic solids that accumulate in the reactor from the influent is equal to the daily influent mass load of this material multiplied by the sludge age. Thus, the influent UPO (Supi) has a direct effect on the mixed liquor solids (VSS) mass in the reactor, and
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therefore on the system volume requirements for a selected mixed liquor solids concentration. Unlike the USO (Sus), of which negligibly little is generated by the biological reactions, UPO material is generated in significant quantities by the biological reactions and also enmeshed (and hence all settleable) in the biological reactor mixed liqour. Owing to the different origin of this unbiodegradable particulate material, called endogenous residue (and denoted XEH), it is accounted for separately from the influent UPOs that accumulate in the reactor (XI), so that in the AS model, the contribution of each of these two UPO materials to the mixed liquor organic concentration in the reactor can be assessed. Biodegradable subfractions. Subdivision of the biodegradable organics (Sbi) into biodegradable soluble (BSO, Sbsi) and particulate (BPO, Sbpi) subfractions depends on the requirements for the system to be designed. For a completely aerobic system, irrespective of whether nitrification is included or not either intentionally or unintentionally, subdivision of the biodegradable organics into its subfractions is not required for design. Knowing the influent biodegradable COD concentration and the average flow per day gives the biodegradable COD load (flux, kgCOD/d) on the plant; knowing the biodegradable COD load (FSbi, where the prefix F denotes flux) and selecting a sludge age, the daily carbonaceous oxygen requirements (FOc, kgO d1) and the active OHO mass (MXBH, where the prefix M denotes mass, kgVSS) and UPO masses (MXEH þ MXI) that make up the mixed liquor organics in the
reactor (MXv ¼ MXBH þ MXEH þ MXI), which conventionally are measured with the VSSs but can be measured with the COD also (as in ASM1 and 2), can be estimated from the steady-state model (see Section 4.14.9.3). However, if denitrification and/or biological P removal are included, then subdivision of influent biodegradable organics (Sbi) into subfractions is essential. The first subdivision of Sbi is into readily biodegradable (soluble) COD (Sbsi) and slowly biodegradable (particulate) COD (Sbpi) (Figure 3). This division has been based on observed biological responses of AS mixed liquor to domestic wastewater (Dold et al., 1980; AS (short sludge age cyclic loading, plugflow reactors, batch tests)); two distinct rates of utilization of domestic wastewater biodegradable COD were apparent with either oxygen or nitrate as electron acceptor (aerobic or anoxic conditions, respectively) (Dold et al., 1980; van Haandel et al., 1981; Ekama et al., 1986; Still et al., 1996; Gabb et al., 1991; Ekama et al., 1996a). A fraction (called readily biodegradable soluble organics, RBSO) was taken up rapidly by the sludge and metabolized, giving rise to a high oxygen or nitrate utilization rate, respectively. The other fraction (called slowly biodegradable particulate organics (BPOs), COD) was taken up much more slowly and metabolized, giving rise to oxygen or nitrate utilization rates about 1/10 of the rate with RBSO. To explain these observations, the RBSO was hypothesized to consist of simple soluble molecules that can be readily absorbed by the organism and metabolized for energy and cell synthesis, whereas the BPO was assumed to be
Total COD (S ti)
Biodegradable (S bi)
Heterotroph active biomass (Z BHi)
Slowly biodegradable (particulate) (Sbpi)
Readily biodegradable (soluble) (Sbsi)
Short-chain fatty acids (SCFA) (Sbsai)
Unbiodegradable (S ui)
Unbiodegradable particulate (Supi)
Unbiodegradable soluble (Susi)
Fermentable RBCOD (F-RBCOD) (Sbsfi)
Figure 3 Complete subdivision of the influent organic material (measured as COD) showing the five fractions required for steady-state design of biological N and P removal systems and dynamic modeling of the fully aerobic and anoxic–aerobic (for N removal) activated sludge systems; although usually zero, an active heterotrophic organism (OHO) concentration is also shown for wastewaters that may contain significant concentrations of this.
Biological Nutrient Removal
made up of particulate/complex organic molecules that require extracellular adsorption and enzymatic breakdown (hydrolysis) prior to absorption and utilization (Dold et al., 1980; Torrijos et al., 1994; van Haandel et al., 1981), that is, the division is a biokinetic one. Under dynamic loading of activated sludge (short sludge age cyclic loading, plugflow reactors, batch tests) two distinct rates of utilization of domestic wastewater biodegradable COD were apparent with either oxygen or nitrate as electron acceptor (aerobic or anoxic conditions respectively) (Dold et al., 1980; van Haandel et al., 1981; Ekama et al., 1986; Still et al., 1985, 1996; Gabb et al., 1991; Ekama et al., 1996a). A fraction (called readily biodegradable COD, RBCOD) was taken up rapidly by the sludge and metabolized, giving rise to a high oxygen or nitrate utilization rate respectively. The other fraction (called slowly biodegradable COD, SBCOD) was taken up much more slowly and metabolized, giving rise to oxygen or nitrate utilization rates about 1/10 of the rate with RBCOD. To explain these observations, the RBCOD was hypothesized to consist of simple soluble molecules that can be readily absorbed by the organism and metabolized for energy and cell synthesis, whereas the SBCOD was assumed to be made up of particulate/complex organic molecules that require extracellular adsorption and enzymatic breakdown (hydrolysis) prior to absorption and utilization (Dold et al., 1980; Torrijos et al., 1994). The difference in molecule size between RBSO and BPO has been used to classify the RBSO as a biodegradable soluble COD and the BPO as a biodegradable particulate COD. This difference in molecule size of the RBSO and BPO has also led to the development of physical separation methods to assist in the quantification of these organic fractions (e.g., Dold et al., 1986; Ekama et al., 1986; Mamais et al., 1993; Wentzel et al., 1995, 1999, 2000; Mbewe et al., 1994, 1998). However, it must be remembered that the physical distinction between soluble and BPOs (COD) is not directly related to the behavior in the secondary settling tank, as is the case for the unbiodegradable COD subfractions, but is an approximation of the difference in response of the organisms to the two biodegradable COD fractions. The RBSO is soluble and therefore is exposed to biological treatment only as long as the liquid remains in the reactor, that is, for the hydraulic retention time, which is relatively short (B6–24 h). However, the rate of RBSO utilization is high and for sludge ages greater than about 3 days, the concentration of RBSO in the effluent is negligible even though the retention time is relatively short. Accordingly, for design of fully aerobic systems, knowledge of the influent RBSO concentration is not required – it can be safely assumed that all the RBSO will be utilized in the system. For the BPO, the extracellular breakdown (hydrolysis) is slow and is the limiting rate in the utilization of BPO. Although the rate of BPO utilization is relatively slow, the BPO does not appear in the effluent. This is because on entry of the influent into the bioreactor, the BPO becomes enmeshed in the mixed liquor, settles out in the secondary settling tank, and is retained in the system. Therefore, unlike the soluble biodegradable organics (RBSO, BSO) which are exposed to biological treatment for only as long as the liquid remains in the system, that is, hydraulic retention time, the biodegradable
417
particulate organics (BPO) are exposed to biological treatment for as long as the particulate (now settleable) material is retained in the system (i.e., for the sludge age). Therefore, even though the utilization of the BPO is around (1/10)th that of the RBSO, because the sludge age of most AS systems is usually more than 10 times longer than the hydraulic retention time, the BPO usually is completely utilized as well. From simulation studies using dynamic kinetic models (such as the one of Dold et al. (1980, 1991)) all the BPO is completely utilized for sludge ages greater than about 2 or 3 days and temperatures greater than about 20 1C (5–6 days at 14 1C). Accordingly, for design of fully aerobic systems using the steady-state model, knowledge of the RBSO (BSO) and BPO (UPO) subdivision is not required – it is sufficient to assume that all the biodegradable COD will be utilized in the system. However, when denitrification and/or BEPR are included, knowledge of the RBSO is essential and a more refined characterization as defined in Figure 3 is required. For denitrification, the rate of denitrification depends on, inter alia, whether RBSO or BPO serves as electron donor (substrate), and the relative proportion of these two organic types will thus influence the amount of N removal. For BEPR the magnitude of the phosphorus removal is strongly linked to the influent RBSO concentration. Furthermore, with BEPR, the RBSO needs to be subdivided into two subfractions (Figure 3) (Wentzel et al., 1990). With BEPR, the organisms mediating BEPR, called PAOs, take up and store intracellularly volatile fatty acids (VFAs) in the anaerobic reactor with associated P release (called sequestration). The amount of VFA that the PAOs take up in the anaerobic reactor determines the proportion of the biodegradable COD that these organisms obtain and therefore their active mass in the system, which in turn determines to a large extent the amount of P removal that is achieved (see Section 4.14.31.1). The VFA is derived from that present in the influent (part of the RBSO) and is also generated in the anaerobic reactor by acid fermentation (Table 1). The rate of VFA uptake is so rapid that it can be assumed that all VFA in the influent will be taken up in the anaerobic reactor by the PAOs (Wentzel et al., 1985). The RBSO that is not in an VFA form is called fermentable RBSO (FBSO) and will be acid-fermented by the OHOs in the anaerobic reactor to VFA which then can be sequestered by the PAOs. The rate of this fermentation reaction is slower than the VFA uptake rate (Wentzel et al., 1985), and the amount of FSBO fermented to VFA depends on the influent FBSO concentration and system design. Thus, for accurate design of BEPR, the RBSO needs to be subdivided into two subfractions, VFA (Sbsai) and FBSO (Sbsfi). Heterotrophic (OHO) active biomass. In some wastewatercharacterization schemes, an active OHO concentration is included as part of the total organic material in the influent (shown dotted in Figure 3). This may be necessary for wastewaters collected in well-aerated sewers (unintentionally in steep sloping ones or intentionally to reduce biocorrosion) so that biological activity that would normally take place in the biological reactor can already take place to a significant extent in the sewer. In some wastewaters, as much at 20% of the total COD has been measured to be active OHO mass (Kappeler and Gujer, 1992). In these aerobic sewer systems, nitrification can also take place to a considerable extent (detectable by
418
Biological Nutrient Removal
nitrate concentrations in the influent) so that the AS system is seeded not only with a significant OHO mass, but also with nitrifiers (ANOs). This seeding effect does not influence the design of the AS system much, particularly when primary sedimentation is included, because most of the biomass settles out as primary sludge. However, should the AS system receive a considerable nitrifier (ANO) organism seed, this would manifest in the system as a higher than usual maximum specific growth rate of nitrifiers at 20 1C (mAm20). Usually, the seeding of OHOs and ANOs is negligible, unless some recycle stream at the plant returns waste activated sludge (WAS) into the influent flow or some other wastewatertreatment plant (WWTP) discharges WAS into the sewer. This has been observed by Mbewe et al. (1994), who developed a batch test procedure for measuring the influent OHO active mass (and the RBSO fraction); they measured that in normal domestic wastewater with a relative short trunk sewer (B5 km), the influent OHO concentration (as COD) was less than 6% of the total COD concentration, whereas at a plant where WAS was recycled to the influent, it was 15% (Wentzel et al., 1995, 1998, Lee et al., 2006). In the steady-state model, the seeding effect of OHOs is considered negligible and ignored because by far the greater proportion of these organisms develop in the biological reactor. In the wastewater-characterization schemes of the UCT (Dold et al., 1991) and IWA (Henze et al., 1987) kinetic simulation models, influent heterotrophic and nitrifier organism fractions are recognized, but may or may not be included in model implementation.
4.14.4.2.2 Quantification of COD fractions Quantification of the COD fractions is based principally on monitoring the response of AS to the different COD fractions (Ekama et al., 1986; Dold et al., 1991; Henze et al., 1994). Measurement of influent unbiodegradable soluble COD (Susi) is relatively simple: for sludge ages greater than 3 days, all the RBSO is utilized in the biological reactor and the BPO and unbiodegradable particulate COD (Supi) are enmeshed in the mixed liquor and will settle out in the secondary settling tank. Also, negligible soluble COD is generated during the biological transformation of the biodegradable COD. Thus, the only soluble COD in the effluent is the USO (Sus) of the influent. By running a steady-state laboratory-scale unit at a long sludge age (410) days and measuring the filtered (0.45 mm) effluent COD, Suse can be determined. At steady state, Susi ¼ Suse. Estimates of Susi can be obtained also by flocculating and filtering samples taken at the end of 24 h aerobic batch tests and measuring filtrate (o0.45 mm) COD (Mbewe et al., 1994; Wentzel et al., 1998). Measurement of UPO (Supi) and influent biodegradable organics COD (BO, Sbi) presents more difficulties because both contribute to the mixed liquor VSS in the biological reactor, UPO (Supi) directly by generating inert VSS mass (XI) and BO (Sbi) indirectly by generating OHO active biomass (XBH) and endogenous mass (XEH). UPO (Supi) and BO (Sbi) can be estimated simultaneously by running a laboratory-scale unit at a long sludge age (415 days), and comparing the measured mixed liquor concentration (conventionally as VSSs, but can also be measured with the COD test) and carbonaceous oxygen utilization with those calculated from the steady-state model with different estimated Supi values (see
Section 4.14.9.3). Provided a 100% COD balance over the experimental system is obtained, the value for Supi that gives a theoretical mixed liquor concentration (VSS or COD) and carbonaceous oxygen utilization equal to measured values will be the Supi concentration for the specific wastewater (Ekama et al., 1986). If the COD mass balance is not 100%, the mixed liquor concentration comparison will provide the more accurate estimate for Supi. A long sludge age is selected because the mixed liquor VSS concentration becomes more sensitive to Supi as the sludge age increases. Having found Supi and Susi, Sbi then can be found by difference. Overall, the Supi, Susi and Sbi must provide consistency between theoretical and measured responses at different sludge ages. Various methods have been developed to quantify RBSO, based on either (1) the different rates of utilization of RBSO and BPO – a bioassay test (short sludge age cyclic loaded systems, batch tests) or (2) the hypothesized difference in molecule size between RBSO and BPO – a physical separation (filtration, flocculation/filtration) (see Ekama et al., 1986; Dold et al., 1986; Henze, 1992; Mamais et al., 1993; Mbewe et al., 1994; Torrijos et al., 1994; Henze et al., 1994, Wentzel et al., 1995; Still et al., 1996; Ekama et al., 1996a; Wentzel et al., 2000). It must be remembered that the physical separation techniques are an approximation only of the biokinetic division of the biodegradable COD. The degree of success with the physical separation techniques depends on the method used: these include filtration through various pore-size paper, glass fiber, or membranes, with or without preflocculation. With preflocculation, all the particulate (settleable and nonsettleable) material is flocculated to settleable or filterable material (mimicking the bioflocculation process in the bioreactor), leaving the filtrate with only the soluble constituents. Without preflocculation, the various filter media, from molecular mass cutoff membranes through to filter paper with relatively large pore sizes (e.g., 1 mm), retain decreasing proportions of the particulate (suspended) material on the filter medium, and produce filtrates with increasing proportions of particulate (suspended) materials. Without preflocculation, even the filtrate that passes through a 0.45-mm filter membrane should not be regarded as entirely soluble material (Dold et al., 1986); it approximates the soluble constituents, but some particulate material does pass through the 0.45 mm membrane and is included with the soluble constituents in measurements of filtrate COD. For this reason, it is imperative that when wastewater is physically characterized by means of various separation tests, the type of separation method used is specified; without this no estimate can be made of the proportion of particulate material included with the soluble material so that no reliance can be attached to the results. Also, irrespective of the physical separation technique used, the filtrate will include both biodegradable and unbiodegradable material; an independent estimation of the unbiodegradable material is required to determine the biodegradable fraction of the soluble COD, as discussed above. In the conventional tests to quantify RBSO outlined above, both VFA and FBSO will be measured as RBSO (Figure 3). Accordingly, where VFAs in the wastewater are appreciable (e.g., in systems with acid fermentation of the primary sludge) and BEPR is to be included in the AS system, an additional test will be required to differentiate the two RBSO fractions –
Biological Nutrient Removal
direct measurement of the VFA (by gas chromatography or acid titration (Moosbrugger et al., 1992)) is the most practical.
419
VSS are related via the COD to VSS ratio (fcv):
XIi ¼ Supi =f cv ¼ f S0 up Sti =f cv
ðmgVSS l1 Þ
ð6Þ
4.14.4.2.3 Analytical formulation for COD For analysis and use in the steady-state model, accepting the influent OHO concentration is zero, the scheme indicated in Figure 3 can be expressed mathematically as follows: Biodegradable and unbiodegradable COD fractions:
Sti ¼ Sui þ Sbi
ð1Þ
where Sti is the total influent COD concentration (mgCOD l1), Sui the unbiodegradable influent COD concentration (mgCOD l1), and Sbi the biodegradable influent COD concentration (mgCOD l1). Each of the two fractions on the right-hand side of Equation (1) is again subdivided. Unbiodegradable COD fractions. The unbiodegradable COD concentration consists of two components, soluble and particulate, that is,
where, XIi is the UPO concentration in the influent expressed as VSS (mgVSS l1) and, fcv the COD to VSS ratio of the UPOs, (1.48 mgCOD/mgVSS). It should be noted that this particulate organic material cannot be directly measured as VSS in the influent. The VSS in the influent consists of both biodegradable and UPOs. This combined particulate organic VSS material can only be separated into its unbiodegradable and biodegradable constituent components by means of biodegradability tests, such as those described by Ekama et al. (1986). The justification for using a uniform COD/VSS ratio for all three components of the reactor VSS is given in Section 4.14.4.3.2. Biodegradable COD fractions. The biodegradable COD concentration is found from Equation (1) as follows:
Sbi ¼ Sti Sui Sui ¼ Susi þ Supi
ðmgCOD l1 Þ
ð7aÞ
ð2Þ and from Equation (5)
where, Susi is the unbiodegradable soluble influent COD concentration (mgCOD l1) and, Supi the unbiodegradable particulate influent COD concentration (mgCOD l1). It is convenient to express Susi and Supi in terms of the total COD concentration Sti, that is, 1
Susi ¼ f S0 us Sti
ðmgCOD l Þ
ð3Þ
Supi ¼ f S0 up Sti
ðmgCOD l1 Þ
ð4Þ
Sbi ¼ Sti Sti ðf S0 up þ f S0 us Þ ðmgCOD l1 Þ ¼ Sti ð1 f S0 up f S0 us Þ ðmgCOD l1 Þ
From Figure 3 the biodegradable COD (Sbi) is divided into readily biodegradable soluble COD (Sbsi) and slowly biodegradable particulate COD (Sbpi). Each can be expressed in terms of Sbi as follows:
Sbsi ¼ f Sb0 s Sbi where fS’us is the fraction of total COD which is unbiodegradable soluble, (mgCOD/mgCOD) and fS’up the fraction of total COD which is unbiodegradable particulate (mgCOD/ mgCOD). Hence, from Equation (2)
Sui ¼ ðf S0 us þ f S0 up ÞSti
ðmgCOD l1 Þ
ð5Þ
The terminology of the symbols defining the different wastewater fractions is as follows: fS’up ¼ fraction (f) of the substrate COD (S) which is (denoted by the prime) unbiodegradable (subscript u) and particulate (subscript p). Because the prime comes immediately after the subscript capital S for COD, total COD is implied. Hence, fSb’s is not the same as fS’bs. The former is the fraction (f) of the biodegradable COD (subscript Sb) which is (prime) soluble (s), whereas the latter is the fraction (f) of the total COD (subscript S) which is (prime) biodegradable and soluble (subscript bs). This difference can be noted in Equation (9). Similarly fN’a is the fraction (f) of the total TKN (subscript N) which is (prime) ammonia (sub a) and fSbp’N is the fraction (f) of the biodegradable particulate COD (subscript Sbp) which is (prime) nitrogen (subscript N). Since by convention the mixed liquor solids concentration in the biological reactor is expressed in terms of VSS units rather than COD units, it is convenient to express the UPO material in terms of its equivalent influent volatile solids concentration (XIi). This is readily accomplished by noting that the COD and
ð7bÞ
ðmgCOD l1 Þ
ð8aÞ
and
Sbpi ¼ ð1 f Sb0 s ÞSbi
ðmgCOD l1 Þ
ð8bÞ
where fSb’s is the fraction of influent biodegradable COD which is readily biodegradable (mgCOD/mgCOD). The readily biodegradable COD can also be expressed in terms of the total COD (Sti); thus, substituting for Sbi from Equation (7) in Equation (8) yields
Sbsi ¼ f Sb0 s ð1 f S0 up f S0 us ÞSti ¼ f S0 bs Sti
mgCOD l1
ð9aÞ ð9bÞ
where fS’bs is the fraction of total COD that is readily biodegradable (mgCOD/mgCOD). If a preflocculated o0.45 mm filtrate is accepted as soluble (which is not unrealistic, Mamais et al., 1993 and Mbewe et al., 1994), then 1. the filtered influent COD concentration gives the sum of the two soluble COD fractions, that is,
Influent : Filt COD ¼ Susi þ Sbsi
ðmgCOD l1 Þ
ð10Þ
2. the difference between the influent unfiltered and filtered COD concentrations gives the two particulate COD
420
Biological Nutrient Removal Table 2 Summary of measurement and calculation procedure for influent organic COD concentrations from experimental results
fractions, that is,
Influent: Total COD filt COD ¼ Sbpi þ Supi
ðmgCOD l1 Þ
COD
3. the filtered effluent COD concentration gives the unbiodegradable soluble COD concentration, that is,
Effluent: Filt COD ¼ Susi
ðmgCOD l1 Þ
ð12Þ
The RBSO concentration (Sbsi) therefore is closely given by the difference between the filtered influent and effluent COD concentration (Equation (10) minus Equation (12), Wentzel et al., 2000) and the BPO (Sbpi) is given by the difference between the total particulate COD concentration (Equation (11)) and the unbiodegradable particulate COD concentration (Supi) known from the experimentally measured (or assumed) fS’up value. For BEPR systems, the readily biodegradable COD (Sbsi) is subdivided in fermentable readily biodegradable COD (Sbsfi) and VFA (Sbsai), that is,
Sbsi ¼ Sbsfi þ Sbsai
ð13Þ
Each of these can be expressed in terms of Sbsi:
Sbsai ¼ f Sbs0 a Sbsi
ð14Þ
Sbsfi ¼ f Sbs0 f Sbsi
ð15Þ
where fSbs’a is the fraction of readily biodegradable COD which is VFA (mgCOD/mgCOD) and fSbs’f the fraction of readily biodegradable COD which is fermentable (mgCOD/ mgCOD). The VFA can be expressed also in terms of the total COD, Sti, that is,
Sbsai ¼ f S0 bsa Sti
Sym
Source/method Unfiltered influent Filtered effluent (i) Flocculation filtrationa Filt infl–Filt effl (2) (ii) Directly by bioassay From steady-state systems (1) (2) (3) (4) Directly by GC or by acid titration (3) (6) Batch testsb
ð11Þ
ð16Þ
where fS’bsa is the fraction of total COD which is readily biodegradable VFA (mgCOD/mgCOD). Calculation of the different COD concentrations from experimental results is summarized in Table 2. The effect of these COD fractions on the AS system is discussed in Section 4.14.7.1.
4.14.4.3 Nitrogenous Materials As for the COD (or electron-donating capacity) of the organic material, the nitrogen content of the organic also is subdivided into different fractions. As noted earlier, fractionation of the N material is necessary only if nitrification, or ND, are included in the plant. The subdivision of the reduced N is shown in Figure 4 and is based on the same principles as those applied to COD. Assessment of these fractions is with the TKN and FSA tests. The TKN test measures both the FSA and the nitrogen bound in organic compounds (i.e., the organic N). In certain wastewaters, nitrate and nitrite (oxidized N) may be present but the TKN test does not include these (only the reduced N). Most municipal wastewaters will not contain nitrate or nitrite because in most sewerage systems the wastewater will be in a
1 2 3
Total USO RBSO
Sti Susi Sbsi
4 5 6 7 8
UPO BPO VFA FBSO ActiveOHO
Supi Sbpi Sbsai Sbsfi ZBHi
a
Flocculation prior to filtration provides more accurate results (Mamais et al., 1993; Mbewe et al., 1994). b ZBHi is the COD of the active OHO concentration in VSS terms (XBHi). The batch test method is described by Wentzel et al. (1995) and Ubisi et al. (1997a, b).
deoxygenated state and any nitrate entering the system is likely to be denitrified before it reaches the WWTP.
4.14.4.3.1 Nitrogenous material fractions The first subdivision of the influent TKN (Nti) is into FSA (Nai) and organically bound N (OrgN, Noi), see Figure 4. The FSA is immediately available for incorporation into the bacterial mass or for nitrification to nitrite or nitrate, if the environmental conditions in the system are appropriate for this. However, the organic N has to be converted to ammonia (FSA) by the action of organisms in the bioreactor (a process called ammonification), before it becomes available for incorporation into the bacterial mass, or for nitrification to nitrite or nitrate. In describing and modeling the ammonification of organic N to ammonia, it is accepted in the steadystate model and ASM2 (not ASM1) that this is linked to the organic material (COD) biological degradation (Henze et al., 1995, 2008). Therefore, each of the organic fractions has associated an organic N content with it. When the biodegradable COD is hydrolyzed/utilized for cell synthesis, the associated organic N is released to the bulk liquid as ammonia, which together with its influent ammonia counterpart participates in further biologically mediated reactions. The unbiodegradable COD fractions are not biologically broken down in the bioreactor and so the associated organic N also will not be affected and remains part of the organic material. Accordingly, the organic N is subdivided into subfractions in exactly the same fashion as the organic material COD is subdivided. Unbiodegradable subfractions. As with the COD, the influent organic unbiodegradable N (Noui) is subdivided into organic unbiodegradable soluble (Nousi) and organic unbiodegradable particulate (Noupi) subfractions (see Figure 4). By implication, these fractions are associated with the unbiodegradable COD fractions Susi and Supi, respectively (Figure 3). Thus, both these organic N fractions are unaffected by biological activity. The Nousi, which is associated with Susi, will pass through the system and be discharged with the effluent. The Noupi, which is associated with Supi and hence also with the unbiodegradable particulate VSS originating from the influent (XIi), is
Biological Nutrient Removal
421
TKN (Nti)
Organic N (Noi)
Ammonia N (Nai)
Heterotroph active biomass N
Biodegradable (Nobi)
Biodegradable particulate (Nobpi)
Unbiodegradable (Noui)
Biodegradable soluble (Nobsi)
Unbiodegradable particulate (Noupi)
Unbiodegradable soluble (Nousi)
Figure 4 Subdivision of the influent organic and inorganic material of N as measured by the TKN test. The organic N component is subdivided in the same way as the organic material as measured by the COD test (see Figure 3).
enmeshed in the sludge, settles out in the secondary settling tank, and is retained in the system. This fraction (denoted fSup’N) basically represents the N content of the UPO material (e.g., paper, hair, and other fibrous material) in the wastewater and exits the system via the waste sludge stream (Figure 2). Just as the COD/VSS ratio of the AS mixed liquor (fcv) is accepted to be the same for the active OHO (XBH), the endogenous residue (XEH) and the inert UPOs that accumulate in the reactor from the influent (XI), so also the TKN/VSS ratio (fN mgN/mgVSS) of these three mixed liquor constituents. The justification for this is the same as for the COD/VSS ratio (fcv) (see Section 4.14.4.3.2). Biodegradable subfractions. The biodegradable organic N (Nobi) is associated with the biodegradable COD (Sbi). Accordingly, Nobi can be subdivided into two subfractions (Figure 4), organic biodegradable soluble (Nobsi) and organic biodegradable particulate (Nobpi), associated with Sbsi and Sbpi, respectively (Figure 3). When the biodegradable COD fractions are utilized for organism metabolism and synthesis, the associated organic N fractions are broken down (ammonified) to FSA which, with its influent counterpart, participates in further biologically mediated reactions. As the Sbi is virtually completely utilized for all sludge ages greater than 3 days for fully aerobic systems and 6–8 days for unaerated–aerated N and P removal systems, it can be assumed that the associated (biodegradable) organic N (Nobi) is virtually completely ammonified to FSA. Consequently, in the steady-state model, subdivision of Nobi into soluble and particulate subfractions is not required for fully aerobic systems or for unaerated–aerated systems for biological N and P removal. Provided the sludge age is long enough, all the Sbi will be utilized in these systems
and so all the associated Nobi will become available as FSA. For the latter systems the subdivision of the biodegradable COD into RBSO (soluble) and BPO (particulate) is not required for quantifying the removal of the biodegradable COD (and organic N) itself, but rather for quantifying the magnitudes of denitrification of nitrate and biological P removal. At short sludge ages and very large unaerated sludge mass fractions, utilization of the biodegradable COD may not be complete. In this case, in terms of the steady-state design models and their use in wastewater characterization presented here, higher unbiodegradable particulate and soluble COD fractions (fS’up and fS’us) and concentrations (Supi and Susi) will be noted, and hence also proportionally higher unbiodegradable particulate and soluble organic N concentrations (Noupi and Nousi) compared with systems in which all the biodegradable COD is utilized.
4.14.4.3.2 Quantification of N fractions (Nti) With regard to measurement of the nitrogen fractions, the TKN (Nti) and FSA (Nai) are measured directly by the tests bearing these names; the organic nitrogen (Noi) is found from the difference between the TKN and FSA test concentrations. Difficulties in quantification arise when subdividing the Noi into unbiodegradable (Noui) and biodegradable (Nobi) subfractions. Numerous comparisons of the observed responses of laboratory-scale systems with those predicted by the steadystate model indicate that subdivision of the organic N into subfractions and quantification of these subfractions are important because they determine the effluent organic N concentration, the N content of the WAS, and the amount of
422
Biological Nutrient Removal
organic N ammonified to ammonia which is then available for incorporation into cell mass or for nitrification to nitrite and nitrate. However, the magnitudes of the two unbiodegradable organic N fractions (Nousi and Noupi) are relatively small compared to the influent TKN (Nti). Also, as mentioned above, the subdivision of the biodegradable organic N (Nobi) into soluble and particulate subfractions (Nobsi and Nobpi, respectively) is of little consequence in the steady-state model because it is accepted that all the biodegradable COD (Sbi) is utilized, with all the associated Nobi becoming available as ammonia. The organic unbiodegradable soluble N (Nousi) can be estimated by following the same procedures used to measure Susi: As all the biodegradable COD (Sbi) is utilized for systems with long sludge ages, all the organic biodegradable N (Nobi) associated with Sbi must have been broken down to FSA (Na). Accordingly, all the organic N present in the filtered (o0.45 mm) effluent from such a system must be due to Nous. Thus, by running a steady-state unit at a sludge age greater than 3 days and measuring the filtered (0.45 mm; effluent samples do not need to be preflocculated because of bioflocculation by the mixed liquor) effluent organic N (effluent TKN-FSA), Nouse is determined; at steady state, Nousi ¼ Nouse because, as for the Sus, it is accepted that no significant unbiodegradable soluble organic N (Nous) is generated by the biological processes taking place in the reactor. From experimental results, Nousi is very low and usually is of little consequence, except where very low effluent N standards based on total N (reduced þ oxidized, TN ¼ FSA þ OrgN þ NOx) are set (Pagilla et al., 2009). The UPO N concentration in the influent (Noupi) is determined from its accepted association with UPO (Supi). The UPO (Supi) accumulates in the reactor as organic VSS solids (XI). Therefore, the N associated with Supi (i.e., Noupi), will also accumulate in the reactor with its associated XI. Because the biodegradable and unbiodegradable particulate COD fractions cannot be individually tested for in the influent wastewater, it follows that this also applies to the biodegradable and UPO N fractions. This can only be done by means of biodegradability tests, the same tests by means of which estimates of the biodegradable and unbiodegradable particulate COD fractions are obtained, by monitoring the nitrogen content of the mixed liquor VSS (or COD) that accumulates in the reactor. From experimental data on the TKN concentration of the mixed liquor in the biological reactor, it was found that the TKN/VSS concentration ratio (fN) remained approximately constant irrespective of sludge age from 3 to 30 days, despite the fact that the proportions of the three constituent fractions of the mixed liquor (i.e., active OHOs (XBH), endogenous residue (XEH) and unbiodegradable particulate (XI)) change appreciably with sludge age. The fraction of XI (i.e., N) is therefore constant and equal to that of the other VSS constituent fractions, at about 0.1 mgN/mgVSS. With the TKN content of the UPOs in terms of VSS (fN) at say 0.10 mgN/ mgVSS, the TKN content of this same material in terms of COD (fSup’N) would be 0.10/1.48 ¼ 0.068 mgN/mgCOD (i.e., fSup’N ¼ fN/fcv or Noupi ¼ fN XIi ¼ fN Supi/fcv ¼ fSup’N Supi mgN l1). Having determined Nousi and Noupi, Nobi can be found by difference (Nobi ¼ Noi Nousi Noupi). For the steady-state
design model, as noted above, it is not required to subdivide Nobi into soluble and particulate subfractions (Nobsi and Nobpi, respectively). However, for the dynamic simulation models, this subdivision is required and can be done with the aid of TKN and FSA test results of filtered and unfiltered influent samples once the two unbiodegradable concentrations are known from experimental systems, analogous to quantifying the COD subdivisions in Section 4.14.4.2.3.
4.14.4.3.3 Analytical formulation For use in the steady-state model, the relationships indicated in Figure 4 can be expressed as follows: The influent total TKN (Nti) is divided into FSA (Nai) and organic N (Noi):
Nti ¼ Nai þ Noi
ðmgN l1 Þ
ð17aÞ
The Nai can be expressed in terms of Nti:
Nai ¼ f N0 a Nti
ðmgN l1 Þ
ð17bÞ
where fN’a is the fraction of influent TKN which is ammonia (mgN/mgN). The Noi can be subdivided in exactly the same way as the COD, that is, soluble, particulate, biodegradable, and unbiodegradable. The first subdivision is into biodegradable (Nobi) and unbiodegradable (Noui) subfractions:
Noi ¼ Nobi þ Noui
ðmgN l1 Þ
ð18Þ
The Noui can be further subdivided into soluble (Nousi) and particulate (Noupi) subfractions:
Noui ¼ Nousi þ Noupi
ðmgN l1 Þ
ð19Þ
It is convenient to express Nousi in terms of Nti:
Nousi ¼ f N0 ous Nti
ðmgN l1 Þ
ð20aÞ
where fN’ous is the fraction of influent TKN which is organic unbiodegradable soluble N (mgN/mgN). Alternatively, Nousi can be expressed in terms of its associated COD fraction, Susi:
Nousi ¼ f Sus0 N Susi
ðmgN l1 Þ
ð20bÞ
where fSus’N is the fraction of unbiodegradable soluble COD which is nitrogen (mgN/mgCOD). The Noupi can be expressed in terms of the influent unbiodegradable particulate COD (Supi), or in terms of its volatile solids counterpart (XIi):
Noupi ¼ f Sup0 N Supi ¼ f Sup0 N XIi f cv
ðmgN l1 Þ ð21Þ
where fSup’N is the fraction of the influent unbiodegradable particulate COD which is nitrogen (E0.068 mgN/mgCOD). From Equations (17) to (21) the organic biodegradable N (Nobi) can be found by subtracting Nai, Nousi, and Noupi
Biological Nutrient Removal
423
Table 3 Summary of measurement and calculation procedure for influent TKN concentrations from experimental results
from Nti:
Nobi ¼ Nti ð1 f N0 a f N0 ous Þ f Sup0 N XIi f cv ¼ Nti ð1 f
N0 a
Þf
Sus0 N
Susi f
Sup0 N
Supi
ð22aÞ ð22bÞ
For the purposes of steady-state design, it is not necessary to subdivide Nobi into soluble (Nobsi) and particulate (Nobpi) subfractions (Section 4.14.21.1), but if required (as included in the dynamic models of Dold et al. (1991)) this can be done as follows:
Nobi ¼ Nobsi þ Nobpi
ðmgN l1 Þ
As done above for the unbiodegradable N subfractions, each of the biodegradable N subfractions can be expressed in terms of either the influent TKN or its associated influent COD subfraction:
ðmgN l1 Þ
Nobsi ¼ f N0 obs Nti ¼ f Sbs0 N Sbsi
ðmgN l1 Þ
ð23aÞ ð23bÞ
where fN’obs is the fraction of the influent TKN which is organic biodegradable soluble (mgN/mgN), fSbs’N the fraction of the influent readily biodegradable COD which is nitrogen, (mgN/mgCOD) and
Nobpi ¼ f N0 obp Nti ¼ f Sbp0 N Sbpi
ðmgN l1 Þ
ðmgN l1 Þ
ð26Þ
3. the difference between the effluent filtered TKN and FSA concentrations gives the unbiodegradable soluble organic N (Nousi):
ðmgN l1 Þ
Unfiltered influent Filtered influent (1) (2) Filtered effluent Filtered effluent (5) (4) From steady-state systems and Supi Floc filt inf TKNa (2) (4) (1) filt inf TKN (5) Batch testsb
a
Preflocculated. From the N content (fN ¼ 0.068 mgN/mgCOD) of the respective COD concentrations.
b
From the above, the soluble biodegradable organic N (Nobsi) is found by difference from Equations (26) and (27). The UPO N (Noupi) can be calculated if the unbiodegradable particulate COD concentration Supi (or fraction, fS’up) is known from Equation (21). The BPO N (Nobpi) can therefore be calculated by difference from Equation (25). Calculation of the different TKN concentrations from experimental results is summarized in Table 3.
Experimental observations on a number of different municipal wastewater flows have indicated that the maximum specific growth rate of the nitrifiers at 20 1C (mA20) can vary greatly in value and appears to be specific to each waste flow, that is, it appears to be a wastewater characteristic. In particular, the value for mA20 appears to be very sensitive to the industrial component of the wastewater presumably because various industries discharge metals, elements, and organics which act in an inhibitory fashion on the ANOs. The magnitude of mA20 can have a significant influence of the design of nutrient removal systems, particularly on the selection of sludge age and unaerated mass fraction. Therefore, for optimal design it is most desirable to have an estimate of mA20. This estimate can be obtained by means of batch tests (Still et al., 1996; Dold et al., 1991) or by running a completely mixed single reactor at a sludge age of about 8 days, imposing a sequence of aerobic and anoxic periods and measuring the rate of nitrate increase during the aerobic period (van Haandel et al., 1981; WRC, 1984). The data thus obtained can be used to calculate mA. By applying the well-established relationship between mA and temperature (see Section 4.14.20.2), the value at a particular temperature within the normal wastewater range (8–281C) can be determined.
2. the difference between the filtered influent TKN and FSA concentrations gives the two soluble organic N fractions, that is,
Nousi ¼ Nte Nae
Nti Nai Noi Nae Nte Nousi Noupi Nobsi Nobpi NXBHi
ð24bÞ
ðmgN l1 Þ ð25Þ
ðmgN l1 Þ
Total FSA OrgN Effluent FSA Effluent TKN USOrgN UPOrgN BSOrgN BPOrgN ActiveOHO
Source/method
4.14.4.3.4 Maximum specific growth rate of nitrifiers at 20 1C
1. the difference between the unfiltered and filtered influent TKN concentrations gives the two particulate organic N fractions, that is,
Filt TKN FSA ¼ Nousi þ Nobsi
1 2 3 4 5 6 7 8 9 10
Sym
ð24aÞ
where fN’obp is the fraction of the influent TKN which is organic biodegradable particulate (mgN/mgN) and, fSbp’N the fraction of the influent slowly biodegradable COD which is nitrogen (mgN/mgCOD). In terms of the TKN characterization block diagram (Figure 4) and accepting that the influent active OHO concentration is zero and that the preflocculated o0.45 mm filtrate is soluble material, then:
Unfilt TKN filt TKN ¼ Noupi þ Nobpi
TKN
ð27Þ
where Nte is the Filtered effluent TKN concentration and, Nae the effluent FSA concentration.
4.14.4.3.5 Typical wastewater TKN characteristics Full waterborne sanitation wastewater N characteristics vary widely due to the varying degree of hydrolysis of the urea in urine (organic N) to ammonia but about 75% of the TKN is FSA, and 25% organic N, which as percentages of the TKN comprise 1–3% unbiodegradable soluble (Nousi), 8–10%
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Biological Nutrient Removal
unbiodegradable particulate (Noupi) and 10–14% biodegradable (Nobi).
4.14.4.4 Phosphorous materials As for the carbonaceous and nitrogenous materials, the phosphorus (P) material is also subdivided into fractions (Figure 5). This fractionation is only necessary if P removal (biological or chemical) is required in the system. The fractionation of the influent phosphorus follows the same scheme used for characterization of the influent nitrogen. Assessment of the influent phosphorus fractions is with the TP and OP tests. The TP test measures soluble orthophosphate, condensed orthophosphates (pyro, meta, and other polyPs), and the phosphorus bound in organic compounds and is denoted Pti. The OP (also orthoP) test measures principally the orthophosphates but a small fraction of some condensed phosphates also may be included. In this chapter, all P concentrations measured by the OP test are termed ‘soluble orthophosphate’ (Psi) and the difference in P concentration between the TP and OP tests is called ‘organic P’. As for the nitrogenous material, the Pti is divided into the same five fractions.
4.14.4.4.1 Phosphorus fractions The first subdivision of total influent phosphorus (Pti) is into soluble orthophosphate (Psi) and organically bound P (Poi) (Figure 5). In both raw and settled municipal wastewaters, the soluble orthophosphate fraction predominates, ranging between 70% and 90% of the TP. The main source of orthophosphates are detergents which can contribute up to
50% of the total phosphate load (Wiechers and Heynike, 1986). Thus, in countries where phosphate-free detergents are used, the relative contribution of the orthophosphates will be lower, so also the influent TP concentration in relation to the COD and TKN (Gleisberg, 1993). The soluble orthoP (Psi) is immediately available for incorporation into bacterial mass and, if the system is appropriately designed, for BEPR. The organic phosphorus (Poi) needs to be converted to orthophosphate (Psi) by the action of organisms in the bioreactor (if possible) before it becomes available. Because of this source of orthophosphate in the biological reactor, it is important in design of BEPR systems to measure the influent phosphorus with the TP test to take account of the organic phosphorus in the influent; if only soluble orthoP (Psi) is measured, the BEPR required by the system to achieve a specified effluent orthoP concentration will be underestimated. For description and modeling the conversion of organic P to orthoP, it is accepted that the conversion is linked to the COD biodegradadation, as was done for the organic N. Therefore, each of the COD fractions has associated with it an organic P content. When the biodegradable COD is utilized for cell synthesis, the associated organic P is released as orthoP. When the unbiodegradable particulate COD is enmeshed in the sludge mass, the associated organic P is similarly enmeshed; when the unbiodegradable soluble COD flows through the system and appears in the effluent, the associated organic P will do likewise. Accordingly, subdivision of the organic P follows the subdivision of the COD and the organic N.
Total P (Pti)
Ortho P (Psi)
Heterotroph active biomass P
Organic P (Poi)
Biodegradable (Pobi)
Biodegradable particulate (Pobpi)
Biodegradable soluble (Pobsi)
Unbiodegradable (Poui)
Unbiodegradable particulate (Poupi)
Unbiodegradable soluble (Pousi)
Figure 5 Subdivision of the influent organic and inorganic material in terms of P as measured by the total P test. The organic P component is subdivided in the same way as the organic material as measured by the COD test (see Figure 3).
Biological Nutrient Removal
Unbiodegradable subfractions. As with the COD, the organic unbiodegradable P (Poui) is subdivided into organic unbiodegradable soluble (Pousi) and organic unbiodegradable particulate (Poupi) subfractions (Figure 5). By implication, these subfractions are associated with the unbiodegradable organics, USO (Susi) and UPO (Supi), respectively. Thus, both these organic P fractions are unaffected by biological activity. The Pousi, associated with Susi, will pass through the system to be discharged in the effluent. The Poupi, associated with Supi, will be enmeshed in the sludge, settle out in the secondary settling tank, and be retained in the system with the associated inert particulate organics (XI); hence, this P fraction leaves the system via the waste sludge stream (Figure 2). Biodegradable subfractions. The organic biodegradable P (Pobi) is associated with the biodegradable COD (Sbi). Accordingly, Pobi can be subdivided into two subfractions (see Figure 5), organic biodegradable soluble (Pobsi) and organic biodegradable particulate (Pobpi) associated with Sbsi and Sbpi, respectively. When the biodegradable COD fractions are utilized for metabolism and synthesis, the associated organic P fractions are broken down to soluble orthoP which, with its influent counterpart, is available for use in biologically mediated reactions. As Sbi is virtually completely utilized for all sludge ages greater than 3 days (see Section 4.14.4.2.1), it can be assumed that the associated biodegradable organic P (Pobi) is virtually completely broken down to soluble orthoP. Consequently, for design, subdivision of the Pobi into subfractions is not required; it can be assumed that all the influent organic biodegradable P (Pobi) will become available as soluble orthoP (Ps) in the bioreactor.
4.14.4.4.2 Quantification of P fractions To measure the phosphorus (P) fractions, the influent soluble orthoP (Psi) is measured directly by a colorimetric test of that name; the influent total P (Pti) can be measured by first subjecting the sample to an acid digestion step in which the organic P is oxidized to orthoP followed by the colorimetric test for orthoP. The organic P (Poi) is found by the difference between Pti and Psi. Subdivision of Poi into subfractions is important because one of them, the unbiodegradable soluble organic P concentration (Pousi), determines the minimum effluent TP concentration that can be achieved. Even if in the BEPR system, all the biodegradable organic P (Pobi) is converted to OP and all the OP is removed, the effluent TP concentration contains Pousi. In this respect, the setting of the effluent P standard can have a profound influence on the design of the plant and whether or not to include effluent filtration. If the standard is specified in terms of TP, then Pousi and the P content of any suspended solids (which can be high in BEPR plants) escaping with the effluent are included in the effluent TP. With 10% P/VSS content, losing an average of about 10 mgSS l1 with the effluent and with a Pousi of 0.3 mgP l1, the effluent P concentration already exceeds 1 mgP l1 TP even if all the OP has been successfully removed. In South Africa, the effluent P standard (where it needs to be met) is set at 1 mgP l1 dissolved OP which excludes Pousi and the P in the effluent suspended solids. For the effluent, the unbiodegradable particulate P concentration (Poupi) is not important because this fraction accumulates with the UPOs in the reactor
425
Table 4 Summary of measurement and calculation procedure for influent TP concentrations from experimental results TP 1 2 3 4 5 6 7 8 9 10
Total P OrthoP OrgP Effluent OP Effluent TP USOrgP UPOrgP BSOrgP BPOrgP ActiveOHO
Symbol
Source/method
Pti Psi Poi Pse Pte Pousi Poupi Pobsi Pobpi PXBHi
Unfiltered influent Filtered influent (1) (2) Filtered effluent Filtered effluent (5) (4) From steady-state systems and Supia Filt inf TPa (2) (4) (1) filt inf TP (5) Batch testsb
a
In contrast to the COD and N, sample NOT preflocculated. From the P content (fP ¼ 0.02 mgP/mgCOD) of the respective COD concentrations.
b
biomass. The magnitudes of the two unbiodegradable organic P concentrations (Pousi and Poupi) are very small compared with the total influent P (Pti), usually less than 0.5 and 3.0 mgP l1, respectively. The subdivision of the organic biodegradable P (Pobi) into soluble and particulate concentrations (Pobsi and Pobi respectively) is of no consequence in the steady-state model – because it is accepted that all the biodegradable COD (Sbi) is utilized, all the associated Pobi becomes available as soluble orthoP (Ps), which either is removed (taken up by biomass) or exits with the effluent. The four organic P fractions (Figure 5) are, like the four organic N fractions (Figure 4), difficult and tedious to measure. If, on the same fully aerobic system in which the low N fractions are measured, unfiltered TP and filtered (o0.45 mm) TP and orthoP concentrations are also measured on the influent and effluent samples, then the four organic P fractions can be calculated (see Table 4). Note that the filtered sample cannot be preflocculated because this will cause flocculation of the soluble P fractions also; however, because the organic P is small relative to the TP, excluding the preflocculation step prior to 0.45 mm filtration is reasonable. As with Noupi, Poupi is determined from its accepted association with Supi. Expressed in terms of COD units Poupi ¼ 0.03 (mgP/mgVSS)XIi (mgVSS l1 influent) ¼ 0.03 (mgP/mgVSS)Supi (mgCOD l1 influent)/1.48 (mgCOD/mgVSS) ¼ 0.02 (mgP/mg unbiodegradable particulate COD) Supi (mgCOD ll influent) Equation (33). Having determined Pousi and Poupi, Pobi can be found by difference: Pobi ¼ Poi Pousi Poupi. In the steady-state model, as noted above, it is not required to subdivide Pobi into soluble and particulate subfractions (Pobsi and Pobpi, respectively), but this can be done if required in the same way as the equivalent COD or N fractions defined above, that is, with the aid of Total P and orthoP analysis of filtered and unfiltered influent and effluent samples from long sludge age fully aerobic AS systems.
4.14.4.4.3 Analytical formulation For use in the steady-state design procedures, the relationships indicated in Figure 5 can be expressed as follows: The influent total P (Pti) is divided into soluble orthoP (Psi) and organic P (Poi):
Pti ¼ Psi þ Poi
ðmgP l1 Þ
ð28Þ
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Biological Nutrient Removal
The Psi can be expressed in terms of Pti:
Psi ¼ f P0 s Pti
1
ðmgP l Þ
ð29Þ
where fP’s is the fraction of influent total P which is soluble orthoP (mgP/mgP). The Poi can be subdivided into biodegradable (Pobi) and unbiodegradable (Poui) subfractions:
Poi ¼ Pobi þ Poui
1
ðmgP l Þ
ð30Þ
The Poui can be further subdivided into soluble (Pousi) and particulate (Poupi) subfractions:
Poui ¼ Pousi þ Poupi
ðmgP l1 Þ
ð31Þ
ðmgP l1 Þ
ð32aÞ
where fP’ous is the fraction of influent total P which is organic unbiodegradable soluble (mgP/mgP). Even though theoretically it would be more logical to define Pousi as a fraction of the organic P (Poi) rather than the total P (Pti) because Pousi is essentially part of Poi, defining Pousi in terms of Pti obviates having to know the organic P concentration (Poi), which requires both the total P and orthoP concentration to be measured. Defining Pousi in terms of Pti requires only Pti to be measured, which is required in any event due to the conversion of organic P to orthoP in the reactor. The Pousi can also be related back to the organic material COD of which it forms part, being a fraction fSus’P of the unbiodegradable soluble organics:
Pousi ¼ f Sus0 P Susi
Pobsi ¼ f P0 obs Pti ¼ f Sbs0 P Sbsi
ðmgP l1 Þ
ðmgP l1 Þ ðmgP l1 Þ
Pobpi ¼ f P0 obp Pti
ðmgP l1 Þ ðmgP l1 Þ
Poupi ¼ f Sup0 P Supi ¼ f Sup0 P XIi f cv
ðmgP l Þ
ð36bÞ
Unfilt TP filt TP ¼ Poupi þ Pobpi
ðmgP l1 Þ
ð37Þ
2. the difference between the influent filtered TP and orthoP concentrations gives the two soluble P concentrations, that is,
ðmgP l1 Þ
ð38Þ
3. the difference between the filtered effluent TP and OP concentrations gives the unbiodegradable soluble organic P (Pousi), that is,
Filt Effl TP Effl OP ¼ Pousi
ðmgP l1 Þ
ð39Þ
ð33Þ
where fSup’P is the fraction of the influent unbiodegradable particulate COD which is phosphorus ( ¼ 0.020 mgP/ mgCOD). From Equations (28) to (33), the organic biodegradable P (Pobi) can be found by subtraction:
Pobi ¼ Pti ð1 f P0 s f P0 ous Þ f Sup0 P XIi f cv
ð34aÞ
¼ Pti ð1 f P0 s Þ f Sus0 P Susi f Sup0 P Supi
ð34bÞ
For the purposes of steady-state design (Section 4.14.21.1), it is not necessary to subdivide Pobi into soluble (Pobsi) and particulate (Pobpi) subfractions because all of it is converted to orthoP, but if required (as included in the dynamic simulation models for BEPR) this can be done as follows:
Pobi ¼ Pobsi þ Pobpi
ð36aÞ
1. the difference between the influent unfiltered and filtered TP concentrations gives the two particulate organic P concentrations, that is,
Filt TP OP ¼ Pousi þ Pobsi
1
ð35bÞ
where fP’obp is the fraction of the total P which is organic biodegradable particulate (mgP/mgP) and fSbp’P the fraction of slowly biodegradable (particulate) COD which is P (mgP/mgCOD). If o0.45 mm (or o0.10 mm, but not preflocculated) filtered samples are considered soluble, then, like for the nitrogenous materials,
ð32bÞ
where fSus’P is the fraction of unbiodegradable soluble COD which is P (mgP/mgCOD). Similarly, the Poupi can be expressed in terms of the influent unbiodegradable particulate COD (Supi), or in terms of its VSS counterpart (XIi):
ð35aÞ
where fP’obs is the fraction of the total P which is organic biodegradable soluble (mgP/mgP), fSbs’P the fraction of readily biodegradable (soluble) COD which is P (mgP/mgCOD), and
¼ f Sbp0 P Sbpi
It is convenient to express Pousi in terms of Pti:
Pousi ¼ f P0 ous Pti
As done above for the organic unbiodegradable N subfractions, each of the biodegradable P subfractions can be expressed in terms of either the influent TP or its associated influent COD subfraction:
From the above, the soluble biodegradable organic P (Pobsi) is found by difference from Equations (38) and (39). The UPO P (Poupi) can be calculated (from Equation (33)) if the unbiodegradable particulate COD concentration, Supi (or fraction fS’up) is known. The BPO P (Pobpi) can therefore be found by difference from Equation (37). Calculation of the different TP concentrations from experimental results is summarized in Table 4.
4.14.4.4.4 Typical wastewater phosphorus characteristics Full waterborne sanitation wastewater phosphorus characteristics also vary widely depending on the P content of detergents. Generally, for raw (unsettled) wastewater about 75% of TP is dissolved orthoP (Psi) and 25% organically bound P (Poi). Of the TP, this latter fraction comprises 0–2% unbiodegradable soluble organic P (Pousi), 10–15% UPO P (Poupi), and 8–12% biodegradable (hydrolyzable to orthoP) organic P (Pobi).
Biological Nutrient Removal 4.14.4.5 Inorganic Dissolved, Settleable, and Nonsettleable Solids The wastewater characterization described above focuses on the organic material. Inorganic material also influences the WWTP. Like the organic material, the inorganic material can be divided into particulate and dissolved constituents (Figure 1). The particulate constituents comprise settleable and nonsettleable material. The concentration of inorganic dissolved solids (IDSs, but conventionally called TDSs, due to wide application to waters with low dissolved organics) does not affect the WWTP except where some of the dissolved constituents are inhibitory, like heavy metals, or excessively high in the case of seawater waste collection systems. For most terrestrial water waste collection systems, the IDSs that enter the WWTP exit it via the effluent. A very low concentration is taken up biologically (5–20 mg l1) for growth, such as Ca, Mg, and K, in particular in BEPR systems. Also under certain circumstances, some dissolved inorganic solids may precipitate in the biological reactor, but this is generally very low as well, unless intentionally increased by chemical precipitation by, for example, simultaneous Fe or Al dosing for chemical P removal (Figure 1). Strictly speaking, FSA is inorganic. This is removed from the wastewater by incorporation into the sludge mass or transfer to gas phase, either as free ammonia via gas stripping or as dinitrogen gas via the biologically mediated processes of nitrification and denitrification. Generally, in activated sludge systems very little ammonia is removed by gas stripping due to the low pH (7–8) relative to that required for ammonia stripping (49.5). The particulate inorganic solids (more usually called suspended solids, Standard Methods (1985) and hence ISS) do affect the AS biological reactor solids concentration. These solids comprise grit, sand, silt, clay, and similar materials. Generally, these solids have a biofilm growth on their surfaces which significantly reduces their settling velocity, especially for the smaller particles. The large grit and sand particles, which enter the collection system via leaking joints and holes in the sewer pipes and groundwater ingress or even as cleaning material from some low-income communities, are usually removed in grit collection units as one of the first unit operations in primary treatment at WWTPs. These solids are highly abrasive and significantly reduce the life of mechanical equipment such as pumps and also can accumulate in settling tanks and biological reactors. The medium-sized influent inorganic particulate solids can be removed by primary sedimentation (and hence is the settleable fraction) and become part of the primary sludge. The concentration of organic and inorganic solids removed in primary sedimentation defines the volatile to total solids ratio (VSS/TSS) of the primary sludge. The inorganic solids not removed in primary sedimentation (nonsettleable particulates), which are the very small particles, enter the biological reactor and become enmeshed with the biological reactor mixed liquor (all settleable) and hence are retained in the system forming the inorganic component of the mixed liquor, usually measured as the ISS concentration in the reactor. If primary sedimentation is not included, then the influent ISS entering the biological reactor is much higher, with the result that the ISS concentration in the reactor is also much higher.
427
The ISSs in the biological reactor (which are all settleable) do not arise only from the influent ISS in the raw or settled wastewater. Also contributing to the measured reactor ISS is intracellular dissolved inorganic solids, which, when dried in the test procedure, precipitate as inorganic solids and hence are measured as ISS. This adds to the ISS from the influent. From an evaluation of experimental data collected over 10 years in the UCT Wastewater Laboratory with real and artificial wastewaters fed to N and N and P removal systems over a range of sludge ages from 8 to 20 days, Ekama and Wentzel (2004) calculated that OHOs contain about 0.15 mg ISS/mg OHO organic (volatile) suspended solids (VSSs). This 0.15 mg ISS/mg OHOVSS value was validated with data from the literature (van Haandel et al., 1998) by Ekama et al. (2006a). For the PAOs, the ISS residue was additionally 3.29 times their polyP content, with a maximum ISS of 1.3 mgISS/mgPAOVSS when their P content was 0.38 mgP/mgPAOVSS under aerobic P uptake BEPR conditions. The consequence of this is that the ISS concentration in the reactor depends on many factors such as (1) influent ISS concentration, (2) sludge age, and (3) hydraulic retention time of the biological reactor, (4) wastewater influent COD concentration and (5) characteristics (fS’up and fS’us), and (6) the magnitude of BEPR. Roughly for an N removal system treating raw wastewater at a long sludge age (15–20 days), the influent ISS contributes about two-thirds of the reactor ISS, the OHOs contributing the remaining onethird. For a N and P removal system treating settled wastewater at a short sludge age, the influent ISS contributes about onesixth of the reactor ISS, the OHOs one-sixth and the PAOs the remaining two-thirds. Because the influent ISS concentration in settled wastewater is much lower (Bone-third) than in raw wastewater, the VSS/TSS ratio of the mixed liquor in the reactor treating raw wastewater is lower (0.75–0.80) than that treating settled wastewater (0.83–0.87). Including BEPR reduces the mixed liquor VSS/TSS ratio significantly below these approximate values. Details on how to determine the influent ISS concentration are given by Ekama and Wentzel (2004) and to calculate the reactor ISS concentration and VSS/TSS ratio are given in Sections 4.14.9.2 and 4.14.31.6.
4.14.4.6 Other Materials Other physical and chemical parameters also influence the AS system and therefore also need to be measured. The main ones are temperature, H2 CO3 alkalinity, and pH. The former has a strong influence on the rates of biological activity: the lower the wastewater temperature, the slower the biological rates; in particular, ND is affected. The H2 CO3 alkalinity and pH also play an important role. Most of the biological reactions in AS proceed optimally around a neutral pH (7–8). Some of the biological reactions (e.g., ND) influence the pH by releasing or taking up hydrogen ions (Hþ), with the result that the ability of the wastewater to resist pH changes (buffer capacity) is important. The H2 CO3 alkalinity plays a central role in establishing the pH buffer capacity of the wastewater. These aspects are discussed in Section 4.14.20.6 which deals with ND, the two biological reactions that most markedly influence H2 CO3 alkalinity and pH. Other inorganic chemical constituents such as magnesium, calcium, potassium, sodium, chlorides, and sulfates are
428
Biological Nutrient Removal
generally of minor significance and need not be routinely measured for wastewater characterization. These inorganic dissolved constituents are needed as trace elements only for biological growth and, in wastewaters, are usually present well in excess of the bacterial requirements, with the result that the greater part of these constituents generally remain dissolved and exit the AS system in the liquid (effluent) stream. However, with BEPR, the cations magnesium and potassium play an important role and quantities greater (5–8 times) than those for normal AS growth are required. Therefore, there may be wastewaters where these cation concentrations in the influent are too low with respect to that required for BEPR, in which event the BEPR would be adversely affected (Lindrea et al., 1994). Many municipal wastewaters also contain potentially toxic metals and elements (PTMEs) such as cadmium (Cd), lead (Pb), nickel (Ni), mercury (Hg), zinc (Zn), copper (Cu), chrome (Cr), cobalt (Co), arsenic (As), fluorine (F), selenium (Se), molybdenum (Mo), and boron (B). The greater part of these PTMEs are in a particulate (nonsettleable or settleable) form and generally accumulate in the sludge mass (primary or secondary) formed at the treatment plant. If the final sludge produced at the treatment plant contains PTMEs exceeding specified limits, then restrictions are placed on the final disposal of the waste sludge.
4.14.4.7 Wastewater Characterization for Plant Wide Modeling Modeling the AS system does not require the carbon composition (total organic carbon, TOC) of the five wastewater organic groups (VFA, FBSO, USO, BPO, and UPO) to be known. Nearly all of the CO2 gas generated by the bioprocesses is stripped from the water via mixing and aeration devices (So¨temann et al., 2005a). It is sufficient to know the COD, TKN, and TP concentrations fractionated into the five organic groups as described in Sections 4.14.4.2–4.4. However, to model the anaerobic digester (AD), the TOC of at least the biodegradable organic groups (VFA, FBSO, and BPO) is also required because the CO2 gas that escapes together with the methane gas establishes the CO2 partial pressure in the AD head space, which with H2CO3 alkalinity (dissolved CO2) establishes the digester pH (So¨temann et al., 2005b, 2005c). So, for the plant-wide models which link ASM1 (Henze et al., 1987) for AS systems and ADM1 (Batstone et al., 2002) for the AD, a compound transformer is interposed between the ASM1 and ADM1 models, which converts the compounds exiting the AS system to the form required for the input to the AD system, while maintaining COD, N and P mass balances. This conversion also adds a carbon composition to the AD input compounds (Volcke et al., 2006). In a different approach, Ekama (2009) established a stoichiometric composition including carbon for the five organic groups in the influent wastewater. The generic composition stoichiometry is in the CxHyOzNaPb form where different x, y, z, a, and b values apply to the five organic groups VFA, FBSO, USO, BPO, and UPO, the first three in their dissolved form and the last two in their settleable and nonsettleable forms. Five unknowns (x, y, z, a, and b) require five facts to determine them. The five facts are the four mass ratios (fcv – (gCOD/g), fC
–(gC/g), fN –(gN/g), and fP – (gP/g)) and the mass balance (fC þ fH þ fO þ fN þ fP ¼ 1) in which the COD replaces fO (gO/ g) and the mass balance replaces fH (gH/g) (Volcke et al., 2006; Ekama, 2009). Therefore, to determine the x, y, z, a, and b values, four ratios need to be determined, viz., COD, TOC, OrgN, and OrgP mass ratios (fcv, fC, fN, and fP) where mass is the VSS for the particulate organics (settleable and nonsettleable) and mass for the dissolved organics. Three of these ratios (fcv, fN, and fP) for the USO and UPO have been in use in AS models already for a long time and were presented in Sections 4.14.4.2–4.4. Also, it was accepted that for the AS system fcv, fN, and fP ratios for the UPO are 1.48 mgCOD/mgVSS, 0.10 mgN/mgVSS, and 0.025 (or 0.03) mgP/mgVSS, respectively, and the same ratios have been accepted for the AS OHO and PAO biomass (without polyP), XBH, XBG, and their endogenous residue, XEH and XEG. Similarly, the N/COD and P/COD ratios of the USO were defined in Sections 4.14.4.2– 4.4. Adding TOC fractionation in exactly the same way as the COD fractionation (Figure 3) and applying the fractionation principles set out above, the fcv, fC, fN, and fP ratios can be determined for all the wastewater organic groups, remembering that the composition of the raw wastewater BPO (settleable þ nonsettleable) is different to that of the settled wastewater BPO (nonsettleable only), due to the removal of the settleable BPO and UPO in the primary sludge (PS). Here, it is interesting to note that the primary settling tanks (PSTs) remove a far greater proportion of UPO (more than twothirds) than BPO (o1/3rd) (Wentzel et al., 2006). This is fortuitous for the AS system because it significantly reduces the UPO, which requires reactor volume (Section 4.14.9.3) and retains a high proportion of BPO, which is required for denitrification. Not all ratios of all the organic groups can be measured, for example, while it is possible to measure the TOC/COD (fC/fcv), OrgN/COD (fN/fcv), and OrgP/COD (fP/fcv) ratios of the USO on filtered WWTP effluents, it is not possible to measure a representative mass of these organics, so one of the four required ratios needs to be guessed – Ekama (2009) used fcv ¼ 1.42 mgCOD mg1). The ratios of the settleable BPO and UPO can be measured in long retention time ADs fed PS, for which the effluent particulate organics comprise mostly UPO (and a very small proportion of AD biomass) and the influent particulate organics comprise the settleable BPO and UPO (Ekama et al., 2006a). An important aspect of this work was determining whether or not wastewater UPO (Supi) and OHO and PAO endogenous residue (XEH and XEG) remained unbiodegradable in the AD. From measured UPO fractions (fS’up) in raw and settled wastewater, the unbiodegradable particulate COD fraction of PS (fSPS’up) can be calculated by mass balance around the PST. In fact, the COD, TKN, and TP concentrations of the PS can be fractionated into the five organic groups represented in the same block diagrams as for raw and settled wastewater (Figures 3–5) by applying the COD, N, P, and VSS (and ISS) mass balances around PST (Wentzel et al., 2006;, So¨temann et al., 2006; Ekama et al., 2006b). The unbiodegradable COD fraction of primary sludge (fSPS’up) so determined matched closely the unbiodegradable COD fraction measured in ADs fed PS (Wentzel et al., 2006, Harding et al., 2009). The same was found for WAS – the unbiodegradable
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COD fraction of the WAS calculated from the steady-state AS model (Section 4.14.9.3) matched closely the unbiodegradable fraction measured in long retention time ADs fed the WAS (Ekama et al., 2006b, Harding et al., 2009). So it could be accepted that the unbiodegradable organics, as measured in the AS system (influent UPO and endogenous residue of each of the OHOs and PAOs), remain unbiodegradable in the AD, which results in these organics passing through the WWTP unchanged. Poinapen and Ekama (2010) applied the stoichiometric organics composition approach for characterizing the BPO in PS to biological sulfate reduction (BSR). From stoichiometric modeling of BSR, they found that most organics, including the BPO in PS, are carbon deficient for BSR in that they can donate more electrons for sulfate reduction than supply carbon for the alkalinity increase in this requires. So all of the COD and C of the utilized BPO (and the very low soluble organics) go to sulfide and H2CO3* alkalinity (dissolved CO2), respectively (and very little to AD biomass). Therefore, PS BPO carbon composition could be determined from the H2CO3* alkalinity generated in the BSR AD. With the stoichiometric composition of the five organic groups known (most measured and some estimated), the influent wastewater characteristics (dissolved, nonsettleable, and settleable) COD, TOC, TKN, TP, and VSS concentrations before (raw) and after (settled) primary settling and of the PS are defined in their block diagram structure by the CxHyOzNaPb composition of the five organics groups (VFA, FBSO, USO, BPO, and UPO) and the FSA and OP concentrations. This allows plant wide stoichiometric modeling with both steady-state (Ekama, 2009) and dynamic kinetic models (Brouckaert et al., 2010), which includes compound products that do not have COD (e.g., CO2) and the proton (Hþ) balance, both of which affect the mixed weak acid/base systems (inorganic carbon, ammonia, VFA, phosphate, and sulfide) that establish the pH in which the bioprocesses occur. Wastewater characterization including stoichiometric CHONP composition therefore opens the way to include, in these models, the three-phase (aqueous, gas, and solid) mixed weak acid/base physical chemistry processes to predict aqueousphase concentrations and pH, gas-phase partial pressure of gases, and mineral precipitation, all very important in anaerobic digestion.
4.14.5.1 Biological Growth Behavior 4.14.5.1.1 Stoichiometry and kinetics To transform a conceptual model into a quantitative kinetic model requires both stoichiometry and kinetics. Stoichiometry gives the quantitative relationships between the various compounds of the conceptual model. For example, for aerobic OHO growth on a simple organic compound such as glucose, the reactants (inputs) are glucose (substrate or electron donor) and oxygen (electron acceptor) and the products are more biomass, water, and CO2. From the bioenergetics of such an aerobic growth process, it can be shown that the mass of organisms formed (anabolism) and oxygen utilized (catabolism) were in a fixed proportion of the mass of organics (substrate) utilized. This fixed proportion is governed by the yield coefficient (YH) and the COD/VSS ratio (fcv) of the organisms. Hence, YH and fcv are stoichiometric constants because they define quantitatively the relationship between the compounds involved in the biological processes (reactions), namely
ðmgVSS l1 Þ
DXBH ¼ YH DSb
DO ¼ ð1 f cv YH Þ DSb
ðmgO l1 Þ
ð40Þ ð41Þ
Although stoichiometry gives the quantitative relationships between the different compounds involved in the biological processes, kinetics considers the rate at which these biological processes take place. Stoichiometric relationships can be changed to kinetic relationships by including the time interval over which changes in the compounds take place, for example, from Equations (40) and (41),
DXBH DSb ¼ YH Dt Dt
ðmgVSS l1 h1 Þ
DO DSb ¼ ð1 f cv YH Þ Dt Dt
ðmgO l1 h1 Þ
ð42Þ ð43Þ
Equations (42) and (43) are kinetic relationships linking the rates of active OHO (XBH) and associated oxygen utilization (O) to the rate of substrate (Sb) utilization via the stoichiometric constants YH and fcv. Mathematically, as the time interval Dt gets infinitesimally small, the finite time interval Dt becomes the derivative dt, viz.,
dXBH dSb ¼ YH dt dt
ðmgVSS l1 h1 Þ
dO dSb ¼ OUR ¼ ð1 f cv YH Þ dt dt
4.14.5 Modeling Biological Behavior To model biological behavior of organisms in WWTPs requires a conceptual model of their behavior in the presence and absence of an external substrate. The former addresses the utilization of biodegradable substrate via anabolism (synthesis of cell mass) and catabolism (generation of energy) by the relevant organism group, such as to the utilization of readily biodegradable organics (RBOs) by the OHOs, or the utilization of FSA by the ANOs. The behavior of organisms in the absence of an external substrate addresses the decline in organism numbers or mass under these conditions. The conceptual models of organism growth and decline on which the AS kinetic models are built are presented below.
429
ðmgO l1 h1 Þ
ð44Þ ð45Þ
Note from Equations (40)–(45) that the COD (or e) balance is conserved because
dSb dXBH dO þ f cv ðÞ ¼0 dt dt dt
ðmgCOD l1 h1 Þ
ð46Þ
In Equation (46), both the substrate and oxygen decrease and the active organism concentration increases as would happen in an aerobic batch test in which soluble organics and OHOs are mixed in aerated water. In the biological growth process, in which the reactants (inputs) are organics and oxygen and the products (outputs) are organisms, carbon dioxide and water,
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the electrons (COD) in the organics are conserved in organisms formed and water produced. However, because it is not possible to measure the concentration of the water produced, the oxygen utilized is measured instead. Oxygen utilization (OU) measurement: (1) is permissible due to the proportionality between oxygen utilized and water produced (4e þ 4Hþ þ O2-2H2O), (2) is relatively simple in AS reactors (e.g., Randall et al. 1991), and (3) is in conformity with the COD test. Because oxygen is a reactant in the growth process (i.e., a decreasing input) and replaces the water product (i.e., an increasing output), the sign of the oxygen utilization term in the COD mass-balance Equation (46) is changed to positive by multiplying it by a negative sign. This is done in all mathematical models which are formulated on the basis of COD balance (such as ASM1, UCTOLD, ASM2, and UCTPHO) to ensure that the COD in fact balances, that is, over a welldefined time interval the COD of the organisms formed (fcv DXBH) plus the oxygen utilized (DO) must be equal to the COD utilized (DSb). When an OHO population under aerobic conditions is brought into contact with biodegradable organics of soluble readily biodegradable (RBSO) and particulate slowly biodegradable (BPO) forms (see Figure 3), their response may be described qualitatively (conceptually) as follows: 1. The soluble (RBOs) passes directly through the cell wall and is metabolized (utilized) at a high rate. 2. The particulate slowly biodegradable organics (BPO) is enmeshed or entrapped in the sludge mass and some of it is adsorbed onto the active organisms. These enmeshment and adsorption reactions are rapid and effectively remove most of the particulate and colloidal organics from the wastewater. The adsorbed organics are broken down via a biologically assisted hydrolysis process to smaller and simpler organics by extracellular enzymes and the simpler organics are transferred through the cell wall and metabolized in the same manner as the RBOs in (1) above. The rate of the hydrolytic enzymatic breakdown is relatively slow and is the limiting (slowest) rate in the overall growth process on slowly biodegradable organics, only about onetenth of the rate for the RBOs.
4.14.5.1.2 Monod growth kinetics for utilization of RBSO The utilization rate of RBSO is described qualitatively via Monod kinetics. Monod kinetics is described in some detail below because it is a very useful mathematical expression in biological process modeling. Not only is it applied directly to model the utilization of all soluble substrates, such as utilization of RBSO by OHOs and FSA by ANOs (nitrification), its form is used to model the hydrolysis/utilization of BPO and also as swithing functions for the progressive phasing-in and -out of biological kinetic processes as environmental conditions in reactor(s) change from aerobic (dissolved oxygen, DO present) to anoxic (DO absent) conditions and vice versa. Monod (1950) defined the term specific growth rate (m) as the increase in organism concentration per unit time (dXBH/ dt) per average concentration of organisms present. Mathematically, this is equivalent to logarithmic growth and is
Monod specific growth rate curve 3.0 Specific growth rate (d−1)
430
Maximum specific growth rate = UHm = 2.5 d−1
2.5 2.0 1.5
Half-saturation coefficient Ks = Substrate concentration at which UH = UHm/2 = 10 mg COD l−1
1.0 0.5 0.0 0
10 20 30 40 50 Substrate concentration (mgCOD l−1)
60
Figure 6 Monod specific growth rate (mH, d1) vs. soluble substrate concentration (Sbs, mgCOD l1) curve.
expressed as
dXBH 1 ¼ mH dt XBH
ðmgOHOVSS=ðmgOHOVSS dÞÞ
ð47Þ
Further, Monod conducted batch experiments in which the soluble biodegradable substrate concentration of defined substrates such as glucose remained essentially constant and, in different chemostat (flow through) reactor tests, observed how the specific growth rate (mH) of certain organism species in pure culture changed at different defined soluble organic substrate concentrations (Sbs). The form of the relationship observed by him between mH and Sbs is shown in Figure 6 and can be expressed mathematically as
dXBH 1 mHm Sbs ¼ mH ¼ dt XBH KS þ Sbs
ð48Þ
where mHm is the maximum specific growth rate (d1) and KS the half-saturation coefficient, which is the concentration of the substrate at which the specific growth rate is half the maximum (mg l1) (Figure 6). Monod determined the mHm and KS constants for different organism species in pure culture growing on various soluble organic substrates and found that the mHm and KS values were different for different organism species–substrate type combinations. Despite this, the Monod equation (Equation (48)) was adopted into biological WWTP modeling in which not only the active OHO population is highly diverse but also the soluble biodegradable organics are innumerable. Because clearly not all the OHO species and soluble organic compounds could be modeled individually in biological WWTP models (level of organization too low), measures are adopted that lump together all the OHO species in a single surrogate organism group, that is, the active part of the volatile settleable solids (OHOVSSs) and the innumerable biodegradable organics into a single substrate, that is, the COD. Therefore, the units for the mHm and KS constants in the Monod equation as applied to modeling biological WWTPs are mgActiveOHOVSS/(mgActiveOHOVSS d) and mgCOD l1 respectively, as shown in Equation (48).
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From Figure 6 and Equation (48), there are two extremes of specific growth behavior described by Monod’s equation: (1) When the substrate concentration is very high in relation to KS so that KS contributes negligibly to the denominator (SbsdKS), changes in Sbs do not change mH which remains constant at mHm. Hence, the specific growth rate is zero order with respect to the substrate concentration, that is, (mH ¼ mHmS0bs ). (2) When the substrate concentration is very low in relation to KS so that Sbs contributes negligibly to the denominator (KSdSbs), changes in Sbs change the rate in proportion to Sbs. Hence, the specific growth rate is first order with respect to the substrate concentration (mH ¼ mHm/KSS1bs ). Monod’s equation (Equation (48)) therefore changes from first order to zero order with respect to the substrate concentration as the substrate concentration increases from zero to several times higher than the KS value. This characteristic is very useful in biological process modeling. First, it is a selfcontrolling kinetic rate; when the substrate concentration is high, the rate is a maximum and when the substrate concentration is zero the rate is zero. Second, the Monod term is a useful switching function if a biological process is required to operate only under aerobic conditions. To achieve this, the biological process rate equation is multiplied by the Monod term O/(KO þ O) where O is the dissolved oxygen (DO) concentration and KO is the sensitivity of the biological process rate to DO. If the DO concentration is high, then the process rate is zero order with respect to DO and process proceeds at its defined rate independent of the DO concentration. If the DO concentration is zero, the biological process rate also is zero as required for an exclusively aerobic process. If the DO concentration is less than KO, then the biological process will be less than half its maximum rate and dependent on the DO concentration and if the DO concentration is greater than KO, the biological process rate will be more than half its maximum rate. The KO value therefore controls the sensitivity of the biological process to DO; if KO is high, say 3 mg l1, then the biological process rate is lower than 75% of its maximum value up to a DO concentration of 9 mgO l1, which is the approximate saturation value for DO in freshwater aerated with air at 1 atm pressure (760 mmHg) and 20 1C; if the KO value is low (0.1 mgO l1), then the biological process rate reaches 95% of its maximum value at a DO concentration of 2 mgO l1. The above switching function can also be applied if an anoxic process needs to be switched on simultaneously as an associated aerobic process is switched off. To achieve this, the anoxic biological process rate is multiplied by 1 the above Monod term, that is, 1 O/ (KO þ O) ¼ KO/(KO þ O). Clearly, the Monod term is a relatively realistic and very convenient mathematical means whereby biological processes can be switched on or off in conformity with the conditions in the biological reactor and hence is extensively used in AS simulation models (e.g., ASM1 and ASM2). Substituting the Monod (equation Equation (48)) for the specific growth rate (mH, Equation (47)) yields the rate of growth of active OHOs as a function of substrate RBSO concentration, viz.,
dXBH mHm Sbs XBH ¼ dt KS þ Sbs
ðmgOHOVSS l1 h1 Þ
ð49Þ
dSbs mHm Sbs XBH ¼ dt YH KS þ Sbs
ðmgCOD l1 h1 Þ
dOs mHm Sbs ¼ OURs ¼ ðÞ ð1 f cv YH Þ XBH dt YH KS þ Sbs ðmgO l1 h1 Þ
431
ð50Þ
ð51Þ
where dOs/dt (or OURs) is the oxygen utilization rate (OUR) for OHO synthesis.
4.14.5.1.3 Active site surface kinetics for hydrolysis/ utilization of BPO The particulate slowly biodegradable organics (BPO, Sbp), comprising all the nondissolved organics, when mixed AS including active OHOs, are enmeshed into the sludge mass by a biologically assisted flocculation process. Because AS has such a strongly flocculent nature, this enmeshment process is efficient and rapid and effectively removes most of the particulate organics, both nonsettleable (colloidal) and settleable, from the wastewater liquid phase onto the AS solids phase (Senm). Therefore, if after a contact time as short as 1 h between the municipal wastewater particulate organics and AS, the sludge mass is allowed to settle, a COD removal of over 80–85% is achieved, the RBSO (Sbs) through rapid utilization and the BPO (Sbp) through rapid enmeshment and adsorption. This is in fact the principle on which the contact stabilization AS is based (Gujer et al., 1975a, 1975b; Alexander et al., 1980). Once enmeshed, some of the BPOs (Sbp) are adsorbed onto the active OHOs (XBH) in the AS. This adsorption process, which is also a physical one like the enmeshment, brings the active organisms into close contact with the particulate organics. The organisms have a finite capacity to absorb particulate organics. Conceptually, the adsorption process is modeled as an active site surface kinetic reaction approach in which particulate organics are adsorbed onto the organism mass until their active sites are all occupied, at which point the organisms are saturated with particulate organics (XS, mgVSS l1) (Dold et al., 1980). Once adsorbed, the particulate organics (XS) are broken down into small RBOs via a biologically assisted enzymatic hydrolysis process. The RBSO product of this hydrolysis process then enters the organism through the cell wall and is utilized by the same anabolic and catabolic growth processes as the RBSO of the wastewater. The adsorption rate is modeled as a first-order rate with respect to the enmeshed biodegradable COD concentration (Senm) and the active OHO concentration (XBH) and also includes a term that describes the degree of saturation of the adsorption sites on the organisms, viz., (fma XS/XBH), where XS /XBH is the concentration ratio of adsorbed biodegradable organics (as VSS) and the active OHOs (also as VSS). When the adsorption sites are all full, the XS /XBH ratio equals fma and the adsorption rate is zero. In the UCTOLD dynamic simulation model (Dold et al., 1991), fma is assigned a value of 1 indicating that the active OHOs can adsorb their own mass of particulate organics (XS), viz.,
dSenm XS ¼ Ka Senm XBH f ma dt XBH
432
Biological Nutrient Removal
where Senm is the enmeshed particulate slowly biodegradable organics concentration (BPO, mgCOD l1), Ka the specific BPO adsorption rate [l/(mgVSS d)], XS the adsorbed particulate slowly biodegradable organics concentration (mgVSS l1), and fma the maximum adsorbed organics to active OHO concentration ratio (mgVSS/mgVSS). In the adsorption, hydrolysis, and utilization processes of BPO, the hydrolysis process is the limiting one because it has the slowest rate. This process therefore governs the overall specific growth rate of the active OHOs on BPOs. Unlike the soluble organics, the utilization rate of which is governed by its concentration in the liquid phase (see Equation (50)), the rate of hydrolysis of adsorbed particulate organics (XS) is governed by the degree of saturation of the active adsorption sites XS/XBH. From a comparison of predicted and experimental results, Dold et al. (1980) found that (1) the Monodtype equation best described the relationship between the specific hydrolysis rate and the XS/XBH ratio; (2) the power on the XS/XBH ratio, which defines the surface to volume ratio of the material on which the reaction takes place, was 1, that is, the active OHOs could be viewed simply as a planer surface rather than a spherical one (for which n ¼ 2/3); and (3) the soluble RBSO and particulate BPO are hydrolyzed/utilized simultaneously by the active OHOs. Because the specific OHO growth rate on the BPO hydrolysis product is as fast as the specific hydrolysis rate, the former is equal to the latter and is expressed by
Kmp ðXS =XBH Þ dXBH 1 ¼ þYH dt XBH ½Ksp þ ðXS =XBH Þ
ð52Þ
where Kmp is the maximum specific hydrolysis rate mgCOD/ (mgOHOVSS d) and Ksp the half-saturation coefficient for hydrolysis of BPO (mg COD/mg COD). The associated OUR is directly proportional to the specific hydrolysis/utilization rate of BPO via the catabolic stoichiometric factor (1 fcvYH) (see Equations (43) and (44)). It is mainly in the hydrolysis/utilization of the BPO that the UCTOLD and IWA ASM No1 models differ (Dold and Marais, 1986, Hu et al., 2003). The former includes instantaneous enmeshment (Senm), a defined kinetic adsorption process (Senm-XS), and the BPO (XS) specific hydrolysis/ utilization rate (XS-XBH) is governed by the XS/XBH ratio. ASM1 (and ASM2) includes only instantaneous enmeshment, which directly becomes adsorbed BPO (Senm ¼ XS), and the specific hydrolysis rate is governed by the ratio Senm/XBH. Furthermore, in ASM1, the BPO hydrolysis product (RBSO, Sbs) is not directly utilized in a combined hydrolysis/utilization process as in the UCTOLD model. Instead, the BPO hydrolysis/products are released to the bulk liquid and utilized as RBSO taken up from the liquid phase like the wastewater RBSO via the Monod kinetics Equation (50). Therefore, in ASM1, only RBSO is utilized for OHO growth and associated oxygen utilization. Further details on the biological growth kinetics of the UCTOLD and ASM No1 simulation models can be found in Dold et al. (1980, 1991) and Henze et al. (1987), respectively, and are compared by Dold and Marais (1986) and Hu et al. (2003).
4.14.5.2 Organism Decline 4.14.5.2.1 Endogenous respiration Experimentally, it has been observed that after all the biodegradable COD, whether of a soluble or particulate form, added to an aerated batch test on real AS has been utilized, oxygen continues to be utilized and the VSS concentration decreases for 20 days or longer. There are various conceptual models which seek to describe this simultaneous oxygen utilization and VSSs decrease in the absence of an externally added substrate. The most widely accepted is called the endogenous respiration and mass loss model. This model dates back to the earliest empirical AS models and is used even today in the dynamic kinetic models for the ANOs and PAOs. The problem with this model was not so much its concept and mathematics but the determination of the kinetic and stoichiometric constants associated with it. Once the determination of endogenous respiration rate independently of the yield coefficient (YH) was established via batch tests on AS harvested from continuous steady-state systems (Marais and Ekama, 1976; van Haandel et al., 1998), this model is acceptable and so is retained for the steady-state OHO AS model. In the endogenous respiration conceptual model, in the absence of an externally available substrate, the organism utilizes the organics (COD) of its own cell mass to generate energy via catabolism for essential cell functions. The utilization of the organism organics in which the e of these organics are passed to oxygen accounts for the continued simultaneous VSS decrease and utilization of oxygen. However, in order to quantitatively match the observed changes in VSS reduction and oxygen utilization, it appears that associated with this endogenous respiration process is the generation of unbiodegradable endogenous VSS residue. So mathematically, organism endogenous mass loss/respiration is modeled as follows: Over a defined period of time, dt, the decrease in active organism concentration (XBH) is proportional to the concentration of organisms present, viz.,
dXBH ¼ bH XBH dt
ðmgOHOVSS l1 h1 Þ
ð53Þ
where bH is the endogenous respiration rate (d1). However, the decrease in active organism VSS concentration is not all biodegradable organics. A fraction, fEH, is unbiodegradable and accumulates as endogenous residue denoted XEH. Hence, the endogenous residue accumulation rate is given by
dXEH ¼ þf EH bH XBH dt
ðmgEVSS l1 h1 Þ
ð54Þ
where fEH is the unbiodegradable fraction of the OHOs. The biodegradable part of the active organism concentration decrease is therefore (1 fEH)bHXBH mgVSS l1 h1. These biodegradable organics the organisms use catabolically to generate energy for maintenance of essential cell functions. All the electrons (e) of these biodegradable organics are therefore passed to oxygen to generate the energy. As (1 fEH)bHXBH is the rate of utilization of biodegradable
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organism organics in VSS units, to express this in oxygen or COD units, it needs to be multiplied by the COD/VSS ratio fcv of the OHOs. Because these organics are used catabolically and none is conserved in new organism mass, all the electrons of the biodegradable organism organics are passed to oxygen. Hence, the utilization rate of organism organics is the OUR for endogenous respiration, viz.,
dOe ¼ OURe ¼ ðÞf cv ð1 f EH ÞbH XBH dt
ðmgO l1 h1 Þ ð55Þ
where dOe/dt (or OURe) is the OUR for endogenous respiration (mgO l1 h1) and fcv the COD/VSS ratio of the organism organics (mgCOD/mgVSS). Again because oxygen is decreasing (utilization to form water), it is negative but as it is a reactant and not a product in the biological process, it is given an additional negative sign to make the overall OUR for endogenous respiration þ ve to maintain the COD balance. The equations (53)–(55) satisfactorily model the changes in VSS concentration and OUR with time in the absence of an externally added substrate. From many aerobic batch tests on AS harvested from continuous flow laboratory scale systems at 8, 14, and 20 1C, and accepting an endogenous residue fraction fEH ¼ 0.20 from McKinney and Symons (1964), Marais and Ekama (1976) found that the bH rate (1) did not change the sludge age, (2) was 0.24/d at 20 1C and (3) was mildly temperature sensitive between 8 and 20 1C with an Ahrrenius constant ybH ¼ 1.029, viz.,
bHT ¼ bH20 ð1:029ÞðT20Þ
ðd1 Þ
ð56Þ
This endogenous respiration rate was subsequently confirmed by van Haandel et al. (1998) and found to be slightly more temperature sensitive between 20 and 30 1C with a ybH ¼ 1.04.
4.14.5.2.2 Death regeneration The second conceptual model for describing the continuous oxygen utilization and VSS reduction with time in the absence of an externally added substrate is called death regeneration. In this model the active organisms die and lyse (release) all their organic material to the bulk liquid. The unbiodegradable part of this released organic material, which is all of a particulate nature, remains as endogenous residue or organism ‘‘skin and bones’’. The biodegradable part is utilized by the remaining active organisms via the identical biological growth processes of anabolism and catabolism as influent wastewater organics to form new active organism mass. The oxygen utilized is now that required for the synthesis of new organism mass from the biodegradable organics of the organisms that have died. The death regeneration model concept was incorporated into the UCTOLD AS model by Dold et al. (1980). From an evaluation of this model it was found that if it was assumed that all the biodegradable organics are utilized, acceptable for the steady-state model, then there was no difference between this model and the endogenous respiration model. However, to reproduce the observed experimental results, the death rate 0 ) (b’H) and unbiodegradable fraction of the organisms (fEH needed to be 0.62/d and 0.08, respectively. The similarity of
433
the two conceptual models for aerobic steady-state conditions is demonstrated in Table 5. Incorporation of the death regeneration model into the dynamic kinetic model, in which the utilization of slowly BPO is modeled with different kinetic rates than the RBSO and not all of these fractions are necessarily completely utilized, it was found that for aerobic conditions the two models yielded virtually identical results provided most of the COD was utilized. To achieve similarity it was necessary to accept that the biodegradable COD released by the dead organisms was the same as influent BPO and that no soluble organics of the biodegradable or unbiodegradable form were released by the dead organisms. Although for aerobic systems there was little to choose between the endogenous respiration and death regeneration models because both yielded equally close predictions, the latter is conceptually more consistent because it does away with the young fat organism and old thin organism outcome of the endogenous respiration model. With death regeneration, 99% of the active organisms are replaced in 7.4 days so there is a continuous generation of new organisms growing on the biodegradable organics of the dead ones. Also, an unbiodegradable fraction of 0.08 is a more reasonable value for microorganisms than the high 0.2 value for endogenous respiration model. Although for aerobic conditions and sludge ages longer than about 3 days (i.e., 495% biodegradable COD utilization) the two conceptual models gave virtually identical results (Dold et al., 1980), the advantage of the death regeneration model became evident when modeling anoxic–aerobic conditions. First, no change in organism behavior, and therefore kinetic equations, was needed in the model when oxygen and/ or nitrate became limiting. Under these conditions, the growth processes (i.e., RBSO and BPO utilization) cease but the death processes continue, leading to an accumulation of enmeshed and adsorbed BPO from the dead organisms until oxygen and/or nitrate again become available. Because this led to improved simulations for anoxic–aerobic systems (Van Haandel et al., 1981), the death regeneration model was incorporated into the dynamic kinetic model (e.g., UCTOLD and ASM1). However, because it is assumed in the steady-state model that all the biodegradable organics are utilized, the endogenous respiration model is retained for the steady-state model as it leads to much simpler process equations (Marais and Ekama, 1976; WRC, 1984; Henze et al., 2008) which are presented in this chapter.
4.14.6 AS System Constraints Basically, all aerobic biological treatment systems operate on the same principles, that is, trickling filters, aerated lagoons, contact stabilization, extended aeration, etc., differ only in the conditions under which the biological reactions are constrained to operate, called system constraints. The AS system comprises the flow regime in the reactor, the size and shape, number and configuration of the reactors, recycle flow, influent flow, and other features either incorporated deliberately or present inadvertently or unavoidably. Whereas the response of the organisms is in accordance with their nature, that is, biological process behavior, that of the system is governed by a
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Biological Nutrient Removal
Table 5 Comparison of the endogenous respiration (End-Res) and death regeneration (Dth-Reg) models for describing organism behavior in the absence of an externally supplied substrate Parameter
End-Res
Dth-Reg
Units
Active organism mass loss rate (bH) Unbiodegradable fraction of organisms (fEH) COD/VSS ratio of active organisms (fcv) At the start of 1 day Hence COD at start ¼ fcv100 Active VSS lost in 1 day ¼ bH100 Active organisms remaining ¼ (1 bH)100 COD of active organisms lost ¼ fcvbH100 Endogenous residue formed ¼ fEHbH100 COD of endogenous residue ¼ fcvfEHbH100 Biodeg. organics released ¼ (1 fEH)bH100 COD of biodeg. org released ¼ fcv(1 fEH)bH100 New active organisms formed ¼ YH 84.4 Oxygen utilized ¼ (1 fcvYH)fcv(1 fEH)bH100 Active mass at the end of 1 day
0.24 0.20 1.48 100 148 24 76 35.5 4.8 7.1 19.2 28.4 0 28.4a 76
0.62 0.08 1.48 100 148 62 38 91.8 4.8 7.1 57.0 84.4 38.0 28.4 76
mgActiveVSS/mgActiveVSS/d mgUnbioVSS/mgActiveVSS mgCOD/mgVSS mgVSS/l active organisms mgCOD/l active organisms mgVSS/l active organisms mgVSS/l active organisms mgCOD/l active organisms mgVSS/l endogenous residue mgCOD/l endogenous residue mgVSS/l biodeg. organics mgCOD/l biodeg. organics mgVSS/l active organisms MgO/l mgVSS/l active organisms
Endogenous respiration model
Death regeneration model
Endogenous residue formed 4.8%
Endogenous residue formed 4.8%
Oxygen utilized 19.2%
New active VSS formed 38%
Active VSS remaining 76%
Active VSS remaining 38%
Active VSS after 1 day 76%
Active VSS after 1 day 76%
Unbiopart of XBH = End residue Active VSS loss d−1 bHXBH
Biodeg part (1−fEH)b HXBH Active VSS remaining (1−bH)XBH
Oxygen utilized 19.2%
Oxygen demand for synthesis ′ )b′HXBH (1−fcvYH)fcv(1−fEH
fEHbHXBH
Oxygen demand for end resp = COD biodeg XBH fcv(1−fEH)b HXBH
New active VSS synthesized ′ )b′HXBH YHfcv(1−fEH
Biodeg COD of dead active VSS ′ )b′HXBH fcv(1−fEH
Net active VSS loss Active VSS gain
Active VSS loss d−1 b′HXBH
Active VSS regrown Active VSS remaining (1−b′H)XBH
Unbio VSS of XBH loss = end residue ′ b′HXBH fEH
a
Oxygen utilized ¼ COD of biodegradable organics released.
combination of the organism behavior and the physical features that define the system, that is, the environmental conditions or system constraints under which the biological processes are constrained to operate.
4.14.6.1 Mixing Regimes In the AS system, the mixing regime in the reactor and the sludge return are part of the system constraints and therefore
Biological Nutrient Removal
influence the response of the system – hence, consideration must be given to reactor mixing regimes. There are two extremes of mixing: completely mixed Figure 7(a) and plugflow Figure 7(b). In the completely mixed regime, the influent is instantaneously and thoroughly mixed (theoretically) with the reactor contents. Hence, the effluent flow from the reactor has the same compound concentrations as the reactor contents. The reactor effluent flow passes to a settling tank; the overflow from the tank is the treated waste stream; the underflow is concentrated sludge and is recycled back to the reactor. In the completely mixed system, the rate of return of the underflow has no effect on the biological reactor except if an undue sludge buildup occurs in the settling tank. The shape of the reactor is approximately square or circular in plan, and mixing is usually by mechanical aerators or diffused air bubble aeration. Examples are extended aeration plants, aerated lagoons, Pasveer ditches, and singlereactor completely mixed AS plants. In a plugflow regime, the reactor usually is a long-channeltype basin. The influent is introduced at one end of the channel, flows along the channel axis, and is mixed by air spargers set along one side of the channel or horizontal shaft surface aerators. Theoretically, each volume element of liquid along the axis is assumed to remain unmixed with the elements leading and following. Discharge to the settling tank takes place at the end of the channel. To inoculate the influent waste flow with organisms, the underflow from the settling tank is returned to the influent end of the channel. This creates an intermediate flow regime deviating from true plugflow conditions depending on the magnitude of the recycled underflow. Conventional AS plants are of the intermediate flow regime type with sludge return recycle ratios varying from 0.25 to 3 times the average influent flow rate. If the recycle ratio is very high, the mixing regime approaches that of completely mixed. Intermediate flow regimes are also achieved by having two or more completely mixed reactors in series, or by step
Aeration
Influent
Aerobic reactor
Waste flow Secondary settling tank Effluent
Sludge recycle
(a)
Aeration Influent
Waste flow Secondary settling tank
Aerobic reactor
435
aeration. In the latter, the influent is fed at a series of points along the axis of the plugflow-type reactor. Both configurations require, for inoculation purposes, recycling of the settled sludge from the settling tank(s) to the start of the channel reactor. The mean kinetic response of an AS system (i.e., sludge mass, daily sludge production, daily oxygen demand, and effluent organics concentration) is adequately, indeed accurately, given by assuming that the system is completely mixed and the influent flow and load are constant. This allows the reactor volume, the mass of sludge wasted daily, and average daily oxygen utilization rate to be determined by relatively simple formulations. Peak OURs which arise under cyclic flow and load conditions can be estimated subsequently quite accurately by applying a factor to the average OUR. These factors have been developed from simulation studies with the simulation models on aerobic and anoxic–aerobic systems operated under cyclic and under constant flow and load conditions (see Section 4.14.27.2).
4.14.6.2 Solids Retention Time or Sludge Age In the schematic diagrams for the AS system (Figure 7), the waste (or surplus) sludge is abstracted directly from the biological reactor. The common practice is that the waste sludge is abstracted from the secondary settling tank underflow. Sludge abstraction directly from the reactor leads to a method of control of the sludge age, called the hydraulic control of sludge age, which has significant advantages for system control compared to abstracting wastage via the underflow. The sludge age, Rs in days, is defined by
Rs ¼
Mass of sludge in reactor Mass of sludge waste per day
ðdaysÞ
ð57Þ
By abstracting the sludge directly from the reactor, the sludge concentrations in the waste flow and biological reactor are the same. If a sludge age of, say, 10 days is required, one-tenth of the volume of the reactor is wasted every day. This can be achieved by a constant waste flow rate, Qw (l d1), where Qw is the volume of sludge to be wasted daily. Hence,
Rs ¼
X Vp Vp ¼ X Qw Qw
ðdaysÞ
ð58Þ
where Vp is the volume of the biological reactor (l) and Qw the waste flow rate from reactor (l d1). Equation (58) assumes that the mass of sludge in the secondary settling tanks is negligible relative to that in the biological reactor. This assumption is reasonable when the system is operated at relatively high recycle ratios (B1:1) and the sludge age is longer than about 3 days (see Section 4.14.14).
4.14.6.3 Nominal Hydraulic Retention Time Effluent
Sludge recycle (b)
Figure 7 (a) Activated sludge system with a completely mixed reactor mixing regime and (b) plugflow/intermediate reactor mixing regime.
In AS theory, the volume of the process per unit of volume of influent flow is known as the nominal hydraulic retention time (HRT), that is,
Rhn ¼
Vp Qi
ðdaysÞ
ð59aÞ
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Biological Nutrient Removal
where Rhn is the average nominal hydraulic retention time (days) and Qi the daily average influent flow rate (l d1). When the sludge return flow from the secondary settling tank (Qs) and any other mixed liquor recycle flow entering the reactor (Qa) are included, the retention time is called the actual hydraulic retention time (Rha), viz.,
Rha ¼
Vp Rhn ¼ Qi þ Qs þ Qa 1 þ s þ a
ðdaysÞ
Mass-balance boundary Influent Q i, S bsi
4.14.6.4 Connection between Sludge Age and Hydraulic Retention Time From the above definitions, it can be seen that there are two parameters that relate to time in the system: (1) the sludge age (Rs), which gives the length of time the particulate material remains in the reactor and (2) the nominal hydraulic retention time (Rhn), which gives the length of time the liquid and dissolved material remains in the reactor. In AS systems which do not have solid–liquid separation with membranes or secondary settling tanks (SSTs), such as aerated lagoons, the sludge age and nominal hydraulic retention time are equal, that is, the liquid/dissolved material and the solids/particulate material remain in the reactor for the same length of time. When solid–liquid separation is included, then the liquid and solid retention times are separated and Rs4Rhn. However, long sludge ages (Rs) lead to large sludge masses in the reactor, which, in turn, lead to large reactor volumes (Vp). Therefore, even with solid–liquid separation, as Rs gets longer, so also does Rhn. This link between Rs and Rhn is neither proportional nor linear and depends on (1) the wastewater organic (COD or BOD5) concentration and (2) the reactor suspended solids concentration (TSS). For biological nutrient removal AS systems, the sludge age is around 10–25 days and the nominal hydraulic retention time around 10–24 h.
4.14.7 Model Development – Completely Mixed Aerobic System The system equations are derived by doing mass balances over the AS system on the compounds identified as important for the model. The mass balance considers the rates of increase and decrease in the mass of a particular compound due to this compound’s inflow to and outflow from the system and its production or degradation via the biological processes in the reactor. To define the system, the inflows to it and outflows from it for the mass balance, a boundary needs to be set. The elements within the boundary constitutes the system. The system boundary can be drawn around the reactor only because the biological processes take place in the reactor. However, when this is done, the recycle flow constitutes an inflow to the system and complicates the final steady-state model equations considerably. Therefore, in the interests of the simplicity, the boundary for the mass balances includes both the biological reactor and the settling tank, and only one
Waste flow Qw, XBH, XEH, Sbs
Vp, XBH XEH, S bs OURc QR
ð59bÞ
where Rha is the actual hydraulic retention time (days), s the sludge underflow recycle ratio (Qs/Qi), and a the mixed liquor recycle ratio (Qa/Qi).
Oxygen in Oi
Effluent Qe = Q i − Q w Sbse
Figure 8 Completely mixed activated sludge system showing system boundary (dotted lines) for mass-balance equations.
inflow (influent, Qi) and two outflows (effluent, Qe, and waste, Qw) need to be considered (Figure 8). In the mass balance, it is assumed that (1) no biological activity takes place in the settling tank, thereby confining biological activity to the reactor only and (2) the water mass flow into the system (entering over the boundary) is equal to the water mass flow out of the system (exiting over the boundary). Water losses due to, for example, evaporation or gains due to, for example, biological activity (from oxygen being utilized as terminal electron acceptor) are considered negligible (o1%) in terms of the estimated accuracy of the model (B5%). Hence, the generic compound mass balance over a short time interval, Dt, is given by
2
3 2 3 2 3 Mass Mass flow Mass flow 6 7 6 7 6 7 4 change 5 ¼ 4 into 5 4 out of 5 in system system system 2 3 2 3 Mass gain Mass loss 6 7 6 7 þ4 by by 54 5 bioprocess bioprocess
ð60Þ
This mass-balance equation is applied to each compound (or system variable) of interest in the model. For the steady-state model for organic material removal only, these are 1. 2. 3. 4. 5. 6.
UPOs (XI) (from Supi in the influent), unbiodegradable soluble organics COD (Sus), active OHO organisms (XBH), endogenous residue (XEH), biodegradable COD (Sb), and oxygen (O).
4.14.7.1 Building Up the Model in Stages To gain greater insight into the behavior of the various components in the AS system, the model is built up in stages by considering first the behavior of the influent particulate (Supi) and soluble (Susi) unbiodegradable organics in the system and then that of the biodegradable organics. The influent UPOs (Supi in COD units or Xli in VSS units) cannot be measured directly in the wastewater because they are mixed with the BPOs; likewise, the influent unbiodegradable soluble organics are mixed with the biodegradable soluble organics. These unbiodegradable constituents can only be determined by treating the wastewater in a long sludge age mostly aerobic AS system to ensure that virtually all the biodegradable organics are
Biological Nutrient Removal
utilized. The proportion of the influent particulate and soluble organics that are not degraded are then accepted to be the influent unbiodegradable particulate and soluble COD concentrations Supi and Susi, respectively. To see how these two unbiodegradable constituents behave in the AS system, the stoichiometric and steady-state equations for these two constituents are derived and discussed first. Thereafter, the biodegradable organics of the wastewater are considered. The complete steady-state model for the AS system for all four wastewater organic fractions is then developed by adding the component equations of the four different wastewater organic constituents. This approach also provides insight into the only reliable method developed to date for determining the UPO concentration (Supi) or fraction (fS’up) (Ekama et al., 1986; Wentzel et al., 1999).
437
flows, (flux) viz.,
MXI ¼ Vp XI ¼ Mass XI in reactor FXIi ¼ Qi XIi ¼ Flux XI into reactor
ðmgVSSÞ
ð64Þ
ðmgVSS d1 Þ
ð65Þ
where the prefix M denotes mass and F denotes flux of the compound that follows it. Note that with this nomenclature all compound concentrations in the reactor are multiplied by the reactor volume to transform these to mass in the reactor (mg) and compound concentrations in the influent (Qi), waste (Qw), and effluent (Qe) flows are multiplied by their respective flows to transform these to fluxes (mg d1). Substituting Vp/Qi for Rhn in Equation (62) and multiplying through by Vp yields
4.14.7.1.1 Unbiodegradable particulate organics (Supi) The material, which comprises hair, cellulose, fibrous material, cloth remnants, and other UPOs, accumulates in the biological reactor. It settles out with the AS in the secondary settling tank and therefore is retained in the system. The only stream by which this material exits the system is via the waste flow Qw. This material is denoted XIi in VSS units in the influent and is denoted XI in VSS units in the reactor. Applying the general mass-balance equation (60) to this material yields
Vp dXI ¼ Qi XIi dt 0 Qw XI dt þ 0 0
ð61Þ
where dXI is the change in XI concentration in the reactor in terms of the VSS (mgVSS l1) and dt the time interval of the mass balance. In the mass balance it is assumed that (1) the settling tank is 100% efficient in that all the XI material is captured and returned to the biological reactor and (2) by definition of unbiodegradable, no gains or losses of this material take place by bioreaction. Dividing both sides of Equation (61) by Vpdt yields
dXI Qi Qw ¼ XIi XI dt Vp Vp and noting that Qi/Vp ¼ 1/Rhn and Qw/Vp ¼ 1/Rs yields
dXI XIi XI ¼ dt Rhn Rs
ð62Þ
Now at steady state, the transient dXI/dt is zero because the XI concentration in the reactor has reached equilibrium between the inflow and outflow masses of this compound with the influent and waste flows, respectively. So, setting dXI/dt ¼ 0 and solving for the reactor UPO concentration XI which accumulates from the influent UPOs (XIi) yields
XI ¼ XIi Rs =Rhn
ðmgVSS=l reactorÞ
ð63Þ
For design Equation (63) is not very useful because it includes the nominal hydraulic retention time Rhn which is not known because the reactor volume is not yet known. This difficulty can be eliminated by considering mass and mass
Vp XI ¼ Qi XIi Rs and applying Equations (64) and (65) yields
MXI ¼ FXIi Rs
ð66Þ
Equation (66) shows that at steady state the mass of UPOs (as VSS) in the reactor is equal to the flux of this material into the reactor times the sludge age Rs. Because XIi is found from the influent unbiodegradable particulate COD concentration, that is, XIi ¼ Supi/fcv, the mass of this material in VSS units that accumulates in the reactor at steady state is given by
MXI ¼ FSupi =f cv Rs ¼ Qi ðSupi =f cv ÞRs
ð67Þ
Dividing Equation (67) through by FSupi gives the mass of UPOs as VSS in the reactor per flux of this material in the influent, viz.,
MXI Rs ¼ FSupi f cv
ðmgVSS=ðmgCOD=dÞÞ
ð68Þ
A plot of MXI/FSupi versus Rs from Equation (68) is shown in Figure 9. It can be seen that the mass of UPOs as VSS increases linearly with sludge age and that the production rate or yield coefficient of this material, given by the slope of the line, is 1/fcv ¼ 1/1.48 ¼ 0.676 mg VSS/mgCOD. In COD units, the production rate or yield of this unbiodegradable organic material fcv1/fcv ¼ 1.0 mgCOD/mgCOD. This means that all the unbiodegradable organics are conserved as VSS (or its COD equivalent) in the reactor and the only reason why the yield is 0.676 mgVSS/mgCOD is because in the influent this material is expressed in COD units, whereas in the reactor it is expressed in VSS units. If the UPOs are expressed in the same units, then the yield is 1 mgVSS/mgVSS or 1 mgCOD/mg COD. This is the result that would be obtained if a suspension of UPO is made and treated in an AS system. The influent would have a COD/VSS ratio of around 1.48 mgCOD/mgVSS (which can be measured by doing the COD and VSS tests on the influent). The mass of UPO which accumulates in the reactor as VSS (provided it all settles out in the settling tank) at steady state would be proportional to the sludge age. The effluent
438
Biological Nutrient Removal Reactor VSS mass For unbiodeg. particluate organics
Hence,
mgVSS/(mgCOD/d)
20
15
10
5
0 0
5
10 15 20 Sludge age (R s)
25
30
Figure 9 Mass of VSS in reactor per mg COD load per day applied to the reactor for unbiodegradable particulate organics and biodegradable organics.
COD concentration would be zero (no dissolved organics) and the OUR in the reactor also zero (no biodegradable organics). For example, if the suspension of UPO in water has COD and VSS concentrations of 112.5 mgCOD l1 and 76.0 mgVSS l1, respectively, that is, a COD/VSS ratio fcv ¼ 1.48 mg COD/mgVSS, and this is fed to a 15.2 l AS reactor at 15 l d1, then from Equations (66) or (67), the mass of UPOs in the reactor at steady state MXI at 20 days sludge age would be MXI ¼ 15(112.5/1.48)20 ¼ 22804 mgVSS. This mass is diluted into the 15.2 l reactor volume giving, from Equation (64), a reactor VSS concentration of XI ¼ MXI/Vp ¼ 22804/ 15.2 ¼1500 mgVSS l1. Because the steady-state mass of XI in the reactor is linear with sludge age, at 10 days sludge age, the VSS concentration would be 750 mgVSS l1.
4.14.7.1.2 Unbiodegradable soluble organics (Susi) Applying the mass-balance equation (60) to the unbiodegradable soluble organics yields
V p dSus ¼ Qi Susi dt Qe Sus dt Qw Sus dt þ 0 0
ð69Þ
where Sus is the unbiodegradable soluble organics concentration in the reactor. Additional subscripts ‘i’ or ‘e’ denote concentration in the influent or effluent, respectively. Noting that
Qe þ Qw ¼ Qi
ð70Þ
from conservation of water mass in the system, Suse ¼ Sus from complete mixing regime, Vp/Qi ¼ Rhn from Equation (59a), dSus/dt ¼ 0 at steady state, and dividing Equation (69) by Vpdt, yields
dSus =dt ¼ ðSusi Sus Þ=Rhn ¼ 0
Sus ¼ Suse ¼ Susi
ðmgCOD l 1 Þ
ðQe þ Qw ÞSuse ¼ Qi Susi FSuse ¼ FSusi
ðmgCOD d 1 Þ ðmgCOD d 1 Þ
ð71Þ
From Equation (71), at steady state the unbiodegradable soluble organics pass through the system unchanged from the influent to the effluent and waste flows, and the reactor (Sus) and effluent (Suse) concentrations are equal to that in the influent. Also, because it is assumed that the water mass is conserved, the flux of unbiodegradable soluble organics exiting the system via effluent and waste flows is equal to the flux of unbiodegradable organics entering the system via the influent flow. For example, if a 15 l d1 wastewater flow with a 52.5 mgCOD l1 unbiodegradable soluble COD concentration (Susi) were treated in a 15.2 l AS reactor at a long sludge age, then the effluent and waste flow unbiodegradable soluble COD concentrations would also be 52.5 mgCOD l1.
4.14.7.1.3 Biodegradable organics The behavior of the system with biodegradable organics is now demonstrated with a hypothetical wastewater comprising only soluble biodegradable organics. This approach is also consistent with the Monod kinetics for modeling OHO growth behavior: these kinetics are valid only for soluble RBOs (see Section 4.14.5.1.2). Once the system equations for soluble biodegradable organics have been developed, they will be modified for real wastewater comprising soluble and BPOs. When a wastewater comprising only soluble biodegradable organics (Sbsi) is treated in an AS system (1) active OHOs increase due to growth on Sbs, (2) Sbs decreases, (3) oxygen is utilized for synthesis of OHOs (Os), (4) OHOs decrease due to endogenous respiration resulting in (5) generation of endogenous residue and (6) oxygen utilization (Oe). So, the variables for which the system equations need to be developed are XBH, XEH, Sbs, and oxygen utilization (Oc ¼ Os þ Oe). These variables, together with the system design parameters (Qi, Vp, Qw, and Qe) are shown diagrammatically in the AS system in Figure 8. Applying the mass-balance to the OHO concentration XBH (mgVSS l1) assuming that: (1) the influent XBH concentration is 0; this is of course not true because there will always be some OHOs in the influent; however, relative to the mass of OHOs that grow in the reactor, the mass of OHOs that are seeded into the system is negligible; (2) the settling tank is 100% efficient resulting in no loss of OHOs with the effluent; this also is not realistic but again the mass of OHOs lost with the effluent under normal circumstances is negligible relative to the mass of OHOs intentionally harvested from the system via the waste flow; (3) the growth rate of OHOs is given by Equation (49); and (4) the loss rate of OHOs by endogenous respiration in mgVSS l1 d1 is given by Equation (53), then
Vp dXBH ¼ 0 0 Qw XBH dt mHm Sbs XBH Vp dt bH XBH V p dt þ KS þ Sbs
Biological Nutrient Removal
Dividing through by Vpdt and noting that Qw/Vp ¼ 1/Rs yields
dXBH XBH mHm Sbs ¼ þ XBH bH XBH dt Rs KS þ Sbs
ð72Þ
Because the biodegradable COD concentration in the reactor (Sbs) will be low in the completely mixed reactor, the full Monod term in Equation (72) can be replaced by the simplified first-order growth rate. Then expressing the OHO specific growth rate in terms of the maximum specific substrate utilization rate Km instead of the maximum specific growth rate mHm yields
dXBH XBH ¼ þ YH Kv Sbs XBH bH XBH dt Rs
ð73Þ
where
Kv ¼ Km =Ks ¼ mHm =ðYH Ks Þ ðl=ðmgVSS dÞ1 Þ Applying the mass-balance equation to the endogenous residue concentration (XEH) and assuming, as for the OHOs, no endogenous residue in the influent, a 100% efficient settling tank, and the rate of XEH generation in mgVSS l d1 given by Equation (54), we obtain
Vp dXEH ¼ 0 0 Qw XEH dt þ f EH bH XBH V p dt Dividing through by Vpdt and noting that Qw/Vp ¼ 1/Rs yields
dXEH XEH ¼ þ f EH bH XBH dt Rs
ðmgVSS l1 d1 Þ
ð74Þ
Applying the mass balance to the soluble biodegradable COD concentration and noting that because of complete mixing conditions in the reactor, the soluble Sbs concentration in the effluent and waste flow is the same as that in the reactor, that is, Sbse ¼ Sbsw ¼ Sbs , yields,
Vp dSbs ¼ Qi Sbsi dt Qw Sbs dt Qw Sbs dt Qe Sbs dt mHm Sbs XBH Vp dt mgCODl1 d1 þ0 YH KS þ Sbs Dividing through by Vpdt and noting that Qi/Vp ¼ 1/Rhn yields
dSbs ðSbsi Sbs Þ mHm Sbs ¼ XBH Rhn dt YH KS þ Sbs
ð75Þ
Again, as was done in Equation (73) for XBH above, substituting the simplified first-order equation for the full Monod equation yields
dSbs ðSbsi Sbs Þ ¼ Kv Sbs XBH Rhn dt
ð76Þ
Equations (72), (74), and (75) are valid differential equations for the active OHO (XBH), endogenous residue (XEH), and soluble biodegradable organic (Sbs) concentrations in the completely mixed AS system. Because these equations include the full Monod equation, simple solutions for XBH, XEH, and Sbs cannot be found. Therefore, to generate solutions for XBH, XEH, and Sbs, these three equations need to be simultaneously integrated numerically for which a computer program or simulation package is required. The solutions so generated will be
439
realistic provided realistic values for the OHO kinetic constants for the specific soluble biodegradable organics (YH, bH, fEH, mHm, and KS) are given as input. When this is done, it will be noticed that after an initial start-up period, which can be several sludge ages long, the solutions for XBH, XEH, and Sbs will reach a constant concentration, if the influent flow and Sbs load on the reactor are constant with time. During the initial start-up period, the transients dXBH/dt, dXEH/dt, and dSbs/dt become smaller and smaller until they reach zero at which time steady state is reached for the constant flow and organic load conditions. If the flow and organic load are not constant but vary cyclically over the day in an identical cyclic pattern for each successive day, then after an initial start-up period, a dynamic steady state is reached. Under dynamic steady-state conditions, XBH, XEH, and Sbs also will vary cyclically over the day in response to the cyclic flow and organic load over the day, and this variation will be identical for each successive day. This variation in XBH, XEH, and Sbs will be around an average XBH, XEH, and Sbs concentrations and these average concentrations will be virtually the same as those achieved under steady-state conditions if the organic load on the system over the day under constant flow and load conditions is the same as for the cyclic flow and load conditions. Mathematically, the cyclic loading average XBH, XEH, and Sbs concentrations are called the particular integral and the daily variation around these averages is called the complementary function (Figure 10). If the flow and organic load over the day are constant, then at steady state there is only a particular integral; the complementary function is zero because there is no variation in XBH, XEH, and Sbs concentrations in the reactor over the day. For the AS model, the particular integral (i.e., the steadystate or cyclic loading average concentrations) is very important because they give a close approximation of the concentrations in the reactor in response to the daily average organic load on the reactor. While Equations (72) and (75) are only valid for soluble biodegradable organics fed to the completely mixed AS system, and therefore will not generate valid numerical solutions for real wastewater comprising soluble and particulate organics (see Sections 4.14.5.1.2 and 4.14.5.1.3), the conclusion regarding the importance of the particular integral or steady-state solution is valid and also applies to the system treating real wastewater. The problem is finding simple equations for these steady-state concentrations for the real wastewater system. This is achieved first by substituting the simplified first-order Monod equation into Equation (72) for XBH and Equation (75) for Sbs to give Equations (73) and (76), respectively, and second by making the reasonable assumption (discussed below) that all the biodegradable organics are broken down in the system. Setting the transients in Equations (73), (74), and (76) to 0 yields
0 ¼ YH Kv Sbs XBH bH XBH XBH =Rs from which
Sbs ¼
ð1 þ bH Rs Þ ðmgCOD l1 Þ Kv YH Rs
ð77Þ
which is the steady-state solution for the soluble biodegradable COD concentration in the reactor (Sbs, mgCOD l1). Note that the Sbs concentration is independent of the influent
440
Biological Nutrient Removal
Start-up period
Steady state
Concentration (mg I−1)
Cyclic flow and load response Steady state Particular integral Transient zero Steady-state solution
Constant flow and load response
Dynamic steady state Complementary function Transient not zero
Initial condition
Time (days) Figure 10 Calculated concentration vs. time response profile from a selected initial condition for constant and cyclic flow and load conditions showing changing concentration during start-up and constant or cyclically varying concentration when steady state is reached.
organic concentration (Sbsi). This implies that the steady-state residual Sbs concentration is the same irrespective of the magnitude of the influent Sbsi concentration. Note also that Sbs is dependent on sludge age (Rs) and not on the nominal hydraulic retention time (Rhn). From Equation (74)
0 ¼ f EH bH XBH XEH =Rs from which in turn
XEH ¼ f EH bH Rs XBH
ðmgVSS l1 Þ
ð78Þ
which is the steady-state solution for the reactor endogenous residue concentration (XEH, mgVSS l1). From Equation (76)
0 ¼ ðSbsi Sbs Þ=Rhn Kv Sbs XBH from which in turn
Kv Sbs ¼ ðSbsi Sbs Þ=ðRhn XBH Þ
ð79Þ
A simple equation for the steady-state OHO concentration cannot be found from Equation (79) alone. Multiplying Equation (77) through by Kv makes the right-hand side equal to KvSbs, which is the same as the right-hand side of Equation (79). Setting the right-hand side of these two equations equal and solving for XBH yield
XBH ¼
YH Rs ðSbsi Sbs Þ ðmgVSS l1 Þ Rhn ð1 þ bH Rs Þ
ð80Þ
which is the steady-state solution for the active OHO concentration in the reactor in mgVSS l1. If Sbs is solved for from Equation (79), it will be found that
Sbs ¼
Sbsi ðmgCOD l1 Þ ð1 þ Kv XBH Rhn Þ
ð81Þ
This equation is similar to that used for fecal coliform die-off in oxidation ponds (Marais, 1974) except that the die-off rate is KvXBH. It appears to indicate that the residual soluble biodegradable organic concentration Sbs is a function of the influent Sbsi concentration and dependent on the nominal hydraulic retention time (Rhn) and not on the sludge age (Rs) in apparent contradiction to Equation (77). However, Sbs does not increase with increase in Sbsi because the effective utilization rate KvXBH is not constant; the higher the Sbsi, the higher the XBH and the higher the utilization rate KvXBH. This increased KvXBH rate compensates for the increased Sbsi so that the same Sbs concentration is obtained. This equation for Sbs is not useful because it requires the (1) active OHO concentration XBH and (2) nominal hydraulic retention time Rhn to be known, which at the design stage are not. Because the influent wastewater comprises only soluble biodegradable organics (Sbsi), the VSSs concentration in the reactor (Xv) is the sum of the OHO and endogenous residue concentrations, viz.,
Xv ¼ XBH þ XEH ðmgVSS l1 reactor volumeÞ ¼ XBH þ f EH bH Rs XBH ¼ XBH ð1 þ f EH bH Rs Þ YH Rs ðSbsi Sbs Þ ð1 þ f EH bH Rs Þ ¼ ð1 þ bH Rs Þ Rhn
ð82Þ
As in Equation (63) for the UPOs accumulating in the reactor from the influent (XI), the appearance of Rhn in Equation (82) is not helpful in design because it requires the reactor volume (Vp) to be known. Therfore, Equation (82) is converted to mass of VSS in the reactor (MXv) by substituting Vp/Qi for Rhn and multiplying through by Vp yielding
Vp Xv ¼ Qi ðSbsi Sbs Þ
YH Rs ð1 þ f EH bH Rs Þ ð1 þ bH Rs Þ
Now VpXv in the equation is the mass of VSSs in the reactor which from the earlier mass nomenclature is MXv mgVSS. The (Sbsi Sbs) is the biodegradable COD concentration change between the influent and effluent denoted DSbs mgCOD l1.
Biological Nutrient Removal
Multiplying this concentration change by the influent flow gives the flux of soluble biodegradable organics degraded per day in the system and is denoted FDSbs mgCOD d1. Hence, at steady state, the mass of VSS in the reactor is given by
MXv ¼ FDSbs
YH Rs ð1 þ f EH bH Rs Þ ð1 þ bH Rs Þ
MXBH
YH Rs ðmgVSSÞ ¼ FDSbs ð1 þ bH Rs Þ
MXEH ¼ f EH bH Rs MXBH
ðmgVSSÞ
soluble biodegradable organics across the system (FDSbs) is equal to the flux (load) of these organics on the system (FSbsi), viz.,
FDSbs ¼ Qi Sbsi ðQw þ Qe ÞSbs ¼ Qi Sbsi Qi Sbse ¼ Qi ðSbsi Sbse Þ
ð83Þ
from which it can be seen that the mass of VSS in the reactor MXv when treating a wastewater with only soluble biodegradable organics is proportional to the flux of biodegradable COD degraded FDSbs. Converting the reactor concentrations XBH and XEH to masses in the reactor by noting that Sbsi Sbs ¼ DSbs, Rhn ¼ Vp/ Qi, VpXBH ¼ MXBH, QiDSbs ¼ FDSbs, and VpXEH ¼ MXEH yields
¼ Qi Sbsi ¼ FSbsi Substituting FSbsi for FDSbs into Equations (84), (85), and (83) yields
MXBH ¼ FSbsi ð84Þ
To determine the masses of OHO VSS (MXBH), endogenous residue VSS (MXEH), and total VSS in the reactor (MXv) from Equations (83)–(85) for the soluble biodegradable organics in wastewater, the flux change (mgCOD d1) in soluble biodegradable organics (FDSbs) needs to be known, which in turn requires the residual soluble biodegradable organics concentration not degraded in the reactor (Sbs ¼ Sbse) to be known. The residual Sbs concentration can be calculated from Equation (77) noting that Kv ¼ mHm/(KsYH) from Equation (73). Assigning appropriate values to mHm, Ks, and YH allows one to determine Kv. For soluble RBOs such as glucose and acetate, typically mHm is between 1 and 5 d1 and Ks between 5 and 20 mgCOD l1 for OHOs isolated from AS (Richard et al., 1982). Accepting mHm ¼ 2.25 d1, Ks ¼ 10 mgCOD l1, and YH ¼ 0.45 mgVSS/mgCOD yields Kv ¼ 0.5 l/(mgOHOVSS d). Accepting bH ¼ 0.24 d1 at 20 1C allows the residual unbiodegraded Sbs concentration to be calculated for a selected sludge age from Equation (77). When this is done, it will be noticed that even at very short sludge ages (B3 d), Sbs is very low at around 2 mgCOD l1. If the influent soluble RBO concentration (Sbsi) is higher than 200 mgCOD l1, which are common soluble biodegradable organic concentrations in real wastewater, then in excess of 99% biodegradable organic material breakdown takes place in the system. Therefore, for soluble biodegradable organics, it is reasonable to assume that all the organics are degraded and that the change in flux of
YH Rs ðmgVSSÞ ð1 þ bH Rs Þ
MXEH ¼ f EH bH Rs MXBH YH Rs f EH bH Rs ¼ FSbsi ð1 þ bH Rs Þ
ð85Þ
4.14.7.1.4 Complete utilization of soluble biodegradable organics
441
MXv ¼ FSbsi
ð86Þ
ðmgVSSÞ
ð87Þ
YH Rs ð1 þ f EH bH Rs Þ ðmgVSSÞ ð1 þ bH Rs Þ
ð88Þ
Knowing the OHO kinetic and stoichiometric constants (see Table 6) allows the mass of VSS in the reactor per mass of biodegradable COD load per day, that is, MXv/FSbsi to be calculated and plotted versus sludge age (Figure 11). From Figure 11, it can be seen that the OHO VSS mass in the reactor increases from about 1.0 to 1.5 kgVSS/(kgCOD/d) with increasing sludge age from 5 to 15 days. Increases in sludge age above 15 days lead to only a marginal increase in OHO VSS mass, that is, 1.6 kgVSS/(kgCOD/d) at 30 days sludge age. In contrast, the endogenous residue VSS mass is very low at short sludge ages and increases almost linearly with sludge age from 0.2 at 5 days to 2.0 kgVSS/(kgCOD/d) at 30 days sludge age. At 25 days sludge age the OHO and endogenous residue VSS masses are equal at 1.6 kgVSS/(kgCOD/d) giving a total VSS mass in the reactor of 3.2 kgVSS/(kgCOD/d). Figure 11 also shows that the proportion of the OHOs in the VSS mass in the reactor decreases as sludge age increases; this proportion is 80% and 50% at 5 and 25 days sludge age, respectively (see Section 4.14.9.5). A comparison of the mass of VSS in the reactor per flux organic load on the reactor (MXv/FSbsi) from biodegradable soluble organics (which are accepted to be all utilized) with that from UPOs is given in Figure 12. It can be seen that much less VSS mass accumulates in the reactor with biodegradable organics (YH ¼ 0.45 mgVSS/mgCOD, bH ¼ 0.24 d1) than
Table 6 Stoichiometric and kinetic constants and their temperature dependency for the OHOs in the steady-state carbonaceous degradation activated sludge model Constant
Symbol
Temperature dependence
Theta (y)
Standard value at 20 1 C
Yield coefficient (mgVSS/mgCOD) Endogenous respiration rate (d1) Endogenous residue fraction (–) COD/VSS ratio (mgCOD/mgVSS)
YH bH fEH fcv
Remains constant bHT ¼ bH20 y (T20) Remains constant Remains constant
1 1.029 1 1
0.45 0.24 0.2 1.48
From Marais GvR and Ekama GA (1976) The activated sludge process part 1 – steady state behaviour. Water SA 2(4) 163–200.
Eq. no.
56
442
Biological Nutrient Removal Reactor VSS mass for biodegradable organics only
4
mgVSS/(mgCOD/d)
Volatile mass, MXV 3
Active OHO mass, MXBH
2
1 Endogenous residue mass, MXEH 0 0
5
10 15 20 Sludge age (R s)
25
30
Figure 11 Active OHO (MXBH), endogenous residue (MXEH), and volatile suspended solids (MXv ¼ MXBH þ MXEH) per mass biodegradable COD per day (MXv/FSbsi) vs. sludge age for wastewater with biodegradable organics only.
Reactor VSS mass for biodeg. and unbiodeg. organics
20
mgVSS/(mgCOD/d)
From unbiodegradable 15
reaction and the carbon (energy) of these organics catabolized exit the system as CO2 (is lost as heat). Only twothirds of the soluble biodegradable organic electrons accumulate in the reactor as particulate OHO organic (VSS) material. 2. If the endogenous respiration rate were zero (bH ¼ 0.0 d1), then according to Equation (86) the mass of OHO per organic load per day (MXBH/FSbsi) would increase linearly with sludge age at a slope of YH ¼ 0.45 mgVSS/mgCOD (see Figure 12, the YH ¼ 0.45 mgVSS/mgCOD, bH ¼ 0.0 line) and VSS mass would be two-thirds of that for UPOs (see (1) above). However, bH ¼ 0.24 d 1 which means that for every day of sludge age, 24% of OHO VSS organics is lost, of which 20% (fEH ¼ 0.20) accumulates as unbiodegradable particulate endogenous residue and increases the VSS mass in the reactor. The electrons of the remaining 80% biodegradable OHO organics VSS lost every day (1 fEH ¼ 0.8) are passed to the oxygen (endogenous respiration oxygen utilization, Oe), the carbon of which exits the system as CO2 and the energy of which is lost as heat. From the above it is clear that the reduced VSS mass in the reactor per organics COD load per day with biodegradable organics is due to oxygen utilization for synthesis and endogenous respiration which are the result of the biological redox reactions in which the electrons, carbon, and energy of biodegradable organics from the influent and OHOs are passed to oxygen, released as CO2 and lost as heat, respectively. The difference in terms of the VSS mass in the reactor per organic load per day between unbiodegradable particulate and soluble biodegradable organics is therefore due to oxygen utilization which conforms to the COD (e) mass balance.
4.14.7.1.5 The mass balance on oxygen Applying the general mass balance (Equation (60)) to the oxygen in the system yields
COD lost via oxygen utilization = energy lost as heat
10
5
3 3 2 3 2 3 2 Mass Mass Mass Mass O2 6 change 7 6 flow O 7 6 flow O 7 6 utilized 7 27 27 6 6 6 7 7 6 6 7 7¼6 76 76 4 in O2 in 5 4 into 5 4 out of 5 4 by OHOs in 5 system system system system 2
From biodegradable 0 0
5
10
15
20
25
30
Vp dOr ¼ FOin dt dOr ðQw þ Qe Þdt OURc Vp dt
Sludge age (R s) Figure 12 Mass VSS (MXv) in reactor per mass COD load /d on the reactor vs. sludge age for unbiodegradable particulate organics and biodegradable organics. The difference is the COD (or e) passed to oxygen which is proportional to the energy in the biodegradable organics lost as heat.
with UPos (YH ¼ 0.676 mgVSS/mgCOD, bH ¼ 0.0 d1) – from about 65% less at 5 days sludge age to 81% less at 25 days sludge age. There are two reasons for this: 1. The yield of OHO VSS electrons is 0.45 mgVSS/mgCOD (or 0.45 1.48 ¼ 0.66 mgCOD/mgCOD) so that one-third of the electrons in the organics (1 fcvYH ¼ 1 0.66 ¼ 0.34) are passed to oxygen (synthesis oxygen utilization, Os) to form water in the OHO catabolic part of the synthesis
where dOr is the dissolved oxygen concentration in the reactor (mgO l1) and, OURc the OHO oxygen utilization rate (mgO l1d1). Normally, the mass of oxygen exiting the reactor via the effluent and waste flows is negligible in comparison with that entering the reactor by the aeration device (FOin, mgO d1). Also even when the reactor DO concentration is changing so that the transient dOr/dt is not zero, this too has a negligible effect in the oxygen mass balance compared with the mass of oxygen transferred to the water by the aeration device and that utilized by the OHOs (OURcVp dt). Hence, the rate of oxygen mass input by the aeration device is equal to the mass rate of oxygen utilized by the OHOs, viz.,
FOin ¼ OURc Vp ¼ FOc
ðmgO d1 Þ
ð89Þ
Biological Nutrient Removal
Now OURc is the OHO OUR, which is the sum of the OURs for synthesis (OURs) and endogenous respiration (OURe):
OURc ¼ OURs þ OURe
1
ðmgO l
substituting Equation (84) for MXBH into Equation (91) yields
2
3
6 FOc ¼ FDSbs 4ð1 f cv YH Þ þ f cv ð1 f EH ÞbH
1
d Þ
Synthesis
ðmgO d Þ
Now from the biological kinetic behavior of growth (Equation (51))
OURs ¼ ð1 f cv YH Þ
dSbs Kms Sbs ¼ ð1 f cv YH Þ XBH dt Ks þ Sbs
ðmgO l1 h1 in reactorÞ
OURs ¼ ð1 f cv YH ÞKv Sbs XBH
and accepting that all the soluble biodegradable organics are degraded, that is, FDSbs ¼ FSbsi yields
FOc ¼ FSbsi ð1 f cv YH Þ þ f cv ð1 f EH ÞbH
YH Rs ð1 þ bH Rs Þ
ðmgO d1 Þ
which for the simplified first-order conditions reduces to
From Equation (79),
Kv Sbs XBH ¼ ðSbsi Sbs Þ=Rhn ¼ Qi ðSbsi Sbs Þ=Vp Therefore,
OURs ¼ ð1 f cv YH ÞQi ðSbsi Sbs Þ=Vp This makes sense from our kinetic understanding, that is, oxygen for synthesis is the catabolic part (1 fcvYH) of the organics degraded. The oxygen utilized for endogenous respiration comes from Equation (55):
OURe ¼ f cv ð1 f EH ÞbH XBH
YH Rs 7 5 ð1 þ bH Rs Þ
Endog: resp:
1
FOc ¼ FOs þ FOe
443
ðmgO l1 d1 in reactorÞ
that is, COD equivalent of the biodegradable OHO VSS concentration that disappears. Hence,
ð92Þ
From Equation (92), it can be seen that the mass of oxygen utilized by the OHOs (FOc) is a function of the OHO stoichiometric and kinetic constants and the sludge age. Knowing the values of the stoichiometric and kinetic constants from Table 6, the mass of oxygen utilized by the OHOs per mass organic load per day [FOc/FSbsi, (mgO/d)/(mgCOD/d)] is plotted versus sludge age (Rs) as shown in Figure 13. From Figure 13, it can be seen that as the sludge age increases so the flux of oxygen utilized (total demand, FOc) per flux organic load increases, but the increase becomes smaller as the sludge age increases. The synthesis oxygen demand (FOs) is constant because all the biodegradable organics are utilized and transformed to OHO VSS mass. The increase in total oxygen demand is due the increasing oxygen demand from endogenous respiration (FOe), which increases because the longer the OHO VSS mass remains in the reactor, the more of this mass is degraded via endogenous respiration, the electrons, carbon, and energy of which are passed to oxygen, changed to CO2, and lost as heat, respectively. Clearly, synthesis is the biological process whereby influent biodegradable organics are transformed to OHO VSS mass (anabolism) with OHO oxygen demand
OURc ¼ ð1 f cv YH ÞQi ðSbsi Sbs Þ=Vp þ f cv ð1 f EH ÞbH XBH
kgO/d per kgCOD/d load on reactor
1.0
ðmgO l1 h1 in reactorÞ ð90Þ
Total FOc
To obtain the flux of oxygen utilized FOc, the rate per liter reactor is multiplied by Vp, that is,
FOc ¼ OURc Vp ¼ OURs Vp þ OURe Vp ¼ ð1 f cv YH ÞQi ðSbsi Sbs Þ þ f cv ð1 f EH ÞbH XBH VP which in terms of the mass and flux nomenclature yields
FOc ¼ ð1 f cv YH ÞFDSbs þ f cv ð1 f EH ÞbH MXBH
kgO/d per kgCOD/d
0.8
Endogenous respiration FOe
0.6
0.4
0.2 1
ðmgO d Þ
Synthesis FOs
ð91Þ
From Equation (91) it can be seen that the synthesis OUR is proportional to the mass of biodegradable organics degraded per day and the endogenous respiration OUR proportional to the mass of active OHOs in the reactor. From Equation (84), the mass of active OHOs in the reactor is related to the flux of biodegradable organics degraded, so
0.0 0
5
20 10 15 Sludge age (R s)
25
30
Figure 13 The synthesis, endogenous respiration, and total OHO oxygen utilization rates (demand) versus sludge age in kgO/d per kg COD/d biodegradable organic load on the reactor.
444
Biological Nutrient Removal
an associated electron transfer to oxygen and an energy loss as heat (catabolism), and endogenous respiration is a process whereby now the organism biodegradable organics are degraded via catabolism to CO2 with a further oxygen demand and energy loss as heat. The electron transfer to oxygen results in the much lower accumulation of VSS mass in the reactor compared with UPOs (Figure 12). In Figure 12, the difference between the YH ¼ 0.676 mgVSS/mgCOD and the YH ¼ 0.45 mgVSS/mgCOD bH ¼ 0.0 d1 line values at a selected sludge age (say 10 days) is the accumulated oxygen consumed for synthesis over the 10 days sludge age period. Similarly, the difference between the YH ¼ 0.45 mgVSS/mgCOD bH ¼ 0.0 d1 and the YH ¼ 0.45 mgVSS/mgCOD and bH ¼ 0.24 d1 line values in Figure 12 at a selected sludge age (say 20 days) is the accumulated oxygen consumed for endogenous respiration over 20 days sludge age period. From this it can be seen that (1) the YH value governs the proportion of the influent biodegradable organic COD (e) that is conserved as an OHO organics (balance passed the oxygen for synthesis) and (2) the bH rate governs the rate of breakdown of the organism biodegradable organics, in which the COD (e) of these organism organics is passed to the oxygen in the endogenous respiration process. At sludge ages of 5, 10, and 25 days, of the influent biodegradable organic COD, 38%, 29%, and 21% remains as sludge solids COD and 62%, 71%, and 79% is oxygen utilized, which represents energy lost as heat. It can be seen that even at short sludge ages, the greater proportion of the influent biodegradable organic COD (e), carbon, and energy is passed oxygen, transformed to CO2 and lost as heat, respectively, in the system.
4.14.7.1.6 Complete utilization of BPOs The influent BPOs, both settleable and suspended (Sbpi), are mostly slowly biodegradable. These slowly BPO become enmeshed and absorbed within the AS flocs and become part of suspended VSS sludge mass in the reactor (see Section 4.14.5.1.3). As part of the VSS sludge mass, these organics settle out with the sludge mass in the setting tank and are returned to the biological reactor. Undegraded particulate organics therefore do not escape with the effluent but remain part of the sludge VSS mass in the system; the only exit route for the undegraded BPOs is via the waste flow (Qw) with the waste sludge. The time available for the breakdown of the particulate slowly biodegradable organics by the OHOs is therefore related to the solids retention time or sludge age of the system. It has been found from experimental and modeling work with real wastewater that if the sludge age is long (43 days for fully aerobic systems at 20 1C), then virtually complete breakdown of the BPOs takes place in the system. Hence, at long sludge ages the residual biodegradable organic concentration, both soluble and particulate, not broken down can be accepted to be very small. From this an important assumption and simplification can be made for the steady-state model, that is, it is not necessary to make a distinction between soluble and BPOs; all are transformed to OHO VSS mass. However, it must be remembered that, although reasonable, this assumption that all the biodegradable organics are
degraded cannot be proved because any residual biodegradable soluble and particulate organics not degraded in the system are implicitly included with the unbiodegradable soluble and particulate organic fractions, respectively.
4.14.7.1.7 Integration of biodegradable and unbiodegradable organics models In Sections 4.14.7.1.1–4.14.7.1.3, the behavior of the four organic constituents of real wastewater in the completely mixed system was examined individually with hypothetical wastewaters containing only the single constituent: 1. 2. 3. 4.
UPOs (Supi, mgCOD l1), unbiodegradable soluble organics (Susi, mgCOD l1), biodegradable particulate organics (Sbpi, mgCOD l1), and biodegradable soluble organics (Sbsi, mgCOD l1).
However, because the biodegradable organics are accepted to be completely utilized at sludge ages longer than about 3 days, these organics can be considered a single constituent as all are transformed to OHO mass through the biological growth process. To obtain the steady-state AS model for the completely mixed system treating the real wastewater, the equations derived for each individual constituents are simply combined. Combining the three constituents into the 15 l d1 wastewater flow (Qi) and treating this in the 15.2 l (Vp) 20 day sludge age (Rs) in a system result in a reactor VSS concentration (Xv) of 3256 mgVSS l1, an oxygen utilization rate (FOc) of 6798 mgO d1 or (Oc) 18.63 mgO l1 h1 and an effluent COD concentration (Suse) of 52.5 mgCOD l1. In the same way as here the answers from the steady-state equations are added; below (Section 4.14.9.3) the equations themselves are added to give the complete steady-state model for real wastewater COD removal. The major conclusion from the above is that if a completely mixed aerobic AS system were operated by hydraulic control at a long sludge age (say 20 days) and fed unknown real wastewater at a constant flow and load for a sufficiently long time to reach steady-state conditions, then by measuring the effluent COD concentration (Suse), the reactor VSS concentration (Xv), and the oxygen utilization rate (OURc or FOc), the wastewater unbiodegradable soluble and particulate organic concentrations (Susi and Supi) or fractions (fS’us and fS’up) can be calculated. In such an experiment, it would be sufficient to measure the effluent COD and the reactor VSS concentrations only to determine Supi. However, it is also important to measure the oxygen utilization rate to allow the COD balance of the experimental results to be checked. The closer the COD balance is to 100%, the more reliable the Supi estimate. Despite considerable research efforts to find simpler and less laborious methods of wastewater Supi or fS’up measurement (Wentzel et al., 1999; Torrijos et al., 1994), the only reliable method to date is that outlined above (Ekama et al., 1986).
4.14.8 The COD (or e) Mass Balance In the AS system, COD theoretically must be conserved so that at steady state the COD flux out of the system must be equal to the COD flux into the system over a defined time interval. The
Biological Nutrient Removal COD (e) of the influent organics is (1) retained in the unbiodegradable particulate and soluble organics, (2) transformed to OHO mass and therefore conserved in a different type of organic material, or (3) passed onto oxygen to form water. Therefore, in general, the COD (or e) balance over the AS system at steady state is given by
"
can be seen that at all sludge ages the influent COD flux is the same and constant (equal to 11 250 and 6750 kgCOD for raw and settled wastewater respectively). Also, the effluent COD flux is the same for the raw and settled wastewater (788 kgCOD d1) because the effluent COD concentration is the same at the USO concentration of 53 mgCOD l1. The flux
# " # Mass of COD ðe Þ Mass of COD ðe Þ ¼ output per day input per day 3 3 2 3 2 2 Mass of Mass of Mass of 7 2 Mass of oxygen utilized 3 2 Mass of 3 7 6 7 6 6 6 soluble 7 6 soluble 7 6 particulate 7 7 6 7 6 7 6 6 7 6 7 7 7 6 7 6 6 6 COD in 7 þ 6 COD in 7 þ 6 COD in 7 þ4 by OHOs for COD 5 ¼ 4 COD input 5 7 7 6 7 6 6 breakdown per day per day 4 effluent 5 4 waste flow 5 4 waste flow 5 per day
per day
445
ð93Þ
per day
Qe Ste þ Qw Ste þ Qw Xv fcv þ Vp OURc ¼ Qi Sti
where Ste is the effluent total soluble COD concentration (mgCOD l1), Xv the VSS concentration in biological reactor (mgVSS l1), and OURc the carbonaceous (for organic material degradation) oxygen utilization rate in reactor (mgO l1 h1). In Equation (93), the first two terms represent the soluble organics that exit the system via effluent and waste flows, the third term the particulate organics that exits the system via the waste flow, and the fourth term represents the mass of oxygen utilized for biodegradable organic material breakdown by the OHOs. Noting that from Equations (70), (58), (89), and (65)
of COD exiting the system via the waste VSS (FXv) decreases as sludge age increases and concomitantly the flux of oxygen utilized (FOc) increases as sludge age increases so that their sum is constant all sludge ages (equal to 11 250 788 ¼10 462 and 6750 788 ¼ 5962 kgCOD d1 for raw and settled wastewater respectively). This transfer from VSS (COD) to oxygen takes place via the endogenous respiration process and because the endogenous respiration process continues for longer, the longer the sludge age, more COD (electrons) is transferred from the VSS sludge mass to oxygen the longer the sludge age.
ðQe þ Qw ÞSte ¼ Qi Ste ¼ FSte
Example wastewaters 12 000
Qw Xv ¼ Vp Xv =Rs ¼ MXv =Rs ¼ FXv 10 000
Qi Sti ¼ FSti the general COD mass balance is given by
ð94Þ
where FSte is the COD flux of soluble organics exiting system via effluent and waste flows (mgCOD d1), fcvMXv/Rs ¼ FXv the COD flux of particulate organics (VSS) exiting system via waste flow (mgCOD d1), and FOc flux of oxygen utilized by OHOs for biodegradable organic material breakdown (carbonaceous) (mgO d1). The carbonaceous oxygen utilization flux FOc is the sum of the oxygen utilized for the synthesis of new organism mass from the biodegradable organics (FOs) and the oxygen utilized for internal generation of energy for essential cell functions from organism biodegradable organics (endogenous respiration (FOe), Equation (91)). The COD balance (Equation 94) is shown for the example raw and settled wastewater in Figure 14 versus sludge age. It
COD flux out (kgCOD d−1)
Vp OURc ¼ FOc
FSte þ f cv MXv =Rs þ FOc ¼ FSti
Raw FSti 22 C 14 C
8000
Settled FSti
FOc Raw
6000
4000
Settled FOc
2000
FXv Raw
Settled FXv Settled FSte
0 0
5
10
Raw 15
20
25
30
Sludge age (days) Figure 14 COD fluxes (kgCOD/d) exiting the fully aerobic activated sludge reactor via effluent (FSte), waste sludge (fcv FXv) and oxygen utilization (FOc) versus sludge age for the example raw and settled wastewaters.
446
Biological Nutrient Removal
The COD mass balance is a very powerful tool for checking not only the data measured on experimental systems (Ekama et al., 1986) but also the results calculated for design from the steady-state model.
In most cases, the effluent VSS and TSS concentrations are too low to measure reliably with the VSS and TSS tests. Alternative methods for measuring low solids concentrations in the effluent have been developed; for the VSS, via the filtered and unfiltered COD concentrations, that is, from Equation (95b).
4.14.9 The AS System Steady-State Equations for Real Wastewater Once it is recognized that all the organics in the influent, except the unbiodegradable soluble COD, either is utilized by the microorganisms to form OHO mass, or remains in the process and accumulates as inert sludge mass, it follows that the mass of sludge produced and the carbonaceous oxygen demand in the system are stoichiometric functions of the flux of COD to be treated daily; the greater the daily COD flux (FSti), the greater the sludge production (FXv) and carbonaceous oxygen demand (FOc). The steady state model equations below give the effluent COD concentration comprising the unbiodegradable soluble organics, the masses of sludge generated in the reactor, and the average daily oxygen demand for organic material removal as a function of the total organic (COD) load per day, the wastewater characteristics, that is, the unbiodegradable soluble and particulate COD fractions (fS’us and fS’up) and the sludge age. The kinetic and stoichiometric constants in the equations, that is, the specific yield coefficient (YH), the specific endogenous mass loss rate (bH), the unbiodegradable fraction of the OHOs (fEH), and the COD/VSS ratio of the sludge (fcv), as well as their temperature dependencies are given in Table 6.
4.14.9.1 Effluent COD Concentration Under normal AS process operating conditions, where the sludge ages are in excess of 8 days (to ensure nitrification and biological nutrient removal), the nature of the influent organics in municipal wastewaters is such that the COD concentration in the effluent is inconsequential in the process design – the soluble readily biodegradable COD fraction is completely utilized in a very short period of time (o1 h) and the particulate COD, whether biodegradable or unbiodegradable, is adsorbed or enmeshed in the sludge flocs, and settles out with the sludge in the secondary settling tanks. Consequently, the effluent COD concentration is comprised virtually wholly of the unbiodegradable soluble COD (from the influent) plus the COD of the sludge particles which escape with the effluent due to imperfect operation of the secondary settling tank. Hence, the effluent COD concentration, Ste, is approximately given by
Ste ¼ Suse
ðmgCOD l1 Þ
ð95aÞ
for filtered samples, or,
Ste ¼ Suse þ f cv Xve
ðmgCOD l1 Þ
ð95bÞ
for unfiltered samples, where Suse ¼ unbiodegradable COD in the effluent ¼ Susi ¼ fS’us Sti (mgCOD l1) (see Equation (3)); Xve ¼ volatile solids concentration in the effluent (mgVSS l1); and fcv ¼ COD/VSS ratio of the volatile solids ¼ 1.48 mgCOD/ mgVSS.
Xve ¼ ðunfiltered Ste filtered Ste Þ=f cv
ð96Þ
and, for TSS, via the turbidity, once a turbidity versus TSS concentration calibration curve for the AS has been prepared (Wahlberg et al., 1994).
4.14.9.2 ISS Concentration The ISS concentration from the influent accumulates in the reactor in the identical way as the UPOs Equation (66), that is, the mass of influent ISS in the reactor is equal to the daily flux of ISS into the reactor FXIOi times the sludge age (Rs), viz.,
MXIO ¼ FXIOi Rs
ðmgISSÞ
ð97aÞ
ðmgISS d1 Þ
ð97bÞ
where
FXIOi ¼ XIOi Qi
and XIOi ¼ influent ISS concentration (mgISS l1). The influent ISS is only part of the ISS measured in the reactor. The OHOs (and PAOs if present) also contribute to this concentration. For fully aerobic and ND systems, where only OHOs comprise the active biomass, the OHOs contribute about 15% of their OHOVSS mass to the ISS (Ekama and Wentzel, 2004). It appears that this ISS mass consists of intracellular dissolved solids, which, when a sludge sample is dried in the TSS procedure, precipitate as ISS. If this is so, then theoretically, this TSS contribution of the OHOs (and PAOs if present) should strictly be ignored even though it manifests in the TSS test, because being intracellular dissolved solids, it does not add to the actual ISS load on the secondary settling tank. However, because this ISS mass has always been implicitly included in the TSS test result in the past, it will be retained because SST design procedures have been based on the measured TSS result. Including the OHO ISS mass yields for non-BEPR systems,
MXIO ¼ MXIOi Rs þ f iOHO MXBH
ðmgISSÞ
ð98aÞ
¼ MXIOi Rs þ f iOHO f avOHO MXv
ðmgISSÞ
ð98bÞ
where favOHO is the fraction of the VSS mass as active OHOs (see Section 4.14.9.4) and fiOHO the inoganic content of the OHOs ( ¼ 0.15 mgISS/mgOHOVSS). For BEPR systems, the ISS in the PAOs needs to be included also. For aerobic P uptake BEPR, fiPAO ¼ 1.30 mgISS/ mgPAOVSS, that is, 9 times higher than for OHOs. For anoxic P uptake BEPR, the P content (fXBGP), and hence the ISS content (fiPAO), of the PAOs is lower and variable. Calculating the reactor ISS concentration for NDBEPR systems with PAOs is presented in Section 4.14.31.4 later.
Biological Nutrient Removal 4.14.9.3 Process Design Equations 4.14.9.3.1 For the influent The flux of total organics (FSti, mgCOD d1), biodegradable organics (FSbi, mgCOD d1), and unbiodegradable particulate organics (FXIi, mgVSS d1) are
FSti ¼ Qi Sti
ð99Þ
FSbi ¼ Qi Sbi
ð100aÞ
¼ Qi Sti ð1 f S0us f S0up Þ ð100bÞ ¼ FSti ð1 f S0us f S0up Þ FXli ¼ Qi Xli
ð100cÞ ð101aÞ
¼ Qi f S0up Sti =f cv
ð101bÞ
¼ FSti f S0up =f cv
ð101cÞ
MXv ¼ MXBH þ MXEH þ MXI YH Rs ð1 þ f EH bH Rs Þ þ FXLi Rs ¼ FSbi ð1 þ bH Rs Þ ð1 f S0us f S0up ÞYH Rs f S0up ¼ FSti Rs ð1 þ f EH bH Rs Þ þ ð1 þ bH Rs Þ f cv ðmgVSSÞ
4.14.9.3.2 For the system
MXBH ¼ XBH Vp
ð102aÞ
MXEH ¼ XEH V p
ð102bÞ
MXi ¼ Xi Vp
ð102cÞ
MXv ¼ Xv Vp
ð102dÞ
ð106Þ
MXIO ¼ FXIOi Rs þ f iOHO MXBH ðmgISSÞ
ð107Þ
and hence
MXt ¼ MXv þ MXIO ð1 f S0us f S0up ÞYH Rs ð1 þ f EH bH Rs þ f iOHO Þ ¼ FSti ð1 þ bH Rs Þ f S0 up XIOi ðmgTSSÞ ð108Þ þ Rs þ f cv Sti The VSS/TSS ratio of the sludge (fi) is
fi ¼
The masses of OHO VSS (MXBH, mgVSS), endogenous residue VSS (MXEH, mgVSS), unbiodegradable organics VSS (MXI, mgVSS), volatile settleable solids VSS (MXv, mgVSS), and the TSS (MXt, mgTSS) in the system are
447
MXv MXt
ðmgVSS=mgTSSÞ
ð109Þ
If the influent ISS concentration (XIOi) is not known, then the reactor TSS mass (MXt) can be calculated from an estimated VSS/TSS ratio (fi) of the sludge, that is,
MXt ¼ MXv =f i
ðmgTSSÞ
ð110Þ
The flux of oxygen utilized (FOc, mgO d1) is
FOc ¼ FSbi ð1 f cv YH Þ þ f cv ð1 f EH ÞbH ¼ FSti ð1 fS0us fS0up Þ ð1 f cv YH Þ þ f cv ð1 f EH ÞbH
YH Rs ð1 þ bH Rs Þ
YH Rs ðmgO d1 Þ ð1 þ bH Rs Þ
ð111Þ FOc ¼ OURc Vp MXBH ¼ FSbi
ð102eÞ
YH Rs ð1 þ bH Rs Þ
¼ FSti ð1 f S0us f S0up Þ
YH Rs ð1 þ bH Rs Þ
ðmgVSSÞ
Vp ¼ MXt =Xt ð103Þ
MXEH ¼ f EH bH Rs MXBH YH Rs f EH bH Rs ¼ FSbi ð1 þ bH Rs Þ ¼ FSti ð1 f S0us f S0up Þ
YH Rs f EH bH Rs ð1 þ bH Rs Þ
ðmgVSSÞ
MXI ¼ FXli Rs ¼ ðmgVSSÞ
Knowing the mass of total settleable solids (MXt) in the reactor, the volume of the reactor is determined from the value specified for the MLSS concentration, Xt (see Section 4.14.10):
ð104Þ f S0up FSupi Rs ¼ FSti Rs f cv f cv ð105Þ
ðl; m3 ; or MlÞ
ð112Þ
Knowing the volume Vp, the nominal hydraulic retention time, Rhn, is found from the design average dry weather flow rate Qi from Equation (59a). The above design equations lead to the following important conclusions for the steady-state model: the mass of VSSs in the reactor is a function principally of the daily flux of COD on it and the sludge age. Similarly, the mass of TSS in the reactor is a function principally of the daily flux of COD and ISS on the reactor and the sludge age. Consequently, insofar as the mass of sludge in the reactor is concerned, it is immaterial whether the flux of COD (and ISS) arises from a low daily flow with a high COD (and ISS) concentration, or a high daily flow with a low COD (and ISS) concentration. Provided FSti (and FXIOi) is the same in both cases, the masses of sludge (TSS) in
448
Biological Nutrient Removal
the reactor will be virtually identical. However, the hydraulic retention times will differ, being long in the first and short in the second case, respectively. The hydraulic retention time, therefore, is incidental to the COD (and ISS) fluxes, the reactor VSS (and TSS) masses, and the daily flow – it serves no basic design function in the steady-state model. Design criteria for the AS reactor volume based on hydraulic retention time should therefore be used with extreme caution because they implicitly incorporate specific wastewater strength and characteristics values typical for the regions for which they were developed.
4.14.9.4 Active Fraction of the Sludge The active VSS mass MXBH in the reactor is the live OHO mass which performs the biodegradation processes of the organic material. The other two volatile solids masses, MXEH and MXI, are inactive and unbiodegradable and do not serve any function insofar as the biodegradation processes in the system are concerned. They are given different symbols because of their different origin, the MXI is UPOs from the influent wastewater and the MXEH is UPOs produced in the reactor via the endogenous respiration process. The active OHO fraction of the volatile solids in the reactor favOHO is given by
f avOHO ¼
MXBH MXv
ðmgOHOVSS=mgVSSÞ
ð113Þ
Substituting Equations (103) and (106) for MXBH and MXv and rearranging yields
f S0up ð1 þ bH Rs Þ 1 ¼ 1 þ f EH bH Rs þ f avOHO f cv YH ð1 f S0up f S0us Þ
ð114Þ
where fatOHO is the active OHO fraction of the VSS mass. If the TSS mass is used as the basis for determining the active fraction, then the active fraction of the sludge mass with respect to the total settleable solids, fatOHO, is given by
f atOHO ¼ f i f avOHO
ð115Þ
where fi is the VSS/TSS ratio of the AS. The active fractions favOHO or fatOHO give an indication of the stability of the waste sludge, which is related to the remaining biodegradable organics in the sludge mass. The only biodegradable organics in the VSS mass are those of the OHOs which in terms of the steady-state model is 80% (1 fEH) of the OHO mass. Hence, the higher the active fraction, the greater the proportion of biodegradable organics in the sludge mass and the greater the utilizable energy content of the sludge mass. For an AS (or digested primary sludge) to be stable, the remaining utilizable organics in it should be low so that it will not generate odors through further significant biological activity. Ekama et al. (2006b) and Harding et al. (2009) have shown experimentally that the UPOs of the influent (UPO, Supi) and the endogenous residue organics (XEH) generated in the AS system remain unbiodegradable in aerobic and anaerobic digestion. Therefore, the biodegradable COD fraction of primary sludge and WAS can be calculated from its wastewater characteristics (fS’up and fS’us) and AS active
fraction, respectively. Sludge used as a soil conditioner needs to be stable because its primary purpose is to provide nutrients and unbiodegradable organic content to the soil (Korentajer, 1991); unstable sludges applied to agricultural land lead to an undesirable high oxygen demand in the soil through significant residual biological activity.
4.14.9.5 Steady-State Design Chart The design equations set out above form the starting point for aerobic and anoxic–aerobic N removal AS system design, from the relatively simple single reactor completely mixed aerobic system to the more complex multireactor anoxic–aerobic systems for biological nitrogen removal. When BEPR is included in the AS system the above equations do not give an accurate estimate for the VSS and TSS masses in the system. With BEPR, a second group of heterotrophic organisms, the PAOs need to be considered, which have different stoichiometric and kinetics constants producing more VSS and TSS mass per mass organics (COD) utilized. Incorporation of the PAOs in the steady-state model is presented in Section 4.14.31. For the more complex anoxic–aerobic systems, the above basic equations apply if the assumptions made in their derivation apply. Provided this is the case, the effects of nitrification (see Section 4.14.17) and ND (see Section 4.14.24) and the associated oxygen demands can be formulated as additional equations to the basic equations above. That the above simplified approach is adequate for design rests principally on the assumption that the biodegradable organics are completely utilized. This has been established by the close correlation achieved between the mean response of the more complex anoxic aerobic systems predicted by the more complex general kinetic model (and validated experimentally), with that calculated by the above basic equations and the additional equations for ND. The close correspondence between this simplified steady-state model and the more complex general kinetic simulation model is demonstrated for anoxic–aerobic systems including anoxic–aerobic digestion by So¨temann et al. (2006). Indeed, the simplified models developed in this chapter form the basis for hand calculations to (1) develop design input information for and (2) check output results from the dynamic simulation models. Other assumptions on which the steady-state model is based are that (1) the mass of active OHOs seeded into the system with the influent is negligible in comparison with that which grows in the reactor and (2) there is no loss of solids in the effluent from the secondary settling tanks, (3) water mass is conserved, (4) a 100% COD balance is achieved, and (5) active OHO loss is modeled as endogenous respiration. It is important to take cognizance of these assumptions in the model. With regard to the assumption of complete utilization of biodegradable organics, if this is not the case, then the mass of sludge produced per day increases and the carbonaceous oxygen demand decreases below those predicted by the basic equations. The reason for these deviations lies in the kinetics of degradation of the slowly biodegradable particulate material – if, for example, the aerobic fraction of the sludge mass is too small, the BPOs are only partially utilized and residual BPOs accumulate in the system as additional VSS like UPOs. Concomitantly, the carbonaceous oxygen demand is
Biological Nutrient Removal
reduced because less biodegradable organics are utilized. Clearly, for such situations the simple steady-state equations are inappropriate – approximate solutions are sometimes obtainable by simulation using the general kinetic models
such as UCTOLD (Dold et al., 1991) or ASM1 (Henze et al., 1987). Graphs of Equations (103)–(111) and (113)–(115) are shown in Figures 15(a)–15(d) per COD flux (FSti). The values
Settled wastewater
Raw wastewater
10
10 T = 20 °C YH = 0.45 bH = 0.24 fcv = 1.48 fEH = 0.20
T = 20 °C YH = 0.45 bH = 0.24 fcv = 1.48 fEH = 0.20
f S′up = 0.15 f S′us = 0.07 f i = 0.75 8 Sludge mass (kg/kgCOD on reactor)
MXt 6 MXv
4 MXi 2
f S′up = 0.04 f S′us = 0.12 f i = 0.83
6 MXt 4 MXv MXi
2
MXEH
MXBH MXEH
MXBH 0
0 20 15 10 Sludge age (days)
5
0 (a)
25
5
0
30
FOc 0.6
0.4
0.4 favOHO 0.2 fatOHO
0.0 10 15 20 Sludge age (days)
25
0.8
FOc 0.6
0.6 favOHO
0.4
0.4 T = 20 °C YH = 0.45 bH = 0.24 fcv = 1.48 fEH = 0.20
0.2
fatOHO 0.2
0.0
0.0 5
30
1.0 f S′up = 0.04 f S′us = 0.12 f i = 0.83
0.8
Active fraction
0.8
Oxygen demand (kg/kgCOD on reactor)
Active fraction
f S′up = 0.15 f S′us = 0.07 f i = 0.75
0.6
0
25
Settled wastewater
0.2
(b)
20
1.0
1.0
0.8
15
Sludge age (days)
Raw wastewater 1.0 T = 20 °C YH = 0.45 bH = 0.24 fcv = 1.48 fEH = 0.20
10
(c)
0.0 0
30 (d)
Oxygen demand (kg/kgCOD on reactor)
Sludge mass (kg/kgCOD on reactor)
8
449
5
10 15 20 Sludge age (days)
25
30
Figure 15 Active (MXBH), endogenous (MXEH), insert (MXI), volatile (MXv), and total (MXt) masses (kg) of settleable solids in the reactor per kgCOD d1 organic flux on the reactor for (a) the example raw wastewaters and flux carbonaceous oxygen demand (FOc, kgO d1) per kgCOD d1 organic flux on the reactor and active fraction with respect to volatile solids (favOHO) and total solids (fatOHO) vs. sludge age for (b) the example raw wastewaters vs. sludge age from 3 to 30 days at 20 1C, (c) the settled wastewaters and flux carbonaceous oxygen demand (FOc, kgO d1) per kgCOD d1 organic flux on the reactor and active fraction with respect to volatile solids (favOHO) and total solids (fatOHO) vs. sludge age for (d) settled wastewaters vs. sludge age from 3 to 30 days at 20 1C.
450
Biological Nutrient Removal
Table 7 Influent wastewater COD characteristics for the example raw and settled wastewaters Wastewater characteristic
Unbiodegradable soluble COD fraction (fS’us) Unbiodegradable particulate COD fraction (fS’up) Influent ISS (XIOi)a mgISS l1
Wastewater type Raw
Settled
0.07 0.15 47.8
0.12 0.04 9.9
a
These influent ISS concentrations give VSS/TSS (fi) values for the raw and settled wastewater systems at 20 days sludge of 0.75 and 0.83 mgVSS/mgTSS, respectively.
of the kinetic constants, YH, bH, fcv, and fEH are listed in Table 6 for 20 1C and have been validated in extensive laboratory and pilot-scale investigations over the years. The values of the unbiodegradable particulate and soluble COD fractions (fS’up and fS’us, respectively) and the AS VSS/TSS ratio (fi) are listed in Table 7 and are values for the example raw and settled municipal wastewaters that will used throughout this chapter. Figures 15(a) and 15(c) give the masses MXBH, MXEH MXI, MXv and MXt in the reactor for a unit flux of COD applied per day to the reactor for sludge ages from 3 to 30 days for the example raw and settled wastewaters, respectively. Figures 15(b) and 15(d) show the associated flux of oxygen to be supplied for organic material degradation (FOc) and the active fraction of the sludge with respect to VSS (favOHO) and TSS (fatOHO) for sludge ages from 3 to 30 days for the example raw and settled wastewaters, respectively. For a particular wastewater, the volume of the reactor at two different sludge ages will be in direct proportion to the TSS mass (MXt) in the reactor if the same reactor concentration Xt (mgTSS l1) is specified. The graphs show that the active mass increases fairly rapidly with an increase in sludge age up to about 10 days, after which it increases only marginally. The carbonaceous oxygen demand shows a similar behavior, increasing significantly from 3 to 10 days sludge age and thereafter more gradually. In contrast, the fractions of endogenous and inert solids increase rapidly relative to the active mass at sludge ages greater than 10 days and consequently in systems treating raw wastewater at sludge ages longer than 10 days; only a relatively small fraction of the sludge mass are active OHOs; most of the VSS sludge mass is unbiodegradable organic material from the influent and the endogenous process.
4.14.9.6 The Calculation Procedure The calculation procedure to generate the design results required for a certain sludge age is as follows: Select the wastewater characteristics fS’up, fS’us, and XIOi which are believed to best represent the unbiodegradable particulate and soluble COD fractions and ISS content of the wastewater. Then calculate 1. Supi (Equation (4)), Susi (Equation (3)), and XIi (Equation (6)) 2. FSti (Equation (99)) and/or FSbi (Equation (100)), FXIi (Equation (101)) and FXIOi ( ¼ Qi XIOi); 3. select the sludge age Rs;
4. MXBH (Equation (103)), MXEH (Equation (104)), MXI (Equation (105)), MXv (Equation (106)), MXIO (Equation (107)), or select fi, MXt (Equation (108) or (110)); 5. Select Xt, Vp (Equation (112)); 6. FOc (Equation (111)) and OURc (Equation (102e)); 7. Rhn, (Equation (59a)); and 8. Ste (Equation (95)). In this design procedure the input COD and its characteristics will be governed by the specific waste flow. The parameter that requires selection is the sludge age; this will depend on the specific requirements from the WWTP such as effluent quality, that is, organic COD removal only, nitrification, N removal, biological P removal, and the envisioned sludge treatment facilities, that is, whether or not primary settling is included, the stability of the WAS, etc. Specification of the sludge age, therefore, is an important design decision and requires special consideration (see Section 4.14.15). For illustrative purposes, however, the effect of sludge age on the various design results such as reactor volume, oxygen demand, and sludge production and nutrient (N and P) requirements for sludge production are demonstrated in the following sections.
4.14.10 Reactor Volume Requirements Once the mass of sludge in the reactor is known from a specified sludge age and influent organic COD flux, the reactor volume is determined by diluting this mass of sludge to a specified TSS concentration (Xt). From the volume, the nominal hydraulic retention time, or aeration time for fully aerobic systems, is fixed (by Equation (59)). Hydraulic retention time therefore is immaterial in the design procedure – it is a consequence of the mass of sludge in the reactor and a selected TSS concentration. This point was mentioned earlier but bears repeating because some design procedures lay stress on retention time or aeration time as a basic design parameter, an approach that can result in serious miscalculation of the reactor volume requirements. Compare, for example, two plants operating at the same sludge age, both receiving the same organic load (kgCOD d1) but the first at high influent COD concentration and low flow and the second at a low concentration and high flow. If designed on a specified hydraulic retention time, the volume of the first will be much smaller than that of the second but the sludge mass in the reactors will be the same. Consequently, the first plant may have an inordinately high TSS concentration which may cause problems in the secondary settling tank. Therefore, retention time is a completely inappropriate basis for design and other purposes such as a criterion for comparing the reactor volume requirements of different plants. Figure 16 shows the reactor volume requirements versus sludge for the example raw and settled wastewaters obtained from Equations (99)–(115). The reactor volume requirements may also be determined from the equivalent COD load per person equivalent (PE), also shown in Figure 16 for a raw WW COD load of 0.10 kg COD/person/day. Hence, treating the example raw WW at a sludge age of 20 days and a TSS concentration of 4 kgTSS m3, a reactor volume of 145 l
Biological Nutrient Removal Reactor volume 350
3.5 Raw wastewater load = 0.10 kgCOD/(PE d) 40% COD removal in PST
2.5
300
250
Reactor TSS Kg m−3
3 200
2.0 4 Raw 1.5
150
5
1.0
Liters per PE
m3/kgCOD per day on treatment plant
3.0
100
3 4
0.5
451
WW) systems – in a survey of 45 full-scale AS plants in the Netherlands, Stofkoper and Trentelman (1982) found significantly higher DSVIs in settled WW systems than in raw WW systems (see Ekama and Marais, 1986). The effect of WW strength and sludge settleability, as well as other factors such as the peak wet weather flow (PWWF) to average dry weather flow (ADWF) ratio (or peak flow factor fq ¼ PWWF/ADWF), WW and AS characteristics (fS’up, fS’us, fi) and construction costs, can all be taken into account by determining the reactor concentration from a construction cost minimization analysis (Ho¨rler, 1969; Dick, 1976; Riddell et al., 1983; Pincince et al., 1995). In such an analysis, the construction cost of the reactor(s) and the SSTs is determined as a function of the reactor TSS concentration. The reactor concentration at which the combined construction cost of the reactor(s) and the SST(s) is a minimum, is the design reactor concentration.
50
5
4.14.11.1 Reactor Cost Settled 0
0.0 0
5
10 15 20 Sludge age (days)
25
30
Figure 16 Reactor volume requirements in m3 kg1 COD raw WW load per day vs. sludge age at different average reactor TSS concentrations for raw and settled WW (assuming 40% COD removal by primary sedimentation). Reactor volume requirements in liter or person equivalent (PE) is also given on the right-hand vertical axis based on a raw WW COD contribution of 0.10 kg COD/person equivalent.
per PE is required or 1.45 m3 kg1 COD applied per day to the WWTP. The comparative reactor volume requirements for settled WW per kgCOD load per day on the WWTP also is shown in Figure 16 taking due consideration of the COD fraction removed by primary sedimentation (40% for the example settled wastewater) and the reduction in settled wastewater UPO fraction (fS’up) this causes (Table 6). From Figure 16, treating settled WW at a sludge age of 20 days and reactor TSS concentration of 4 kgTSS m3 requires a reactor volume of 0.55 m3 kg1 raw WW COD load per day on the WWTP or 55 l per PE. Therefore, a significant reduction in reactor volume can be obtained by means of primary sedimentation – 62% for the example raw and settled WWs at 20 days sludge age.
4.14.11 Determination of Reactor TSS Concentration The choice of the reactor concentration can be done empirically from past experience with similar WWs or selected from design guidelines such as those from Metcalf and Eddy (1991), for example, for conventional systems (with primary sedimentation) 1500–3000 mgTSS l1 or extended aeration (without primary sedimentation) 3000–6000 mgTSS l1. Differences in the reactor TSS concentration for raw and settled WWs arise because (1) the WW flow per kgCOD load on the reactor for raw WW is significantly greater than that for settled WWs and (2) sludge settleability in conventional (settled WW) systems can be poorer than with extended aeration (raw
For selected WW and AS characteristics (fS’up, fS’us, XIOi, or fi), sludge age (Rs), and organic COD load on the reactor (FSti Reactor), the mass of TSS in the reactor (MXt) can be determined from the steady-state model (Section 4.14.9) and remains constant at a fixed sludge age. The reactor volume as a function of the reactor TSS concentration Xt is found from Equation (112), viz., Vp ¼ MXt/Xt m3, where Xt is the reactor concentration in kgTSS m3. To estimate the cost of the reactor from the volume, empirical functions relating the construction cost of the reactor to the volume are required. Such functions show (1) that as the reactor becomes smaller so its construction cost gets lower (Figure 17) and (2) a benefit of scale effect in that it is cheaper (per m3) to build a large reactor than a small one.
4.14.11.2 SST Cost On the basis of the flux theory, Ekama et al. (1997) show that provided the underflow recycle ratio R is above the critical minimum value, the maximum overflow rate at PWWF (qPWWF, m h1) of the SST is a function of only the reactor (or feed) solids concentration (Xt) and the sludge settleability. Therefore, the maximum overflow rate in the SST must not exceed the settling velocity of the AS at the feed concentration (Xt). For a selected sludge settleability, if the reactor (Xt) concentration increases, the settling velocity of the sludge decreases, with the result that the maximum overflow rate in the SST must be lower for higher Xt. Hence, the required surface area for the SSTs (AST) gets larger as the reactor concentration increases. Therefore, as the biological reactor becomes smaller with increasing concentration, the SST area and its construction increase.
4.14.11.3 Total Cost The total cost of the reactor–SST system is the sum of the reactor and SST costs. Qualitative results for the example raw and settled WWs are given in Figure 17, ignoring that the reactor volume and SST diameter may have upper and lower size restrictions. For real WWTPs, the reactor and/or SST may
452
Biological Nutrient Removal Cost minimization 20
Cost minimization
20 RAW WW Rs = 20 days DSVI = 120 ml g−1 Reactor conc. for minimum cost
10
Reactor conc. for minimum
15
Cost
Cost
15
Settled
Total
10
Reactor
Total
5
5 Reactor SST
SST
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0 0
2
4
6
8
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Reactor conc. (kg TSS m−3)
0
2
4
6
8
10
Reactor conc. (kg TSS m−3)
Figure 17 Reactor, secondary settling tank, and total construction costs to estimate the reactor TSS concentration for minimum total cost in single reactor and SST units for the example raw and settled wastewaters.
need to be split into two or more equal sized modules to bring the volume and diameter within the limit ranges. Basically, the reactor volume depends on organic load (FSti) and the SST surface area on hydraulic load (PWWF). This is the reason why in Figure 17 the cost of the SST for the raw and settled wastewaters is the same but the cost of the reactor for the raw wastewater is higher than for the settled wastewater. From cost minimization analyses such as that above, generally it will be found that the range of reactor concentration for minimum construction cost (1) is higher for higher influent WW strengths (BOD5, COD), (2) is higher for longer sludge ages, and (3) is higher for raw WW than settled WW at the same strength, because these three changes all increase the size of the biological reactor relative to that of the settling tank, (4) is lower for higher peak flow factors (fq), and (5) is lower for poorer settling sludges because these two changes all increase the size of the settling tank relative to that of the biological reactor. A universal optimum therefore cannot be specified. In countries with low WW strengths and short sludge age plants (e.g., North America), the reactor concentration tends to be low (2000–3000 mgTSS l1) and in countries with high WW strengths and long sludge age plants (e.g., South Africa), the reactor concentration tends to be high (4000–6000 mgTSS l1) as the example WWs demonstrate.
4.14.12 Carbonaceous Oxygen Demand 4.14.12.1 Steady-State (Daily Average) Conditions The mean daily carbonaceous oxygen demand per kgCOD load on the reactor (FOc/FSti Reactor) is calculated from Equation (111). For sludge ages longer than 15 days the increase in FOc/FSti Reactor is small with further increase in sludge age for both raw and settled wastewater (Figures 15(b) and 15(d)). The FOc/FSti Reactor for raw and settled wastewater is usually within 10% of each other, with the demand for settled wastewater being the higher value. This is because compared to raw wastewater, a higher percentage of the total organics
(COD) load in settled wastewater is biodegradable. For example, the wastewaters at 20 days sludge age, the FOc/FSti Reactor is 0.604 kgO/kgCOD for raw wastewater and 0.653 kgO/ kgCOD for settled wastewater. Although there is only a small difference in FOc/FSti Reactor between raw and settled wastewaters, there is a large difference in the oxygen demand (FOc) because primary settling removes a significant proportion of the WWTP organic load as primary sludge (PS). For settled wastewater, this is given by 0.653 (1– 0.40) for 40% COD removal in PSTs, which gives 0.38 kgO/ kgCOD load on the WWTP. For the raw wastewater, it would remain 0.604 kgO/kgCOD load on the treatment plant, making the settled wastewater oxygen demand 37% lower than that for the raw wastewater. Clearly, primary sedimentation leads to significant aeration energy savings – because primary settling tanks remove about 30–50% of the raw influent COD, the carbonaceous oxygen demand for settled wastewater generally will be about 30–50% lower than that for raw wastewater. The carbonaceous oxygen demand is the oxygen demand for the oxidation of the influent organics (COD) and the associated OHO endogenous process only. In N removal systems, oxygen is also required for nitrification, which is the biological oxidation of ammonia to nitrate by autotrophic nitrifiers. However, with denitrification, which is the biological reduction of nitrate to nitrogen gas by facultative heterotrophic organisms, some of the biodegradable organics are utilized with nitrate as electron acceptor, for which oxygen is then not required. Thus, denitrification leads to a reduction in the oxygen demand. The total oxygen demand for a N removal system therefore is the sum of the carbonaceous and nitrification oxygen demands less that saved by denitrification. The procedures for calculating the oxygen demand for nitrification and the oxygen saved by denitrification are discussed in Sections 4.14.22.3 and 4.14.27.2. The equations given here for calculating the carbonaceous oxygen demand are based on the assumption that all the biodegradable organics are utilized with oxygen as electron acceptor, that is, for fully aerobic systems.
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Sludge production
4.14.12.2 Daily Cyclic (Dynamic) Conditions Owing to the daily cyclic nature of the organic (COD) load on the reactor, the carbonaceous oxygen demand will vary concomitantly over the day. The TKN load on the reactor also varies over the day in an approximately similar fashion as the organic load. Generally, the COD and TKN loads on the reactor increase in the morning due to increases in both flow and COD and TKN concentration reaching a peak at around noon. Thereafter, the COD and TKN loads decrease reaching a minimum during the nighttime hours of 2–4 am due to decreases in both flow and COD and TKN concentration. The peak-to-average and minimum-to-average load ratios and the time of day these occur depend on the catchment that the particular treatment plant serves, such as size of population, layout of the catchment and industrial activity. Generally, the smaller the catchment, the lower the flow and COD and TKN loads but the greater the peak to average flow and load ratios and the lower the minimum to average flow and load ratios. As the TKN load and its variation over the day and the nitrification process have a profound influence on the daily average and peak total oxygen demands, empirical methods to estimate the peak oxygen demand from the average for fully aerobic nitrifying systems are discussed in Section 4.14.27.2. Fully aerobic AS systems with sludge ages longer than 3 days are likely to nitrify at temperatures 414 1C. Moreover, a sludge age of 3 days is around the limit of validity for the steady-state AS model because at sludge ages lower than this the assumption that all the biodegradable organics are utilized is not valid. Therefore, there is little merit in developing empirical methods for estimating the peak oxygen demand for fully aerobic systems without nitrification.
4.14.13 Daily Sludge Production The mass of sludge produced per day by the AS system is equal to the mass of sludge wasted per day from it via the waste flow and is called WAS or secondary sludge. From the definition of sludge age (see Equation (58)), the mass of sludge TSS produced per day (flux) FXt is given by the mass of sludge in the system MXt divided by the sludge age (Rs), that is,
FXt ¼ MXt =Rs
ðmgTSS d1 Þ
ð116Þ
Substituting Equation (108) for MXt and simplifying yield the sludge produced per day per mgCOD load on the biological reactor, that is, ð1 f S0 us f S0 up YH Þ f S0 up XIOi FXt ð1 þ f EH bH Rs þ f iOHO Þ þ ¼ þ FSti f cv Sti ð1 þ bH Rs Þ
ðmgTSS=dÞ=ðmgCOD=dÞ
ð117Þ
A plot of the daily TSS produced per unit COD load on the biological reactor (Equation (117)) versus sludge age (Rs) is shown in Figure 18 for the example raw and settled wastewaters. It can be seen that the mass of sludge produced in the AS system (per unit COD load on the biological reactor) decreases as the sludge age increases for both raw and settled wastewater but the rate of decrease is negligible at sludge ages longer than about 20 days. Treating settled wastewater results
kgTSS/d per kgCOD/d on reactor
0.5
0.4
Raw 0.3
Settled
0.2
T = 20 °C YH = 0.45 bH = 0.24 f cv = 1.24 f EH = 0.20
0.1
0.0 0
5
10
15
20
25
30
Sludge age (days) Figure 18 Daily sludge production in kgVSS d1 and kgTSS d1 per kgCOD load per day on the biological reactor for the example raw and settled wastewaters at 14 1C.
in lower secondary sludge production per unit COD load on the biological reactor than treating raw wastewater. This is because the unbiodegradable particulate COD fraction (fS’up) and inorganic content (XIOi/Sti) in settled wastewater are significantly lower than that in raw wastewater. Temperature effects on secondary sludge production are small – sludge production at 14 1C is about 5% greater than at 22 1C, a difference which is completely masked by the uncertainty in the estimates of the wastewater characteristic fS’up and the VSS/TSS ratio (fi) of the sludge if the influent ISS concentration (XIOi) is not measured. Although the secondary sludge production treating settled wastewater is lower than that treating raw wastewater, the total sludge mass treating settled wastewater is higher because the total sludge production includes both the primary and secondary sludges; at plants treating raw wastewater, only secondary sludge is produced. In systems treating raw wastewater, the primary sludge is in effect treated in the AS reactor itself. From the COD balance, the more the oxygen utilized in the system, the lower the sludge production and the lower the active fraction of the sludge (Figures 14 and 15). Therefore, because the carbonaceous oxygen demand is much higher when treating raw wastewater, the overall sludge production is much lower compared with settled wastewater. Generalizing the above observations and taking into account the active fraction of the WAS as an indication of the remaining biodegradable organics, there are two extremes in approach to designing WWTPs with AS for biological treatment, viz., 1. Treating settled wastewater at a short sludge age (say 5–8 days). This results in a very small AS system with low oxygen
454
Biological Nutrient Removal
demand and a high sludge production with high energy content, that is, high remaining biodegradable organics in both the primary and secondary (waste activated) sludges requiring further stabilization treatment before disposal, or 2. Treating raw wastewater at a long sludge age (say 30 days). This results in a very large AS system with high oxygen demand and a low sludge production with a low energy content, that is, no primary sludge and low remaining biodegradable organics (low active fraction) in the secondary sludge, not requiring further stabilization treatment before disposal. The daily production of secondary and primary sludges is the mass of sludge that needs to be treated and disposed of by downstream sludge handling methods. Sludge treatment and disposal for BNR systems, in particular, should not be seen as separate from the design of the AS system. In fact, all unit operations of the WWTP from raw wastewater pumping to ultimate disposal of the sludge should be viewed as an integrated system where the design of one unit operation depends on the unit operations before it, and decisions on its design may affect the design of unit operations following it.
4.14.14 System Design and Control The parameter of fundamental importance in the design and control of the AS system is the sludge age, which governs the mass of sludge to be wasted daily from the system. The sludge age can and should replace completely the food to microorganism ratio (F/M, kgBOD or COD load per day per kgMLSS or MLVSS in reactor) or equivalently the load factor (LF) as a reference and control parameter, in particular if nitrification is required. The sludge age can be fixed by a simple control procedure if the system is appropriately designed. This control procedure is simpler and operationally more practical and reliable than procedures based on the F/M or LF, which basically seek to control the mass of sludge in the system by controlling the reactor MLSS concentration at some specified value.
4.14.14.1 System Sludge Mass Control By far the most common AS system control procedure involves keeping the sludge MLSS concentration in the reactor at some specified value. At best this sludge concentration is specified by design or at worst, established from operational experience on the plant behavior, which is usually the concentration that can be contained in the system by the SSTs. This approach does not control sludge age, only the mass of sludge in the system. In fact, in some instances, it may not even be the sludge mass that is controlled via the reactor concentration, but the settled volume at 30 min (SV30) in the 1l measuring cylinder. If the SV30 is greater than say 450 ml l1, then sludge is wasted until it reaches this value again. This approach was developed to obviate the need for measuring the reactor sludge concentration and with it, the sludge concentration in the reactor varied with the sludge settleability (SVI). This approach was acceptable before nitrification became obligatory and at least ensured that the sludge could be contained in the system while maintaining a low effluent suspended solids (ESS)
concentration. However, with this method there is no control of the F/M, the LF, the sludge mass, the reactor concentration, or the sludge age, a situation which is completely untenable when nitrification is required. While nitrification is a simple process to cater for in design – just make the sludge age long enough and provide sufficient oxygen – it imposes a completely different control regime on the operation of the system. It requires the sludge age to be controlled at a fixed value. If the F/M or LF are controlled, then to keep these parameters within the desired limits, not only does the reactor concentration need to be measured regularly, but also the daily BOD5 (or COD) load. This requires extensive sampling and testing of the influent BOD5 (or COD) concentration and flow pattern over the day to determine the daily COD (or BOD) mass load. Controlling the sludge age requires measuring the reactor MLSS concentration and the mass of sludge wasted per day. Usually, the waste sludge is abstracted from the SST underflow to benefit from its thickening function. However, the sludge concentration of the underflow varies considerably over the day with the daily cyclic flow through the plant (Figures 19(a) and 19(b)). Therefore, to know the sludge mass wasted via the underflow, it is necessary to measure the underflow concentration, waste flow rate, and duration each time sludge is wasted. Therefore, to know the LF or sludge age, intensive testing of the influent and/or reactor and underflow concentrations are required. This may be manageable at large plants where the technical capacity is adequate, but on small plants, both the LF and sludge age usually are not known. As a consequence, nitrification is sporadic, partial, or stops altogether during periods of poor sludge settleability, which results in high sludge wastage and hence short sludge age. Even if the reactor concentration were accurately controlled with modern control equipment such as automated wasting and on-line reactor concentration measurement, this still does not control the sludge age. With reactor concentration control at the same value throughout the year and a stable organic load on the plant (zero urban development), the sludge age decreases during winter because sludge production per kgCOD load increases with decrease in temperature due to the lower endogenous respiration rate. Although the decrease is relatively small, decreasing the sludge age is nevertheless the opposite of what should be done to the system in winter to keep the ammonia concentration low. This is particularly relevant to plants operated at sludge ages close to the minimum for nitrification, (Section 4.14.20.3), which is common practice in developed countries to squeeze as much capacity as possible out of the plant because space for extensions is limited. If the reactor concentration is controlled and the organic load on the plant progressively increases, which is usually the case in developing countries where often urban growth is constrained by WWTP capacity, the sludge age decreases progressively with time (Figure 20a). Inevitably, on one cold winter day, nitrification will have stopped. The operators think the cause is a toxic batch of wastewater (which it could be but is not, in this instance) and hope that nitrification returns soon – it does not, because the sludge age is too short, certainly not until summer when the wastewater temperature increases. Thereafter, if sludge concentration is controlled, loss of nitrification will become a seasonal occurrence.
Biological Nutrient Removal
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Reactor TSS conc.
16 000
5
4 8000 Reactor Xt 4000 Low flow Low XR
High flow High XR
0 24h00
06h00
18h00
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Low flow period high recycle ratio (for constant QR)
Xt
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Constant MLSS conc
3
Increasing organic load
2
Decreasing sludge age
0 0
QI + QR QI
1.3 1.2 1.1
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(a)
QI + QR
12h00
Reactor MLSS (gTSS l−1)
Concentration (mg I−1)
SST underflow, XR 12 000
5
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10
20
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Sludge age (Rs)
(a)
QI
Reactor TSS conc. 5
Settled sludge XR
QR
Q I = influent flow rate (m3 h−1) Q R = recycle flow rate (m3 h−1)
(b)
4
XR
X t = reactor concentration (mgTSS I−1) X R = recycle concentration (mgTSS I−1) = (Q I + Q R)X t/Q R = (1 + R)Xt /R R = recycle ratio = Q R/Q I
Figure 19 (a) Experimental data from a full-scale activated sludge plant illustrating the virtually constant reactor concentration compared with the widely varying SST underflow concentration over the day. From Nicholls HA (1975a). (b) Increased sludge accumulation and higher underflow sludge concentration at high influent flow periods (right) than at low influent flow periods (left) at constant recycle flow rate.
Reactor MLSS (gTSS l−1)
QR
1.3 1.2 1.1 1.0
3
Increasing organic load
2
1 Constant sludge age 0 0 (b)
When nitrification is required, not only should sludge age be controlled, but also the SST can no longer serve the dual purpose of clarifier and WAS thickener. To obtain high WAS concentrations, the underflow recycle ratio must be low (o0.25:1), which results in long sludge residence times in the SST (Figure 19(b)). The long sludge residence time stimulates denitrification in the SSTs causing floating (or rising) sludge on the SST surface, in particular in summer when wastewater temperature is high (420 1C). In fact, in the tropics, the climatic region of most developing countries, it may not be possible to operate an AS system that does not nitrify even at very short sludge ages. So the problem of rising sludge due to denitrification can take place in plants even where nitrification is not a requirement. This happened at Brazilia WWTP, which had a low return sludge ratio (0.25:1) – it nitrified even at 3 days sludge age and suffered from floating sludge all the time. If the sludge age was reduced to stop nitrification, the COD removal deteriorated below an acceptable level. So once
Increasing reactor MLSS
5
10
15
20
25
30
Sludge age (Rs)
Figure 20 (a) With reactor TSS controlled at a constant concentration, sludge age decreases as sludge production increases due to increasing organic load or wastewater temperature decrease. (b) With sludge age control, the reactor TSS concentration increases with increase in sludge production due to increase in organic load or decrease in wastewater temperature.
nitrification takes place, whether intentionally by design or unavoidably, one must cater to denitrification in appropriate reactor (anoxic) zones and increase the underflow recycle ratio (B1:1) to minimize rising sludge in the SSTs due to denitrification. Clearly, once nitrification takes place, whether as a requirement for N removal or unavoidably due to system conditions, one is forced to abandon using the SSTs as WAS thickeners. If one has to thicken WAS in a separate unit, whether from the underflow or reactor, one may as well waste sludge directly from the reactor and derive the significant
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Biological Nutrient Removal
operational benefit of hydraulic control of sludge age. It is simple, requires very little testing, and establishes the sludge age almost exactly. It results in stable year-round nitrification and is strongly recommended for AS systems where nitrification is required, even where sophisticated reactor concentration control measures can be applied.
4.14.14.2 Hydraulic Control of Sludge Age Hydraulic control of sludge age was first proposed and implemented in a generalized form by Garrett in 1958, and is based on a method of modified wastewater aeration implemented by Setter et al. (1945). If a sludge age of 10 days is specified, (1/10)th of the reactor volume is wasted daily, if 20 days, 1/20th is wasted daily, that is, Qw ¼ Vp/Rs (Equation (58)). For plants with low levels of technical support, a satellite settling tank or a dewatering drying bed completely independent of the SSTs can be provided to which the daily WAS flow from the reactor is discharged – for plants with a higher level of technical support, a dissolved air flotation unit would be best (Bratby, 1978, Bratby et al., 2008), which also minimizes P release from BEPR sludges (Pitman, 1999). The supernatant is returned to the reactor and the thickened sludge is pumped to the sludge treatment/disposal part of the plant. This procedure establishes very closely the desired sludge age because the mixed liquor concentration does not change significantly over the day (Figure 19(a)). An important point about hydraulic control of the sludge age is that irrespective of the flow through the plant, if a fixed fraction of the volume of the reactor is wasted every day, the sludge age is fixed. If the COD mass load per day on the plant remains constant, the sludge concentration will remain constant automatically. If the COD mass load increases, the sludge concentration will increase automatically, to maintain the same sludge age (Figure 20b). Thus, by monitoring the reactor concentration and its changes at a fixed sludge age, an indirect measure is obtained of the long-term changes in COD load on the plant. With time, the reactor concentration may increase indicating that the organic load on the plant is increasing. Hydraulic control of sludge age is very easy for the operator – (s)he just needs to check that the flume/pipe is not blocked and is running at the correct flow rate; the reactor MLSS concentration does not even have to be measured very often. By means of the hydraulic control procedure, the sludge age may be changed by simply changing the volume wasted per day. If say, the sludge age is reduced from 25 to 20 days by hydraulic control, the full effect of the change will become apparent only after about half a sludge age. Thus, the biomass has an opportunity to adapt gradually to the change in F/M and LF. Hydraulic control of sludge age is particularly relevant to plants with sludge ages longer than about 5 days because for these plants the mass of sludge contained in the SSTs is a relatively small fraction of the total mass of sludge in the system. At sludge ages shorter than 5 days the mass of sludge in the SSTs can become appreciable with respect to the total mass of sludge in the system, particularly when the sludge settleability gets poor (DSVI4150 ml g1). When the mass of sludge in the SSTs is significant, hydraulic control will have to
take cognizance of this and accuracy of the control will require additional testing. Hydraulic control of sludge age devolves a greater responsibility on the designer and removes responsibility from the plant operator – oftentimes operator ingenuity had to work around design inadequacies by force fitting the biological processes into the designed constraints to achieve the best effluent quality. It becomes essential that the designer calculates the sludge mass more exactly, to provide sufficient reactor volume under the design organic load to allow for the required reactor concentration at the specified sludge age. Also, the settling tank surface area, underflow recycle ratio, and aeration capacity must be accurately sized for the particular wastewater and sludge age of the system. If these aspects are catered for adequately, then with hydraulic control of the sludge age, plant control is simplified and, on small-scale plants, may even do away with the requirements for solids and SVI tests except at long intervals. Hydraulic control of sludge age makes parameters such as LF and F/M redundant and introduces an entirely different attitude to system control. It is eminently practical and establishes the desired sludge age to ensure year-round nitrification. When nitrification is a requirement, sludge age control also becomes a requirement, and then the hydraulic control of sludge age is the easiest and most practical way. Moreover, with hydraulic control of sludge age the mode of failure of the plant is completely different than with solids mass control. With the solids mass control the plant fails by nitrification stopping and a high effluent ammonia concentration, a nonvisible dissolved constituent which also is difficult to remove by other means. With sludge age control, the plant fails more obviously – sludge solids over the secondary settling tank effluent weirs and high effluent suspended solids concentration. At plants managed with low levels of technical capacity, this is more like to prompt remedial action.
4.14.14.3 Flow and Load Equalization Tanks In BNR systems of any sludge age, aeration control is a particularly vexing problem under cyclic flow and load conditions, because the system is affected by too high or too low DO concentrations in the aerobic zone. Too high DO concentrations are unnecessarily expensive and result in oxygen recycle to the anoxic (and the anaerobic zone if BEPR is included), thereby reducing the potential for N and P removal; too low DO concentrations cause nitrification efficiency to decline and possibly poor settling sludges to develop. Although some good DO control systems have been developed over the years, the cost of providing aeration capacity and SST surface area for the peak flow has prompted research into alternative control solutions such as flow and load equalization. Furthermore, most of the diurnal variation in system variables such as ammonia, nitrate, and phosphate concentration is not induced by the biological processes but by the hydraulic flow variation. To minimize hydraulic flow variation, an equalization tank is provided upstream of the AS system and outflow from this tank is controlled in such a manner that the cyclic fluctuations in flow and load are damped to very small values. The tank is controlled by
Biological Nutrient Removal
microcomputer which calculates the tank outflow rate that best damps the projected inflow of the next 24 h. This flow equalization approach has been tested at the Goudkoppies BNR plant (Johannesburg, RSA) and showed great potential for reducing aeration and other control problems in nutrient removal plants (Dold et al., 1982, 1984).
4.14.15 Selection of Sludge Age Selection of the sludge age is the most fundamental and important decision in the design of an AS system, particularly when biological nutrient removal is included. The sludge age selected for a plant depends on many factors, some of which are listed in Table 8 such as stability of the system, sludge settleability, whether or not the waste sludge should be suitable for direct discharge to drying beds, and most important of all, the quality of effluent required, that is, is COD removal only acceptable, must the effluent be nitrified, are N and P removal required. Several of the factors have already been discussed earlier and will not be repeated here. Only a few clarifying and additional comments on Table 8 are made in the following sections.
Table 8
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4.14.15.1 Short Sludge Ages (1–5 days) 4.14.15.1.1 Conventional plants These plants are operated in the conventional configuration, that is, a semi-plugflow configuration, but modified systems such as contact stabilization, step aeration, step feed, and others are also implemented. Short sludge age plants have been extensively used in Europe and North America before N (and P) removal became requirements. Their main objective is COD removal only, for which sludge ages of 1–3 days are sufficient. BOD5 or COD reductions range from 75% to 90%. The removal achieved depends on the wastewater characteristics, the operation of the plant (in particular, the management of the transfer of the sludge between the reactor and SSTs), and the efficiency of the SSTs. Because predatory activity of protozoan organisms on the free swimming bacteria is limited at short sludge ages, the nonsettling component (or dispersion) of the AS flocs is high which causes turbidity and high effluent COD (Chao and Keinath, 1979; Parker et al., 1971). It is accepted in Table 8 that short sludge age plants would not normally nitrify. For temperate and high-latitude regions, where wastewater temperatures are generally below 20 1C, this would be the case. However, in tropical and low-latitude
Some important considerations in the selection of sludge age for the activated sludge system
Sludge age
Short (2–5 days)
Intermediate (8–15 days)
Long (4 25 days)
Types
High rate, step feed, aerated lagoons, contact stabilization, pure oxygen
Extended aeration, orbal, carousel, BNR systems
Objectives
COD removal only
Similar to high rate but with nitrification and sometimes denitrification. BNR systems COD removal, nitrification, biological N removal, and/or biological P removal
Effluent quality
Low COD, high ammonia, high phosphate, variable
Low COD, low ammonia, low nitrate high/low phosphate, relatively stable
Low COD, low ammonia, low nitrate, low phosphate, usually stable
Primary settling
Generally included
Usually included
Usually excluded
Activated sludge quality
High sludge production, very active, stabilization required
Medium sludge production, quite active, stabilization required
Low sludge production, inactive, no stabilization required
Oxygen demand
Very low
High due to nitrification
Very high due to nitrification and long sludge age
Reactor volume
Very small
Medium to large
Very large
Sludge settleability
Generally good, but bulking by nonlow F/M filaments like S. natans, 1701, Thiothrix possible
Good at low sludge age and high aerobic mass fractions; but generally poor due to low F/M filament growth like M. parvicella
Can be good with high aerobic mass fractions, but generally poor due to low F/M filament growth, particularly M. parvicella
Operation
Very complex due to AS system variability and 11 and 21 sludge treatment
Very complex with BNR and 11 and 21 sludge treatment
Simple if without 11 and 21 sludge treatment, but BNR system is complex
Advantages
Low capital costs, energy self-sufficient with anaerobic digestion
Good biological N (and P?) removal at relatively low capital cost
Good biological N (and P?) removal No 11 and stable 21 sludge Low sludge handling costs
Disadvantages
High operation costs, effluent quality variation
Complex and expensive sludge handling costs
Large reactor, high oxygen demand, high capital cost
COD removal, biological N removal, biological P removal
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Biological Nutrient Removal
regions, where wastewater temperatures can exceed 25–30 1C, short sludge systems would normally nitrify; in fact, it would be difficult to stop them doing so. For these situations, it is best to accept nitrification as inevitable and design the system accordingly. Furthermore, it would be advantageous to include a small primary anoxic zone (B15–20% anoxic mass fraction, see Section 4.14.24) in the system to denitrify a considerable proportion of the nitrate generated even if N removal is not required – this increases the minimum sludge age for nitrification, reduces oxygen demand, recovers alkalinity, and reduces the risk sludge flotation and high effluent COD due to denitrification on the SST bottom. Biological P removal is possible at short sludge ages of 3–5 days – the PAOs are relatively fast-growing heterotrophs (compared with the ANOs). In the absence of nitrification, an unaerated zone would be anaerobic (i.e., no nitrate or oxygen present or entering it) and provided the readily biodegradable (RB) COD and short-chain fatty acids (VFAs) are available from the influent, BEPR will take place. The original Phoredox system developed by Barnard (1976) is based on a two-reactor anaerobic–aerobic system. The minimum sludge age for BEPR is temperature dependent, increasing as temperature decreases and is around 3–5 days at 14–20 1C (Mamais and Jenkins, 1992). At these temperatures, the minimum sludge age for nitrification is significantly longer than that for BEPR, so that nitrification generally would not take place with the result that the adverse effect of nitrate on the BEPR would be absent. However, in warmer climates the minimum sludge age for nitrification and BEPR is similar, and ensuring a low nitrate recycle to the anaerobic reactor by including also anoxic zones is essential if BEPR is required (Burke et al., 1989). If BEPR is not required, the nitrification changes the two-reactor unaerated – aerated system from a P removal one to an N removal one.
4.14.15.1.2 Aerated lagoons Aerated lagoons, as distinct from aerated oxidation ponds in which oxygenation is supplemented by algae, are essentially high-rate AS systems because the oxygen demand is supplied wholly by aerators. There are essentially two types of aerated lagoons, suspension mixed and facultative. Suspension mixed aerated lagoons have sufficient energy input per unit volume by the aeration equipment to keep the sludge in suspension. In facultative lagoons this energy input is insufficient and settlement of solids onto the lagoon floor takes place. The biodegradable solids in the sludge layer so formed degrade anaerobically, as in an oxidation pond. Kinetically, suspension mixed lagoons are flow-through AS systems, and can be modeled as such. Their nominal hydraulic retention time equals their sludge age, and the waste (Qw) and effluent (Qe) flows are one and the same and equal to the influent flow (Qi). Hence, the volume of the aerated lagoon per unit COD load is very large compared with the conventional short sludge age systems, which have hydraulic retention times about 1/20th of the sludge age. The effluent from a suspension mixed aerated lagoon has the same constituents as the mixed liquor in the lagoon. The COD removed from the system via the oxygen demand is relatively small so that the COD in the effluent is generally
unacceptable for discharge to receiving waters. In fact, the principal objective of all short-age plants is to act as biologically assisted flocculators, which (biologically) transforms the influent soluble biodegradable organics to settleable organism mass and enmesh with this the influent BPO and UPOs to a form a settleable sludge that allows effective liquid– solid separation. In conventional short sludge age plants, the waste sludge is transferred to the sludge treatment facility; in the aerated lagoon systems, the effluent (with the waste sludge) usually flows to a second pond, that is, an oxidation pond or a facultative aerated lagoon, to allow the now readily settleable particulate material to settle to the lagoon floor to produce a relatively solids free and low COD effluent. The sludge that accumulates on the tank floor undergoes anaerobic stabilization. Aerated lagoons find application principally as low-technology industrial waste-treatment systems where organic strengths are high, the load varies seasonally, and nitrification is not required.
4.14.15.2 Intermediate Sludge Ages (10–15 Days) Where nitrification is obligatory because of a low effluent FSA concentration standard, this will govern the minimum sludge age of the AS system. For nitrification, the sludge ages required are 5–8 times longer than those for COD removal only, depending on the temperature. In the temperate regions where water temperatures may fall below 14 1C, the sludge age is not likely to be less than 10–15 days, taking due consideration of some unaerated zones in the reactor for denitrification (and biological P removal). In this range of sludge age, the effluent COD concentration no longer plays a role in the design. For sludge ages longer than about 4 days, protozoan organism predation of free swimming bacteria is high and flocculation good so particle dispersion is low. Also, virtually all soluble biodegradable organics are broken down, with the result that the effluent COD (or BOD) concentration remains approximately constant at its lowest achievable value, that is, the unbiodegradable soluble COD concentration. The effluent ammonia concentration also plays a minor role in design because the nitrification kinetics are such that once nitrification is achieved, it is virtually complete provided sufficient oxygen is supplied. Even though the effluent standards may require an effluent ammonia concentration, say o10 mgFSAN l1, once nitrification takes place the concentration is not likely to be greater than 2–4 mgN l1. Consequently for nitrification, the sludge age of the system is fixed principally by the requirement for nitrification. The method for calculating the minimum sludge age for nitrification is given in Section 4.14.20.3. Once a sludge age of say 25% longer than the minimum is selected, the effluent FSA concentration is mostly affected by the system operating conditions than by the nitrification process itself, that is, oxygen supply limitations, variation in ammonia load, uncontrolled loss of sludge, and pH of the mixed liquor. With low alkalinity wastewaters, nitrification can cause a significant reduction in effluent pH, often as low as 5. This not only causes problems with the nitrification process itself, that is, noncompliance with the effluent ammonia standard, but
Biological Nutrient Removal
also produces aggressive effluents that can do considerable damage to concrete surfaces. To reduce these problems and derive the other advantages of oxygen and alkalinity recovery (see Section 4.14.24.2), the policy of deliberate biological denitrification is advocated whenever nitrification is likely, even if N removal is not required. However, once nitrification is required and biological denitrification is incorporated in the system, sludge ages longer than 10–15 days may be required and the system falls into the long sludge age category. In nitrifying aerobic AS plants, there is always the possibility of denitrification in the SST. This problem is exacerbated by the system control procedure of abstracting the waste sludge from the settling tank underflow (Section 4.14.14.2). At low underflow recycle ratios, sludge retention in the SST is long leading to denitrification (Figure 19(b)). Henze et al. (1993) estimated that between 6–8 and 8–10 mgN l1 nitrate needs to be denitrified to cause sludge flotation at 10 and 20 1C, respectively. The concentration of nitrate denitrified increases as (1) sludge retention time in the SST increases, which is dependent on the recycle ratio and peak flow conditions; (2) active fraction of the sludge increases, that is, greater at shorter sludge ages (Figure 15); (3) temperature increases; and (4) mass of unutilized enmeshed biodegradable organics increases which is higher at shorter sludge ages and greatest at the peak load condition (Ekama et al., 1997). The above demonstrates that, for plants where nitrification takes place, the SST should not serve the dual purpose of solid–liquid separation and waste sludge thickening; the hydraulic control of sludge age should be employed; and deliberate denitrification should be included in the system (Section 4.14.14.2). These modifications will ameliorate the problem of sludge flotation by denitrification in the SST, but may not completely eliminate the root cause, that is, high nitrate concentrations in the mixed liquor. In order to reduce the construction cost of the AS system, reductions sludge age need to be made. Moreover, a reduction in sludge age also increases both biological N and P removal per mass organic load (WRC, 1984; Wentzel et al., 1990) and this would be particularly beneficial for low temperature wastewaters (10–15 1C) where nitrification is required. To try to reduce the sludge age required for nitrification, and hence the biological reactor volume per Ml WW treated, internal fixed media have been placed in the aerobic reactor (Wanner et al., 1988; Sen et al., 1994; Ekama and Wentzel, 1999b). The nitrifiers that grow on the fixed media are not subject to the mixed liquor sludge age and aerobic mass fraction with the result that both can be reduced. However, the effectiveness of the internal fixed media has not been as good as expected, and they yield a rather low cost/benefit ratio. Successful reduction of sludge age down to 8–10 days has been achieved with external nitrification (Bortone et al., 1996; Sorm et al., 1996; Hu et al., 2000, 2001) and this system starts finding application at full scale (Vestner and Gu¨nthert, 2001; Muller et al., 2006). With external nitrification, the nitrification process is removed completely from the suspended AS and transferred to an external fixed medium system like a trickling filter. With nitrification independent of the BNRAS mixed liquor, the sludge age can be reduced to around 8–10 days. Such a reduction reduces the biological reactor volume requirement per M l WW treated by about a one-third
459
without negatively impacting either biological N or P removal. Moreover, the sludge settleability improves significantly (DSVIB60–80 ml g1) compared with conventional BNR systems, which further increases the capacity of the system (Hu et al., 2000). Source separation of urine may also produce wastewaters, comprising feces flush water (brown), bathroom and kitchen water (gray), with sufficiently low influent TKN concentrations to obviate nitrification in the WWTP (Wilsenach, 2006; Mbaya et al., 2010). Comparing intermediate sludge age plants with high rate plants, the oxygen demand per kgCOD (including nitrification) is doubled (except with external nitrification, for which it is halved), the system volume is 3–4 times larger, the daily sludge mass wasted is reduced by 40%, and active fraction is much lower. Intermediate sludge age plants are much more stable than high-rate plants, requiring less sophisticated control techniques or operator intervention (excepting external nitrification), thereby making these plants more suitable for general application. At intermediate sludge ages, the active fraction of the waste sludge is still too high for direct discharge to drying beds. Consequently, some form of waste sludge stabilization would need to be incorporated in the WWTP, for example, aerobic or anaerobic digestion. The former has the advantage of ease of operation but the disadvantage of energy costs for oxygen supply; the latter has the advantage of energy generation from the biogas but the disadvantage of complexity of operation. Even with energy recovery by anaerobic digestion of waste sludge, because of the low mass of sludge wasted from the AS plant and high oxygen demand per kgCOD load, energy self-sufficiency at intermediate sludge ages is not possible. However, on large plants (B500 000 PE) where technical supervision and operator expertise are of a high level, energy costs can be reduced by gas production from AD, particularly if energy costs continue to increase as they have over the past decade. Ekama (2009) found that the green house gas emission (CO2) from two widely differing WWTPs treating the same wastewater is virtually the same if the residual biodegradable organics (COD) in the final disposed sludge is the same, viz., (1) a long sludge age (30 days) extended aeration AS system treating raw wastewater and (2) a short sludge age (8 days) AS system treating settled wastewater with anaerobic digestion of primary sludge and aerobic or anaerobic digestion of WAS with beneficial combustion/flaring of methane gas. In any event, the CO2 emitted by WWTPs (B20 g CO2-C/PE/d), together with that generated by their energy consumption at a fossil-fuel power station (B40 g CO2-C/PE/d at 800 kWh Ml1) is very low in comparison with that generated by (1) domestic energy consumption (B4000 gCO2-C/PE/d at 10 kWh/PE/d), which is only one-third of the total PE energy consumption, (2) motor car driving (B2000 gCO2-C/PE/d and 30 km/PE/d) or (3) even just breathing (B180 gCO2-C/ PE/d for a 6000 kJ d1 diet). So, from a sustainability point of view, the treated water produced at the WWTP has a far greater value than trying to maximize energy recovery from WWTP at the expense of effluent quality. ‘‘Minimization of energy requirement for wastewater treatment is an important goal but has a lower priority than the human and environmental health which is closely related to efficient water quality management’’ (Svardal and Kroiss, 2009).
460
Biological Nutrient Removal % Biodegradability of sludges residual biodegradable organics (COD)
% Biodegradability
80
4.14.15.3.2 Anoxic–aerobic plants
70
Particulate
60
Soluble
50 40 30 20 10 0
1
2
3
4
5 6 7 Sludge type
8
9
10
Figure 21 % residual biodegradable organics remaining in stabilized wastewater sludges treated with different stabilization system types: (1) Raw unsettled wastewater; (2) Zimpro humus þ 11 high soluble COD; (3) anaerobically digested 11 þ WAS high VFA; (4) anaerobically digested 11 only high VFA; (5) anaerobically digested 11, 1st stage low VFA; (6) Zimpro humus þ 11 low soluble COD; (7) anaerobically digested 11, 2nd stage low VFA; (8) DAF thickened WAS; (9) an. digested 11 þ WAS, single stage low VFA; and (10) aerobically digested WAS. 11, primary sludge; VFA, volatile fatty acids; WAS, waste activated sludge.
4.14.15.3 Long Sludge Ages (20 Days or More) 4.14.15.3.1 Aerobic plants Long sludge age aerobic plants are usually called extended aeration plants. The principal objective of long sludge systems is to obviate primary (11) and secondary (21) sludge treatment. These plants therefore treat raw wastewater and the sludge age is chosen so that the active fraction (or residual biodegradable organics) of the waste sludge is sufficiently low to allow its direct discharge to sludge drying beds. The sludge age required to produce a sludge sufficiently stable so as not to generate odor problems is uncertain and will depend on the temperature and climatic conditions, that is, whether or not the sludge can be dried sufficiently quickly before it starts stinking, but probably exceeds 30 days. Interestingly, from a survey of the residual biodegradable organics in wastewater sludges treated by different sludge stabilization systems, Samson and Ekama (2000) found that aerobically digested WAS contained the lowest residual biodegradable organics (10%) compared with wet air oxidized (Zimpro) and anaerobically digested primary sludges (25–60%, Figure 21). Extended aeration plants are very stable in operation and require less supervision than their short sludge age counterparts. Although the volume requirements and oxygen demand per unit COD are very high, the relative ease of operation makes the system the preferred one for small communities. The reactor configurations that can be operated in the extended aeration mode are many and the particular one that suits the specific application is chosen by the designer. These include single completely mixed reactor systems, Orbal, oxidation ditch and Carousel multichannel systems, and multireactor anaerobic anoxic and aerobic systems for BNR.
Once the sludge age exceeds 20–25 days, nitrification is inevitable and is advisable for reasons cited above to incorporate denitrification in the system, which at these long sludge ages would not affect the stability of nitrification. Furthermore, if required, BEPR can also be included for little extra cost. In fact, biological N and P removal are significantly greater with raw wastewater than the settled wastewater due to the higher organic load. To include N (and P) removal, the reactor is subdivided into unaerated (anoxic and anaerobic) and aerated zones in a variety of configurations. Denitrification takes place in the unaerated but mixed zones receiving nitrified mixed liquor via recycles from the aerated zones to give the socalled ND systems. The ND systems include: the four stage Bardenpho, which incorporates primary and secondary anoxic reactors; the modified Ludzack Ettinger (MLE), which incorporates only a primary anoxic reactor; the Orbal, Carousel, and oxidation ditch systems in which the anoxic zones are created along different lengths of the same long channel reactor; and the intermittently decanted extended aeration (IDEA) systems in which aerators are swithched off for different time periods over the day. Although incorporation of denitrification imposes some additional constraints on the design, at long sludge age, these are minor provided the aeration capacity of the plant is sufficient to ensure efficient nitrification under all expected conditions (Section 4.14.20).
4.14.15.3.3 Anaerobic–anoxic–aerobic plants When the BEPR is required, an initial anaerobic reactor is included in the configuration that receives the influent wastewater but minimal oxygen and nitrate via the sludge recycles. For BEPR, assurance of a zero nitrate discharge to the anaerobic zone is critical for achieving good P removal and is an additional constraint on the design when including BEPR in extended aeration systems. The extent of BEPR achieved will depend on a number of factors, mainly the influent RB COD concentration, the TP/COD ratio, and the degree to which nitrate can be excluded from anaerobic reactor, which depends on the influent TKN/COD ratio. The waste sludge from extended aeration systems, including BEPR, has the potential to release high P concentrations. This can be dealt with in specially designed dewatering/drying beds with sand filter under drains and weir overflows, which allow the drying bed also to operate as a dewatering system. While discharging waste sludge directly to the drying bed, the under drain and overflow are monitored for P concentration (with kit dipsticks) and when this gets to say 5 mgP l1, sludge wastage to the drying bed and the return of supernatant to the head of the works is stopped. The relatively small volume of high P liquor that drains from the drying bed thereafter is either chemically treated or irrigated at the WWTP site. The dewatering capability of the drying bed allows significantly more sludge to be discharged to it than drying beds without these dewatering features.
4.14.16 Sludge Age – The Dominant Driver for Size For organic material (C) removal only, the sludge age of the system is short and hence the reactor volume small
Biological Nutrient Removal Dominant drivers for size of activated sludge system
(8) Sludge settleability
Apply uncertainty/ sensitivity analysis to 1−8 O r g a n i c
N I T
N D
E B P R
Kinetic parameters (change only rarely)
(1−7) Wastewater characteristics
(6) Influent RBCOD fraction
(7) Influent P load FP ti − kgP/d
Clarifier models
FX t
Oxygen demand FO2 − kgO2 d−1
(1) organic load FS ti − kgCOD/d (2) Unbio part COD fraction − f S’up (3) Nitrogen load FN ti − kgTKN-N/d (4) Nitrifier max (5) WW temperature
Reactor conc. (Xt)
461
Reactor volume (Vp)
FO2
SST area (A sst)
Mass TSS in reactor (MX t) Sludge production (FXt − kg TSS d−1) Reactor volume (V p) S L U D G E
A G E
f xa
SST area (Asst)
f xd Recycle: high (1:1)
Anoxic mass fraction (f xd)
Effluent NO3
Anaerobic and anoxic mass fractions (f xm = f xa + f xd) Effluent NO3 and PO4
Figure 22 Important wastewater characteristics required to be known for different activated sludge systems – fully aerobic, nitrification, nitrification– denitrification, and biological excess P removal – and the interrelationships that affect sludge age and effluent quality.
(Figure 22). Essentially, only the organic (COD) load and unbiodegradable particulate (fS’up) and soluble (fS’up) COD fractions need to be known (Table 9). The organic load and unbiodegradable particulate COD concentration (Supi) strongly affect sludge mass in the reactor and daily sludge production. The unbiodegradable soluble COD concentration fixes the filtered effluent COD concentration from the system. Also, the organic load and sludge age fix the daily oxygen demand (kgO d1) and the peak hydraulic load fixes the secondary settling tank surface area and peak oxygen demand. If nitrification is required from the system, more wastewater characteristics are required to be known (Table 9). The most important of these are the maximum specific growth rate of the nitrifiers at the standard wastewater temperature of 20 1C (mAm20) and the minimum wastewater temperature (Tmin), both of which fix the minimum sludge age for nitrification (Rsm). The system sludge age (Rs) must be selected longer than the minimum for nitrification and the higher the system to minimum sludge age ratio (Rs/Rsm), the lower the effluent ammonia concentration and the more damped its variation in response to diurnal nitrogen load variation. Also required for nitrifying systems is the daily nitrogen load (both TKN and FSA) so that the components making up the N material in the influent can be determined. For nitrification, the maximum specific growth rate of the nitrifiers is regarded a wastewater characteristic and not a model kinetic constant because it is different in different wastewaters.
With biological nitrogen removal (ND), a part of the biological reactor volume is intentionally not aerated. The sludge mass in the unaerated (anoxic) reactor as a fraction of the sludge mass in the whole reactor is the anoxic mass fraction (fxd, Figure 22). The larger the anoxic mass fraction, the more nitrate can be denitrified but the longer the minimum sludge age for nitrification becomes over that for fully aerobic conditions. So for ND systems, the biological reactor gets larger because the required sludge ages get longer. Also an additional wastewater characteristic needs to be known, that is, the influent RB COD concentration (or fraction, fS’bs) because a high proportion (up to half) of the nitrate denitrified in the primary anoxic reactor is due to this wastewater constituent – if the influent RBSO (COD) concentration is not known, the effluent nitrate concentration cannot be calculated accurately with either steady-state or dynamic simulation models. With BEPR, the daily wastewater phosphorus load (both total P and orthoP) needs to be known so that the components making up the P material in the influent can be determined. With BEPR the influent RBSO concentration (which includes the short-chain fatty acids, VFA) is very important and establishes the extent of biological P removal that can be achieved. If the influent RBSO concentration is not known, the biological P removal that can be achieved cannot be calculated accurately. The influent RBSO is indirectly the food source for the PAOs that mediate the BEPR process. The
462
Biological Nutrient Removal
Table 9 Wastewater characteristics requiring specification for different single sludge activated sludge systems. Characteristics marked NB (nota bene) are very important and required to be accurately known for accurate design. Wastewater characteristics
Units
Symbol
System type Fully aerobic
1. 2. 3.
4.
5. 6.
7. 8. 9.
Mean influent COD concentration Average flow Average temperature Maximum Minimum Influent COD fractions Unbiodegradable soluble (USO) Unbiodegradable particulate (UPO) Readily biodegradable (RBSO) Fermentable readily biodegradable Volatile fatty acids (VFAs) Mean influent TKN concentration Influent TKN fractions Ammonia Soluble unbiodegradable organic N Influent total P concentration ANO max. specific growth rate at 20 1C Influent inorganic suspended solids
Non Nit.
Nit.
ND
NDBEPR
mgCOD l1 l d1
Sti Qi
|NB |NB
|NB |NB
|NB |NB
|NB |NB
1C 1C
Tmax Tmin
| |
| |NB
| |NB
| |NB
mgN l1
fS0 us fS0 up fS0 bs fSbs0 f fSbs0 a Nti
| | -
| | |NB
| | |NB |NB
| | |NB | | |NB
mgP l1 d1 mgISS l1
fN0 a fN0 ous Pti mAm20 XIOi
|
| | |NB |
| | |NB |
| | |NB |NB |
purpose of the anaerobic zone, which receives the influent wastewater, is to allow the PAOs to take up the VFA fermentation products generated from the influent RBSO. Nitrate (or DO) which enters the anaerobic zone results in utilization of some of the influent RBSO by OHOs, which reduces the VFA products available to the PAOs and hence the biological P removal. The difference between the influent P concentration and the BEPR that can be achieved establishes the effluent P concentration. Very low nitrate (and DO) concentrations in the recycles entering the anaerobic zone are essential for maximum BEPR. This imposes important requirements on the denitrification required in the anoxic zones. If the influent TKN/COD concentration ratio is too high, then low nitrate concentrations cannot be achieved in the anoxic zone(s) and methanol dosing may be required. High N removals in the anoxic zones requires large anoxic reactor(s), which together with the anaerobic zone results in large unaerated mass fractions, which in turn requires long sludge ages to ensure nitrification. Unless specific strategies are applied to keep the sludge age low, such as external nitrification (Sorm et al., 1996; Hu et al., 2000; Muller et al., 2006) or adding fixed media into the aerobic zone (Wanner et al., 1988; Sen et al., 1994) to reduce the system sensitivity to the minimum sludge age for nitrification, NDBEPR systems will have long sludge ages, especially where wastewater minimum temperatures are low. The above demonstrates that wastewater characteristic determination is the most important aspect of modeling WWTPs, whether using steady-state or dynamic simulation models. Uncertainty in wastewater characteristics (and sludge settleability) results in a commensurate uncertainty in calculated oxygen demand, sludge production, reactor volume, and effluent quality. So, uncertainty/sensitivity analyses should be applied to the wastewater characteristics rather than to the
kinetic and stoichiometric parameters of the model(s). In fact, only rarely should the kinetic and stoichiometric parameters of the model be changed (except the maximum specific growth rate of nitrifiers which is regarded a wastewater characteristic). Fitting all the effluent quality concentrations, sludge production, and oxygen demand to laboratory, pilotand full-scale plant data can be achieved by changing the wastewater characteristics only, provided the data conform to mass balances (water, COD, N, and P). More often than not, model predictions cannot be made to conform to measured data because the measured data do not conform to mass balance and continuity principles. Only when the data conform to mass-balance and continuity principles and changing the wastewater characteristics cannot yield a good correlation between model predictions and measured data, should kinetic and stoichiometric parameters of the model be changed, but such change(s) should be based on bioprocess basics and not simply because ‘it makes the model fit’.
4.14.17 Nitrification – Introduction The term nitrification describes the biological process whereby FSA is oxidized to nitrite and nitrate. Nitrification is mediated by specific chemical autotrophic organisms with behavioral characteristics that differ significantly from the heterotrophic (OHO) ones. Whereas the OHOs obtain their carbon (anabolism) and energy (catabolism) requirements for biomass synthesis from the same organic compound(s), the autotrophic nitrifying organisms obtain their carbon requirement (anabolism) from dissolved CO2 and their energy requirement (catabolism) for biomass synthesis from oxidizing ammonia to nitrite and nitrite to nitrate. This difference results in the autotrophic nitrifiers having much lower biomass growth
Biological Nutrient Removal
coefficients (one-fifth) than the OHOs. The objectives in this chapter are to review briefly the kinetics of nitrification, to highlight the factors that influence this biological process, and set out the procedure for designing a nitrifying aerobic or anoxic-aerobic AS system. It has been well established that nitrification is due to two specific genera of autotrophic bacteria, the ammonia oxidizing organisms (ANOs) and the nitrite oxidizing organisms (NNOs). Originally, it was thought that only nitrosomonas and nitrobacter mediated nitrification but recent molecular techniques have shown that there are several genera of nitrifying organisms. Nitrification takes place in two sequential oxidation steps: (1) ANOs convert FSA to nitrite and (2) NNOs convert nitrite to nitrate. The nitrifiers utilize ammonia and nitrite principally for synthesis energy requirements (catabolism) but some ammonia is also used anabolically for synthesis of cell mass nitrogen requirements. The ammonia requirement for synthesis, however, is a negligible fraction of the total ammonia nitrified to nitrate by the nitrifiers, at the most 1%. Consequently, in steady-state models it is usual to neglect the synthesis nitrogen requirements of the nitrifiers and to consider the nitrifiers simply to act as biological catalysts in the nitrification process. This stoichiometric approach greatly simplifies the description of the kinetics of the process. The two basic stoichiometric redox reactions in nitrification are:
NH4 þ þ ð3=2ÞO2 ðANOsÞ-NO2 þ H2 O þ 2Hþ
NO2 þ ð1=2ÞO2 ðNNOsÞ-NO3
ð118aÞ
463
nitrification sequence is therefore the ammonia conversion to nitrite by the ANOs. So from a steady-state modeling point of view, one needs to consider the kinetics of this organism group only. Because the nitrite produced is virtually immediately further nitrified to nitrate, it is assumed that the ANOs nitrify ammonia to nitrate directly and the kinetics of nitrification reduce to the kinetic behavior of the ANOs. Experimental investigations by Downing et al. (1964) showed that the nitrification rate can be formulated in terms of the Monod equation. In fact, Monod kinetics was applied to nitrification before it was applied to model the kinetics of organic material breakdown by heterotrophic organisms. The successful application to nitrification prompted Lawrence and McCarty (1972) to apply it to AS. Monod established that (1) the mass of organisms generated is a fixed fraction of the mass of substrate (in this case ammonia) utilized and (2) the specific rate of growth, (i.e., the rate of growth per unit mass of organisms per unit time) is related to the concentration of substrate surrounding the organisms. From (1),
MDXBA ¼ YA MDNa
ð119Þ
where MDXBA is the mass of nitrifiers generated (mgVSS), MDNa the mass of ammonia as N utilized (mgFSA-N), and YA the nitrifier yield coefficient mgVSS/mgN. Taking the changes over a time interval Dt and assuming the changes are very small, one can write
dXBA dNa ¼ YA ðmgANOVSS l1 d1 Þ dt dt
ð120Þ
ð118bÞ
Stoichiometrically, the oxygen requirements for the first and second reactions are 3/2 32/14 ¼ 3.43 and 1/2 32/ 14 ¼1.14 mgO/mgN (also written as mgO/mgFSA-N). Hence, the stoichiometric conversion of ammonia to nitrate, both expressed as N, requires 2 32/14 ¼ 4.57 mgO/mgN utilized. Taking into account the ammonia utilized for synthesis of nitrifier cell mass, the oxygen requirement per mgFSA-N nitrified is slightly less, with reported values down to 4.42 mgO/ mgFSA (Ekama, 2009). This approach is adopted in the simulation models such as ASM1 (Henze et al., 1987) and is one reason for the small difference in the predicted results between steady-state stoichiometric models and the more complex simulation models.
4.14.18 Nitrification Biological Kinetics 4.14.18.1 Growth In order to formulate the nitrification behavior, it is necessary to understand the basic biological growth kinetics of ANOs. The rate of conversion of ammonia to nitrite, by the ANOs is generally much slower than that of nitrite to nitrate by the NNOs. Therefore, under most circumstances in municipal WWTPs, any nitrite that is formed is converted virtually immediately to nitrate. As a consequence generally very little nitrite (o1 mgN l1) is observed in the effluent from a plant operating on an influent that does not contain substances that inhibit the NNOs. The limiting rate in the two-step
From (2) Downing et al. (1964) developed the following relationship, known as the Monod equation
mA ¼
mAm Na Kn þ Na
ðmgVSS=mgVSS dÞ
ð121Þ
where mA is the specific growth rate at ammonia concentration Na (mgANOVSS/mgANOVSS/d), mAm the maximum specific growth rate (mgANOVSS/mgANOVSS/d), Kn the halfsaturation constant, that is, the concentration at which mA ¼ 1/2 mAm (mgN l1), and Na the bulk liquid ammonia concentration (mgN l1). The Monod constants maximum specific growth rate mAm and half-saturation coefficient (also known as the affinity coefficient) Kn for the ANOs are sensitive to temperature, generally decreasing as temperature decreases. An additional subscript T on the symbols refers to temperature (1C). The growth rate is given by the product of the specific growth rate and the ANO concentration (XBA):
dXBA mAmT Na ¼ mAT XBA ¼ XBA dt KnT þ Na ðmgANOVSS l1 d1 Þ
ð122Þ
The rate of ammonia conversion is found by combining Equations (120) and (122), viz.,
dNa 1 mAmT Na ¼ XBA YA KnT þ Na dt
ðmgFSA-N l1 d1 Þ
ð123Þ
Biological Nutrient Removal
dNn dNa 1 mAmT Na ¼ XBA dt dt YA KnT þ Na
ðmgNO3 -N l
1
1
d Þ ð124Þ
where Nn is nitrate concentration (mgNO3 N l1). The oxygen utilization rate associated with nitrification is based on the stoichiometric oxygen requirement of 4.57 mgO/ mgFSA-N nitrified to nitrate calculated above, viz.,
dOn dNa dNn ¼ OURn ¼ 4:57 ¼ 4:57 dt dt dt
0.5
5
UAm or KAm
4
0.4 UA = UAm Na /(Kn + Na) UAm = Maximum specific growth rate Kn = Half-saturation coefficient
0.3
3
2
0.2
0.1 Kn
UAm20 = 0.45/d; KAn = UA20 /YA YA = 0.10 mgVSS/mgN Kn = 1.0 mgN l−1
1
0.0
ðmgO l1 d1 Þ
0 0
ð125Þ Assuming stoichiometric conversion of FSA to nitrate as in Equations (124) and (125) slightly overestimate the nitrate generation and oxygen consumption because a small proportion (1%) of the FSA taken up by the nitrifiers is used for cell synthesis. Based on the empirical organism cell mass formula C5H7O2N, Ekama (2009) shows that for 1 mgFSA-N taken up, 0.99 mgN nitrate and 0.076 mgANOVSS is generated and 4.42 mgO is utilized. Application of the Monod growth kinetics to nitrification by Downing et al. (1964) is probably one of most successful applications of microbiological kinetic research to wastewater treatment, so much so that the Monod kinetics is commonly used today to express the rates of many biological processes in terms of the growth limiting nutrient concentrations. Monod growth kinetics requires three constants to be known: the yield coefficient (YA), the maximum specific growth rate (mAm), and the half-saturation coefficient (Kn). The yield coefficient for nitrifying organisms represents the net organism mass produced per unit mass of substrate nitrogen utilized. Evidence that this coefficient is not constant but can vary with the conditions of growth was presented in the 1960s when the nitrification model was developed. However, Downing et al. (1964) stated that the different VSS concentrations obtained from different YA values are inconsequential to the experimentally determined maximum specific growth rate, mAm, provided a consistent pair of mAm and YA are used. This is because the mAm is obtained from an observed maximum specific nitrification rate, KAm mgFSA-N nitrified/(mgANOVSS d), which is equal to mAm/YA. If YA is selected low, the mAm will be high and vice versa. To avoid confusion about the experimentally determined mAm rates, a standard YA ¼ 0.10 mgVSS/mgFSA or 0.15 mgCOD/mgFSA has been adopted in steady-state and dynamic simulation AS models for municipal WWTPs.
4.14.18.2 Growth Behavior In Figure 23 the relationship between the specific growth rate, mA, the specific substrate (FSA) utilization or nitrification rate, KA, and the bulk liquid FSA concentration, Na, is shown, as described by the Monod equation (Equation (122)). The rate constants selected are mAm20 ¼ 0.45 d1, YA ¼ 0.10 mgANOVSS formed/mgFSA-N nitrified, making KAm ¼ 4.5 mgFSA-N/
Specific nitrification rate
Because in the steady-state model the nitrification process is accepted to be stoichiometric, that is, the nitrifying organisms act only as a catalyst to the process, the rate of nitrate formation is equal to the rate of FSA conversion, that is,
Specific growth rate (d−1)
464
10 20 30 40 Ammonia concentration (mgN I−1)
50
Figure 23 The Monod specific growth rate equation for nitrification at 20 1C.
(mgANOVSS d) and Kn20 ¼1.0 mgN l1. The interesting feature of this nitrifier growth behavior is that, because Kn is so low at B1 mg(FSA-N) l1, the nitrification rate is virtually at maximum for concentrations 42 mgFSA-N l1. However, at concentrations o 2mgN l1, the rate rapidly declines to zero. The implication of this is that when nitrification takes place, it will be nearly complete (provided all other requirements are met – see below) but the ammonia concentration is not readily reduced to zero.
4.14.18.3 Endogenous Respiration It is generally accepted that all organisms undergo some form of biomass loss due to maintenance or endogenous energy requirements. This behavior manifests when a biomass has completely utilized its external substrate – its VSS decreases and it continues to utilize oxygen with time. This process is called endogenous respiration. Different organisms have different endogenous respiration rates. For the OHOs, it is quite high (bH20 ¼ 0.24 d1), whereas for the nitrifiers (ANOs), it is low (bA20 ¼ 0.04 d1). The endogenous respiration process for the ANOs is modeled in exactly the same way as that for the OHOs, that is,
dXBA ¼ bAT XBA dt
ðmgANOVSS l1 d1 Þ
ð126Þ
where bAT is the specific endogenous mass loss rate for nitrifiers at T (1C), mgANOVSS/(mgANOVSS d).
4.14.19 Nitrification Process Kinetics The basic AS system modeled for nitrification is the single completely mixed reactor system with hydraulic control of sludge age (see Figure 2). This system under steady-state conditions provides the information necessary for design of nitrification. The principal steady-state solution required for this is the effluent ammonia concentration (Nae). This solution forms the basis for the analysis of the nitrification process behavior and provides the information for the design of an AS
Biological Nutrient Removal
465
system including this process. This information is also sufficient to understand the modeling of the nitrification process in AS simulation models such as ASM1.
Table 10 Kinetic constants and their temperature sensitivity for ANOs accepted in most activated sludge models Kinetic constant
at 20 1 C
Temp. coeff.
4.14.19.1 Effluent Ammonia Concentration
Yield coefficient, YA (mgVSS/mgFSA) Endogenous respiration rate, bA (d1) Half-saturation coefficient, Kn (mgFSA l1) Maximum specific growth rate mAm (d1)
0.1 0.04 1 Varies
1 1.029 1.123 1.123
A mass balance on the change in nitrifier mass MDXBA over the completely mixed system at steady state is given by
MDXBA ¼ Vp DXBA ¼
mAmT Na XBA Vp Dt bAT XBA Vp Dt KnT þ Na XBA Qw Dt ðmgANOVSSÞ
60 Rs = Rsm
where Vp is the reactor volume (l) and Qw the waste sludge flow rate from the reactor (l d1). Dividing by VpDt yields
ð127Þ
Under steady-state (constant flow and load) conditions, DXBA/ Dt is zero and from Equation (58), Qw/Vp ¼ Rs. Substituting these and solving for the reactor ammonia concentration (Na), and therefore also from the definition of completely mixed conditions, the effluent ammonia concentration (Nae) yields
50 Ammonia (mgN I−1)
DXBA mAmT Na Qw XBA bAT XBA XBA ¼ Dt KnT þ Na Vp
Rs < Rsm
Rs > Rsm Am20 = 0.33 d−1 Kn20 = 1.0 mgN I−1
40
Influent ammonia concs
30
Rsm = For different influent ammonia concentrations
20
10
KnT ðbAT þ 1=Rs Þ Na ¼ Nae ¼ mAmT ðbAT þ 1=Rs Þ
1
ðmgN l Þ
ð128Þ 0
From Equation (128), the ammonia concentration (Na) in the reactor and effluent (Nae) are independent of the specific yield coefficient (YA) and the influent ammonia concentration (Nai). Using mAm20 ¼ 0.33 d1 and Kn20 ¼1.0 mgN l1 at 20 1C, and taking bAT ¼ 0.04 d1 (Table 10), a plot of Equation (128) with Nae versus sludge age Rs is given in Figure 24. At long sludge ages Nae is very low and remains so until the sludge age is lowered to about 4 days. Below 4 days, Nae increases rapidly and in terms of Equation (128) can exceed the influent FSA concentration, Nai. This clearly is not possible so the limit of validity of Equation (128) is Na ¼ Nai. Substituting Nai for Na in Equation (128) and solving for Rs give the minimum sludge age for nitrification, Rsm below which theoretically, nitrification cannot be achieved, that is,
Rsm ¼
1 ½mAmT =ð1 þ ðKnT =Nai ÞÞ bAT
ð129Þ
This minimum sludge age varies slightly with the magnitude of Nai (Figure 24) – higher Nai gives a slightly lower Rsm. The effect of Nai on RSm is very small because the magnitude of KnT is very small relative to Nai (o5%). So for Nai 420 mgN l1 (rarely will it be lower than this), and noting that Kn20B1 mgN l1, then KnT/Nai is negligibly small with respect to 1 (o5%). So substituting zero for KnT/Nai in Equation (129) yields
Rsm ¼
1 mAmT bAT
ðdaysÞ
ð130Þ
For all practical purposes, taking into account the uncertainty in mAm, Equation (130) adequately defines the minimum sludge age for nitrification. Conceptually, Equation (130)
0
2
4 6 Sludge age (days)
8
10
Figure 24 Effluent ammonia concentration vs. sludge age for the steady-state nitrification model.
states that if the net nitrifier multiplication rate (inverse of the net maximum specific growth rate, mAm bA) is slower than the harvesting rate of the nitrifiers via the sludge waste flow rate, then the nitrifiers cannot be sustained in the system and nitrification cannot take place. At sludge ages lower than the minimum for nitrification, nitrifiers are washed out of the system and so are called washout sludge ages. This concept of washout can be applied to any group of organisms in a bioreactor, and defines the sludge age below which the bioprocess will not take place because the organisms mediating this process are not sustained in the system. The virtually constant value for Rsm insofar as the influent FSA concentration is concerned (for the fixed values of mAmT and bAT) and the rapid decrease in effluent FSA concentration at sludge ages slightly longer than RSm is due to the very low Monod half saturation concentration for the nitrifiers (Kn20). This feature causes that in a particular plant, as the sludge age is increased, once Rs4Rsm, a high efficiency of nitrification will be observed, provided the FSA is the growth limiting nutrient for the ANOs, that is, all other requirements such as oxygen are met. Consequently, under steady-state conditions with increasing sludge age, kinetically, one would expect an AS system either not to nitrify at all, or, if it nitrifies, to nitrify virtually completely depending on whether the sludge age is
466
Biological Nutrient Removal
shorter or longer than the minimum (Rsm), respectively. Conversely, as sludge age decreases, one would expect an AS system to nitrify virtually completely and then quite suddenly cease to nitrify depending on whether the sludge age is shorter or longer than the minimum (Rsm), respectively. This behavior sometimes occurs in full-scale AS systems, where for many years the system has nitrified virtually completely, and suddenly one winter it stops nitrifying and produces very high effluent FSA concentrations. Provided the oxygen supply is not limiting, what happens in these situations is that over the years, the organic (COD) load on the system has increased and in order to maintain the reactor VSS concentration at the required level, the sludge wastage rate (Qw) has been increased, which reduced the sludge age. Then, coupled with a low winter temperature, the system sludge age falls below the minimum and nitrification ceases. This cannot happen with hydraulic control of sludge age, where a fixed proportion of the reactor volume is wasted daily to establish a constant sludge age. However, the secondary settling tank may become overloaded as the reactor TSS concentration increases with time, depending on the settleability with the AS (see Section 4.14.14). An operator therefore can choose the way an AS system fails with increasing organic loading – it does not have to be with nitrification, and so also with N removal.
4.14.20 Factors Influencing Nitrification From the discussion above, it can be seen that there are a number of factors that affect the nitrification process, the minimum sludge age required to achieve it, and the effluent FSA concentration from the AS system: 1. the magnitude of the kinetic constant mAm20 because this rate can vary considerably in different wastewaters; 2. temperature because it decreases the mAm20 rate and Kn20 coefficient; 3. unaerated zones in the reactor because ANOs are obligate aerobes and can grow only under aerobic conditions; 4. DO concentration because Monod kinetics presumes that FSA is the growth-limiting nutrient implying that the oxygen supply must be adequate; 5. cyclic flow and load conditions because FSA is dissolved and therefore the reactor (and effluent) FSA concentration is affected by the instantaneous actual hydraulic retention time; most FSA not nitrified during the actual hydraulic retention time escapes with the effluent; and 6. pH in the reactor because the mAm20 is strongly suppressed by pH outside the 7–8 range. These six factors are discussed further below.
4.14.20.1 Influent Source The maximum specific growth rate constant mAmT has been observed to be quite specific for the wastewater and also to vary between different batches of the same wastewater source. This specificity is so marked that mnmT should not be classified as a kinetic constant but rather as a wastewater characteristic. The effect appears to be of an inhibitory nature due to some substance(s) in the influent wastewater. It does not appear to
be a toxicity problem because a high efficiency of nitrification can be achieved even with a low mAmT value if the sludge age is increased sufficiently. These inhibitory substances are more likely to be present in municipal wastewater flows having some industrial contribution. In general, the higher the industrial contribution, the lower mAmT tends to be, but the specific chemical compounds that cause the reduction of mAmT have not been clearly defined. A standard temperature of 20 1C has been adopted for reporting mAm rates to take into account the effect of temperature. A range mAm20 values have been reported in the range of 0.30–0.75 d1 for municipal wastewaters. These two limits will have a significant effect on the magnitude of the minimum sludge age for nitrification. Two systems, having these respective mAm20 values, will have Rsm values differing by 250%. Clearly due to the link between the sludge age and mAmT, the latter’s value should always be estimated experimentally for optimal design. In the absence of such a measurement, a low value for mAmT necessarily will need to be selected to ensure that nitrification takes place. If the actual mAm is higher, the sludge age of the system will be longer and the reactor volume larger than necessary. However, the investment in the large reactor is not lost because in the future the plant will be able to treat a higher organic load at a shorter sludge age. Experimental procedures to determine mAm20 are given in the literature (e.g., WRC, 1984). The bn20 rate is taken as constant for all municipal wastewater flows at bn20 ¼ 0.04 d1. Its effect is small so that there is no need to enquire closely into all the factors affecting it. Little information on effects of inhibitory agents on KnT is available; very likely KnT will increase with inhibition.
4.14.20.2 Temperature The mAmT, KnT, and bAT constants are sensitive to temperature with a high-temperature sensitivity for the first two, while the endogenous rate is accepted to have the same low-temperature sensitivity as that for OHOs, viz.,
mAmT ¼ mAm20 ðyn ÞðT20Þ ðd1 Þ
ð131aÞ
KnT ¼ Kn20 ðyn ÞðT20Þ ðmgN l1 Þ
ð131bÞ
bAT ¼ bA20 ðyb ÞðT20Þ ðd1 Þ
ð131cÞ
where yn is the temperature sensitivity for nitrification ( ¼ 1.123) and yb the temperature sensitivity for endogenous respiration for ANOs ¼ 1.029. The effect of temperature on mAmT is particularly strong. For every 6 1C drop in temperature, the mAmT value halves which means that the minimum sludge age for nitrification doubles. Design of systems for nitrification, therefore, should be based on the minimum expected system temperature. The temperature sensitivity of KnT is also strong, doubling for every 6 1C increase in temperature. This does not affect the minimum sludge age for nitrification, but it does affect the effluent FSA concentration – the higher the Kn value, the higher the effluent FSA at Rs b Rsm. However, the faster mAmT rate at the higher temperature compensates for the higher KnT value so that the effluent FSA decreases with increase in temperature.
Biological Nutrient Removal
The effect of unaerated zones on nitrification can be formulated based on the following assumptions: 1. Nitrifiers, being obligate aerobes, grow only in the aerobic zones of a system. 2. Endogenous mass loss of the nitrifiers occurs under both aerobic and unaerated conditions. 3. The proportion of the ANOs in the VSS in the unaerated and aerated zones is the same so that the sludge mass fractions of the different zones of the system reflect the distribution of the nitrifier mass as well. From 1–3 above, it can be shown that if a fraction fxt of the total sludge mass is unaerated (i.e., (1 fxt) is aerated), the effluent ammonia is given by
KnT ðbAT þ 1=Rs Þ Nae ¼ mAmT ð1 f xt Þ ðbAT þ 1=Rs Þ
ð132Þ
Equation (132) is identical in structure to Equation (128), if one views the effect of the unaerated mass (fxt) as reducing the value of mAmT to mAmT(1 fxt), which conforms with (1) to (3) above. This sludge mass fraction approach is compatible with the nitrification kinetics in the AS kinetic models such as ASM1 and ASM2 (Henze et al., 1987, 1995) and UCTOLD and UCTPHO (Dold et al., 1991; Wentzel et al., 1992). In these models, nitrifier growth takes place only in the aerobic zone and endogenous respiration in all the zones. This sludge mass fraction approach is not compatible with the aerobic sludge age approach, which is used in some ND AS system design procedures (WEF, 1998; Metcalf and Eddy, 1991). In the aerobic sludge age approach, it is assumed that the growth and endogenous processes of the nitrifiers are active only in the aerobic zone, with neither processes active in the unaerated zone(s). This aerobic sludge age approach is not compatible with kinetic models and so significantly different predictions can be expected for the nitrification behavior from the aerobic sludge age-based design procedures and kinetic models. Following the same reasoning as that preceding Equation (132), it can be shown that the minimum sludge age for nitrification Rsm in an ND system having an unaerated mass fraction, fxt, is
Rsm ¼
1 mAmT ð1 f xt Þ bAT
ð133Þ
Alternatively, if Rs is specified, then the minimum aerobic sludge mass fraction (1 fxm) that must be present for nitrification to take place is found by substituting Rs for Rsm and fxm for fxt in Equation (133) and solving for (1 fxm), that is,
ð1 f xm Þ ¼ ðbAT þ 1=Rs Þ=mAmT
ð134Þ
or equivalently, from Equation (134), the maximum allowable unaerated sludge mass fraction at a sludge age of Rs is
f xm ¼ 1 ðbAT þ 1=Rs Þ=mAmT
ð135Þ
For a fixed sludge age, Rs, the design value for the minimum aerobic sludge mass fraction (1 fxm) should always be significantly higher than that given by Equation (134), because
nitrification becomes unstable and the effluent ammonia concentration increases when the aerated sludge mass fraction decreases to near the minimum value as given by Equation (134) in the same way as when the sludge age (Rs) approaches the minimum for nitrification (Rsm). This situation is exacerbated by cyclic flow and ammonia load conditions (see below). Consequently to ensure low effluent ammonia concentrations, the maximum specific growth rate of nitrifiers must be decreased by a factor of safety, Sf, to give the minimum design aerobic sludge mass fraction; from Equation (134),
ð1 f xm Þ ¼ ðbAT þ 1=Rs Þ=ðmAmT =Sf Þ
ð136aÞ
The corresponding maximum design unaerated sludge mass fraction, from Equation (136a), is
f xm ¼ 1 Sf ðbAT þ 1=Rs Þ=mAmT
ð136bÞ
With the aid of the temperature dependency equations for nitrification (Equation (131)), the maximum unaerated sludge mass fraction (fxm) from Equation (136b) is shown in Figure 25 for Sf ¼ 1.25 and mAm20 rates from 0.25 to 0.50 at 14 1C. This shows that fxm is very sensitive to mAmT. Unless a sufficiently large aerobic sludge mass fraction (1 fxm) is provided, nitrification will not take place and consequently nitrogen removal by denitrification is not possible. In fact, the selection of the maximum unaerated sludge mass fraction to achieve near complete nitrification and a required degree of N removal is the single most important decision that is made in the design of the BNR AS system because it defines the system sludge age and, for a selected reactor MLSS concentration, also the reactor volume. From Equations (132) and (136), it can be shown that at fxm for constant flow and ammonia load (i.e., steady-state conditions)
Nae ¼ KnT =ðSf 1Þ ðmgN l1 Þ
ð137Þ
From Equation (137), if Sf is selected at say 1.25 or greater at the minimum wastewater temperature, the effluent ammonia
0.80 Maximum unaerated sludge mass fraction
4.14.20.3 Unaerated Zones
467
Temperature = 14 °C
Factor of safety = 1.25
Recommended maximum 0.60 50
0.
40 36 0.
0.
0.40
30
0.
25
0.
0.20
Am20
0.00 0
10
20 30 Sludge age (days)
40
Figure 25 Maximum unaerated sludge mass fraction required to ensure nitrification vs. sludge age for maximum specific growth rates of nitrifiers mAm20 of 0.25–0.50 d1 at 14 1C for Sf ¼ 1.25.
468
Biological Nutrient Removal
concentration (Nae) will be lower than 2 mgFSA-N l1 at 14 1C for Kn20 ¼1.0 mgN l1. Although Kn is higher at higher temperature, Nae will decrease with increase in temperature because at constant sludge age, Sf increases with increase in mAmT. Consequently, for design the lower expected temperature should be selected to determine the sludge age and the aerobic mass fraction. If this is done, using say Sf ¼ 1.25, then it can be accepted from Equation (137) that the effluent ammonia concentration is below 2 mgN l1 at the lowest temperature and around 1 mgN l1 at 20 1C. In this way, explicitly calculating Nae with Equation (132) is not necessary because provision for near complete nitrification has been made by the selection of Sf. Clearly, selection of the mAm20 and Sf values has major consequences on the effluent FSA concentration and economics (size) of the ND AS system.
4.14.20.3.1 Maximum allowable unaerated mass fraction The above equations allow the two most important decisions in the design of an NDAS system to be made, the maximum unaerated sludge mass fraction and sludge age to ensure near complete nitrification. Evidently from Figure 25, for mAm20 4 0.50 the unaerated mass fraction at 14 1C can be as high as 0.7 at a sludge age of 40 days. Such a high unaerated mass fraction is apparently also acceptable at RsZ10 days at 20 1C. However, there are additional considerations that constrain the unaerated mass fraction – sludge age selection. 1. Experience with laboratory-scale ND (and NDBEPR) systems has shown that at unaerated mass fractions greater than 0.40, the filamentous bulking can become a problem, in particular at low temperatures (o16 1C). Systems with low unaerated mass fractions of o0.30 show greater tendency for good settling sludges (Musvoto et al., 1994; Ekama and Wentzel, 1999a; Tsai et al., 2003). 2. For design of BNR plants for high N and P removal, the unaerated sludge mass fraction fxm usually needs to be high (440%). If the mAm20 value is low (o 0.40 d1, which will be the usual case in designs where insufficient information on the mAm20 is available), the necessary high fxm magnitudes will be obtained only at long sludge ages (Figure 25). For example, if mAm20 ¼ 0.35 d1, then with Sf ¼ 1.3 at Tmin ¼ 14 1C, an fxm ¼ 0.45 (Equation (136b)) gives a sludge age of 25 days and for fxm ¼ 0.55 a sludge age of 37 days. Long sludge ages require large reactor volumes – increasing Rs from 25 to 37 days increases the reactor volume by 40%, whereas fxm increased only 22%. Also, for the same P content in the sludge mass, the P removal is reduced as the sludge age increases because the mass of sludge wasted daily decreases as the sludge age increases. Consequently, for low mAm20 values, the increase in N and P removal that can be obtained by increasing the unaerated sludge mass fraction above 0.50–0.60 might not be economical due to the large reactor volumes this will require, and might even be counterproductive insofar as it affects P removal. A sludge age of 30 days probably is near the limit of economic practicality which, for low mnm14 ¼ 0.16 values, will limit the unaerated mass fraction to about 0.5. At higher mnm14 values, the sludge ages allowing 50% unaerated mass fractions decrease significantly again indicating the advantages of determining
experimentally the value of mAm20 to check whether a higher value is acceptable. 3. An upper limit to the unaerated mass fraction is evident also from experimental and theoretical modeling of the BNR system. Experimentally at 20 1C with Rs ¼ 20 days, if fxm 40.70, the mass of sludge generated is found to increase sharply. Theoretically, this happens for fxm 4 0.60 at T ¼ 14 1C and Rs ¼ 20 days. The reason is that for such a high fxm, the exposure of the sludge to aerobic conditions becomes insufficient to utilize the adsorbed and enmeshed BPOs. This leads to a decrease in active mass and oxygen demand and a buildup of enmeshed nondegraded organics. When this happens, the system still functions in that the COD is removed from the wastewater, but the degradation of the COD is reduced; the system begins to behave as a contact reactor of a contact-stabilization system, that is, a bio-flocculation with minimal degradation. This critical state occurs at lower fxm as the temperature is decreased and the sludge age is reduced. From the above discussion, it would appear that the unaerated mass fraction should not be increased above an upper limit of about 60%, as indicated in Figure 25, unless there is a specific reason for this (Tsai et al., 2003).
4.14.20.4 DO Concentration High DO concentrations, up to 33 mg l1, do not appear to affect nitrification rates significantly. However, low oxygen concentrations reduce the nitrification rate. Stenstrom and Poduska (1980) have suggested formulating this effect as follows:
mAmO ¼ mAm
O ðd1 Þ KO þ O
ð138Þ
where O is oxygen concentration in liquid (mgO l1), KO the half-saturation constant (mgO l1), mAmo the maximum specific growth rate (d1), and mAO the specific growth rate at DO of O mg l1. The value of KO ranges from 0.3 to 2 mgO l1, that is, at DO values below KO the growth rate will decline to less than half the rate where oxygen is present in adequate concentrations. The wide range of KO probably has arisen because the concentration of DO in the bulk liquid is not necessarily the same as inside the biological floc where the oxygen consumption takes place. Consequently, the value will depend on the floc size, mixing intensity, and oxygen diffusion rate into the floc. Furthermore, in a full-scale reactor the DO will vary over the reactor volume due to the discrete points of oxygen input (with mechanical aeration) and the impossibility of achieving instantaneous and complete mixing. For these reasons, it is not really possible to establish a generally applicable minimum oxygen value – each reactor will have a value specific to the conditions prevailing in it. In nitrifying reactors with bubble aeration a popular DO lower limit, to ensure unimpeded nitrification, is 2 mgO l1 at the surface of the mixed liquor. Under cyclic flow and load conditions the difficulties of ensuring an oxygen supply matching the oxygen demand and a lower limit for the DO concentration are difficult.
Biological Nutrient Removal
4
2
3
2
1
0.0 0
0 0 (a)
Max. effl. FSA/steady-state FSA ratio
Amplitude of influent flow and FSA conc. 1.00 0.75 0.50 0.25
6
Steady-state FSA conc. (mgN I−1)
Max. effl. FSA/steady-state FSA ratio
4
8
1
2
3
4
5
T = 22 °C: Raw sewage
10
5
8
0.8
0.0 Amplitude of influent flow and FSA
6
0.6
1.00 4
0.4
0.75 0.50
2
0.2
0.25
0
0.0 2
6
R s /R sm ratio
1.0
Steady-state effluent FSA
(b)
Steady-state FSA conc. (mgN I−1)
T = 14 °C: Raw sewage Steady-state effluent FSA
10
469
4
6
8
10
12
14
16
R s /R sm ratio
Figure 26 (a) Maximum to steady-state effluent FSA concentration ratio vs. sludge age to minimum sludge age for nitrification ratio for influent flow and ammonia concentration amplitude (in phase) of 0.0 (steady state) 0.25, 0.50, 0.75, and 1.0 at 14 1C. (b) Maximum to steady-state effluent FSA concentration ratio vs. sludge age to minimum sludge age for nitrification ratio for influent flow and ammonia concentration amplitude (in phase) of 0.0 (steady-state) 0.25, 0.50, 0.75, and 1.0 at 22 1C.
Where storm flows are not of long duration, flow equalization is a practical way to facilitate control of the DO concentration in the reactor. In fact, most of the diurnal variations in reactor dissolved concentrations are a direct consequence of diurnal flow variation – negligibly little is due to the kinetic rates of the biological processes, especially at long sludge ages. In the absence of flow equalization, amelioration of the adverse effects of low DO concentration during peak oxygen demand periods occurs by increasing the sludge age to significantly longer than the minimum necessary for nitrification, that is, by effectively increasing Sf.
4.14.20.5 Cyclic Flow and Load It is well known both experimentally and theoretically with simulation models that under cyclic flow and load conditions the nitrification efficiency of the AS system decreases compared with that under steady-state conditions. From simulation studies, during the high flow and/or load period, even though the nitrifiers are operating at their maximum rate, it is not possible to oxidize all the ammonia available, and an increased ammonia concentration is discharged in the effluent. This in turn reduces the mass of nitrifiers formed in the system. Equivalently, the effect of diurnal variation in flow and load is to reduce the system sludge age. The average effluent ammonia concentration from a system under cyclic flow and load conditions is therefore higher than that from the same system under constant flow and load (steady-state conditions). The adverse effect of the diurnal flow variation becomes more marked as the fractional amplitude of the flow and load variation increase and is ameliorated as the safety factor Sf increases. Simulation studies of the diurnal flow effect show a relatively consistent trend between the maximum or average effluent FSA concentrations under diurnal conditions and the steady-state effluent FSA concentration versus the ratio of system sludge age and the minimum sludge age for nitrification (Rs/Rsm). For mAm20 ¼ 0.45 d1 (other constants in
Table 10), Figures 26(a) (for 14 1C) and 26(b) (for 22 1C) show the maximum (average not shown) effluent FSA concentration as a ratio of the steady-state effluent FSA concentration versus the system sludge age as a ratio of the minimum sludge age for nitrification (Rs/Rsm) for a single reactor fully aerobic system receiving cyclic influent flow and FSA load as in-phase sinusoidally varying flow and ammonia concentration, both with amplitudes of 0.25, 0.50, 0.75, 1.00, and 0.0 (steady state). For example, at 14 1C (Figure 26(a)) if the system sludge age is 2 times the minimum for nitrification, the maximum effluent FSA concentration is 8 times the steadystate value. From Figure 26(a), the latter is 0.8 mgN l1 so the maximum is 8 0.8 ¼ 6.4 mgN l1. From Figures 26(a) and 26(b), clearly the greater the diurnal flow variation and the lower the temperature, the higher the maximum (and average) effluent ammonia concentrations. This can be compensated for by increasing Sf, which has the effect of increasing the sludge age or decreasing the unaerated mass fraction of the system. This obviously has an impact on the effluent quality and/or economics of the system. The importance of the selection of mAm cannot be overemphasized. If the value of mAm is selected higher than the actual value, even with a safety factor Sf of 1.25–1.35, the plant is likely to produce a fluctuating effluent ammonia concentration, with reduced mean efficiency in nitrification. Hence, conservative estimates of mAm (low) and Sf (high) are essential for ensuring nitrification and low effluent ammonia concentration.
4.14.20.6 pH and Alkalinity The mAm rate is very sensitive to the pH of the mixed liquor outside the 7–8 range. It seems that the free ammonia (NH3) and nitrous acid (HNO2) act inhibitorily when their respective concentrations increase too high. This happens when the pH increases above 8.5 (increasing (NH3)) or decreases below 7
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Biological Nutrient Removal
(increasing (HNO2)); optimal nitrification rates are expected for 7opHo8.5 with sharp declines outside this range. From the overall stoichiometric equations for nitrification (Equation (118a)), nitrification releases hydrogen ions which in turn decreases H2CO3* Alkalinity of the mixed liquor. For every 1 mgFSA that is nitrified 2 50/14 ¼ 7.14 mg Alkalinity (as CaCO3) is consumed. Based on equilibrium chemistry of the carbonate system (Loewenthal and Marais, 1977), equations linking the pH with H2CO3* Alkalinity for any dissolved carbon dioxide concentration can be developed. These relationships are plotted in Figure 27. When the H2CO3* Alkalinity falls below about 50 mg l1 as CaCO3 then, irrespective of the carbon dioxide concentration, the pH becomes unstable and decreases to low values. Generally, if nitrification causes the H2CO3*Alkalinity to drop below about 50 mg l1 (as CaCO3), problems associated with low pH will arise at a plant, such as poor nitrification efficiency, effluents aggressive to concrete, and the possibility of development of bulking (poor settling) sludges (Jenkins et al., 1993). For any particular wastewater, the effect of nitrification on pH can be readily assessed, as follows: for example if a wastewater has a H2CO3*Alkalinity of 200 mg l1 as CaCO3 and the expected production of nitrate is 24 mgN l1, then the expected H2CO3*Alkalinity in the effluent will be (200 7.14 24) ¼ 29 mg l1 as CaCO3. From Figure 27, such an effluent will have a pH o7.0. Wastewaters having low Alkalinity (capital A denotes H2CO3* Alkalinity) are often encountered where the municipal supply is drawn from areas underlain with sandstone. A practical approach to treating such wastewaters is to (1) dose lime or better (2) create an anoxic zone(s) to denitrify some or all of the nitrate generated. In contrast to nitrification, denitrification takes up hydrogen ions which is equivalent to generating Alkalinity (see Section 4.14.24.2). By considering nitrate as electron acceptor, it can be shown that for every milligram of nitrate denitrified, there is an increase of 1 50/ 14 ¼ 3.57 mg Alkalinity as CaCO3. Hence, incorporating denitrification in a nitrification system causes the net loss of
10 0.5 1.0 2.0 5.0 10.0
Mixed liquor pH value
8 Carbon dioxide concentration (mg I−1 as CaCO3)
6
Saturation ~ 0.5 mg I−1 as CaCO3
4
Alkalinity to be reduced usually sufficiently to maintain the Alkalinity above 50 mg l1 as CaCO3 and consequently the pH above 7. In the example above, where the Alkalinity in the system is expected to decline to 29 mg l1 as CaCO3, if 50% of the nitrate were denitrified, the gain in Alkalinity would be (0.5 24 3.57) ¼ 43 mg l1 as CaCO3 and will result in an Alkalinity of (29 þ 43) ¼ 72 mg l1 as CaCO3 in the system. In this event the pH will remain above 7. For low Alkalinity wastewaters, it is imperative, therefore, that denitrification be built into nitrifying plants, even if N removal is not required. Incorporation of unaerated zones in the system influences the sludge age of the system at which nitrification takes place so that cognizance must be taken of the effect of an anoxic or unaerated zone in establishing the sludge age of a nitrifying– denitrifying plant (see Section 4.14.20.3). In the AS systems treating reasonably well buffered wastewaters, quantifying the effect of pH on nitrification is not critical because pH reduction can be limited or completely obviated by including anoxic zones, thereby ensuring Alkalinity recovery via denitrification. However, in poorly buffered wastewaters, or wastewaters with high influent N (such as AD liquors), the interaction between the biological processes, pH, and nitrification is the single most important one for the N removal AS system. Hence, it is essential to include the effect of pH on the nitrification rate for such wastewaters to quantify this important interaction. From Equation (121), the specific growth rate of the ANOs (mA) is a function of both mAm and Kn. It was shown above that the minimum sludge age is dominated by the magnitude of mAmT; it is only very weakly influenced by KnT. At RscRsm, the effluent ammonia concentration (Nae), although low, is, relatively speaking, significantly higher for larger KnT values: for example, if KnT increases by a factor of 2, the effluent ammonia concentration will increase correspondingly by the same factor (Equation (132)). Consequently, the value of KnT is significant insofar as it governs the effluent ammonia concentration once nitrification takes place at RscRsm. Several investigations have been made to understand the effect of pH on mAmT. These investigations generally have not separated out the effect of pH on mAmT and KnT so that most data are in effect lumped parameter estimates of mAmT. Almost no information is available on the effect of pH on KnT by itself. Quantitative modeling of the effect of pH on mAm has been hampered by the difficulty of accurately measuring the effects of pH on nitrification. Studies have shown that mAm can be expressed as a percentage of the highest value at optimum pH. Accepting this approach and that mAm is highest and remains approximately constant in the pH range for 7.2opHo8.0 but decreases as the pH decreases below 7.2 (Downing et al., 1964; Loveless and Painter, 1968), So¨temann et al. (2005a) modeled the mA pH dependency as For 5opHo7.2,
2
mAmpH ¼ mAm7:2 yns ðpH7:2Þ 0 −100
0
100
200 −1
Alkalinity (mg I
300
as CaCO3)
Figure 27 Mixed liquor pH vs. H2CO3* alkalinity for different concentrations of carbon dioxide.
400
ð139aÞ
where yns is the pH sensitivity coefficient (E2.35). Declining mAm values at pH48.0 have been observed and it has been noted that nitrification effectively ceases at a pH of about 9.5 (Malan and Gouws, 1966; Wild et al., 1971; Antoniou et al., 1990). Accordingly, for pH47.2, So¨temann et al.
Biological Nutrient Removal
(2005a) proposed Equation (139b) to model the decline in the mAm from pH 47.2 to 9.5 as a function of mAm7.2 using inhibition kinetics as follows:
mAmpH ¼ mAm7:2 KI
Kmax pH Kmax þ KII pH
ð139bÞ
where KI ¼ 1.13, Kmax ¼ 9.5, KIIE0.3. The overall effect of pH on mAm is modeled by combining Equations (139a) and (139b), which is given by Equation (139c) and shown in Figure 28. It can be seen that in the range pH ¼ 7.2–8.3, the change in mAmpH is small, with mAmpH/mAm7.2 40.9:
mAmpH ¼ mAm7:2 2:35ðpH7:2Þ KI
Kmax pH Kmax þ KII pH
ð139cÞ
where 2.35(pH7.2) is set ¼ 1 for pH47.2,
KI
Kmax pH ¼1 Kmax þ KII pH
for pH o7.2 and mAmpH ¼ 0 for pH49.5. Experimental data from the literature are also shown in Figure 28 to provide some quantitative support for Equation (139c). At low pH (o7.2), data from Wild et al. (1971) and Antoniou et al. (1990) fit the equation reasonably well. Very few data are available for pH48.5, but the few points from Antoniou et al. (1990) show reasonable agreement with Equation (139c). Accordingly, Equation (139c) was accepted to calculate mAmpH in the pH range 5.5–9.5. From Equation (139c), the minimum sludge age for nitrification (Rsm) at different pH and temperature (T) and unaerated mass fraction (fxm) is given by
Rsm ¼ 1=½mApHT ð1 f xm Þ bnT
ðdaysÞ
ð140Þ
The problem with nitrification in low alkalinity wastewater is that the pH obtained is not known, because it is interactively 1.2
Fraction Unm/Umm7.2
1 Eq (139b) 0.8 0.6 0.4 Eq (139a) Eq (139b) 0.2 0 4
5
6
7
8
9
10
471
established between the degree of nitrification, loss of alkalinity, pH, and mApHT. To investigate this interaction, the biological kinetic ASM1 model for carbon (C) and nitrogen (N) removal was integrated by So¨temann et al. (2005a) with a two-phase (aqueous-gas) mixed weak acid/base chemistry kinetic model to extend application of ASM1 to situations where an estimate for pH in the biological reactor is important. This integration, which included CO2 (and N2) gas generation by the biological processes and their stripping by aeration, made a number of additions to ASM1, inter alia the above effect of pH on the autotrophic nitrifiers (ANOs). From simulation of a long sludge age ND AS system with incrementally decreasing influent H2CO3* Alkalinity, when the effluent H2CO3* alkalinity fell below about 50 mg l1 as CaCO3, the aerobic reactor pH dropped below 6.3, which severely retarded nitrification and caused the minimum sludge age for nitrification (Rsm) to increase up to the operating sludge age of the system. The simulation confirmed the earlier conclusion that when treating low H2CO3* alkalinity wastewater (1) the minimum sludge age for nitrification (Rsm) varies with temperature and reactor pH and (2) for low effluent H2CO3* alkalinity (o50 mg l1 as CaCO3), nitrification becomes unstable and sensitive to dynamic loading conditions resulting in increases in effluent ammonia concentration, reduced nitrification efficiency, and as a result lower N removal. For effluent H2CO3* alkalinity o50 mg l1, lime should be dosed to the influent to raise the aerobic reactor pH and stabilize nitrification and N removal.
4.14.21 Nutrient Requirements for Sludge Production All live biological material and some unbiodegradable organic compounds contain nitrogen (N) and phosphorus (P). The organic sludge mass (VSS) that accumulates in the biological reactor comprises active organisms (XBH), endogenous residue (XEH), and UPOs (XI), each of which contains N and P. From TKN and VSS tests conducted on AS, it has been found that the N content (as N with respect to VSS, fn, mgN/mgVSS) ranges between 0.09 and 0.12 with an average of about 0.10 mgN/ mgVSS. Similarly, from total P and VSS tests, the P content (as P with respect to VSS, fp, mgP/mgVSS) of AS in purely aerobic and anoxic aerobic systems ranges between 0.01 and 0.03 with an average of about 0.025 mgP/mgVSS. From the steady-state model, the relative proportions of the active organisms (XBH), endogenous residue (XEH), and UPOs (XI) change with sludge age. Yet, it has been found that the fn value of the VSS is relatively constant at 0.10 mgN/ mgVSS. This indicates that the N content of the active organisms (XBH), endogenous residue (XEH), and UPOs (XI) is closely the same; if they were significantly different, it would be observed that fn changes in a consistent manner with sludge age. Likewise, for fully aerobic systems, the P content of the three constituents of AS is approximately similar at 0.025 mgP/mgVSS.
pH Figure 28 Maximum specific growth rate of nitrifiers, as a fraction of the rate at pH 7.2, vs. pH of the mixed liquor. (F), Model; (), Malan and Gouws (1966); ( ), Downing et al. (1964); ( ), Wild et al (1971); and (m), Antoniou et al. (1990).
4.14.21.1 Nitrogen Requirements The mass of N (or P) incorporated into the sludge mass is calculated from a N balance over the completely mixed AS
472
Biological Nutrient Removal
system (Figure 2) under steady-state daily conditions, viz., TKN flux out ¼ TKN flux in TKN flux in ¼ Qi Nti (mgN d1) TKN flux out ¼ TKN flux in Qe and Qw
Noting that Qw þ Qe ¼ Qi and Qw ¼ Vp/Rs yields
Qi Nte ¼ Qi Nti f n Xv Vp =Rs from which
Nte ¼ Nti f n MXv =ðRs Qi Þ ðmgN l1 Þ
ð141Þ
where Nte is the effluent TKN concentration (mgN l1). The term fnMXv/(RsQi) is denoted Ns and is the concentration of influent TKN in mgN l1 that is incorporated into sludge mass and removed from the system bound in the particulate sludge mass in the waste flow (Qw): 1
Ns ¼ f n MXv =ðRs Qi Þ ðmgN l
influentÞ
ðmgN l1 Þ
ð143Þ
From Equation (141), under daily average conditions, the concentration of N per liter influent required for incorporation into sludge mass is equal to the N content of the mass of sludge (VSS) wasted per day divided by the influent flow. Substituting Equation (106) relating the mass of sludge (VSS) in the reactor (MXv) to the daily average organic load on the reactor (FSti), cancelling Qi and dividing by Sti yields the concentration of N required per liter influent for sludge production per mgCOD/l organic load on the reactor, viz.,
ð1 f S0 us f S0 up ÞYH f S0 up Ns ¼ fn ð1 þ f EH bH Rs Þ þ Sti ð1 þ bH Rs Þ f cv ðmgN=mgCODÞ
Nae ¼ Nai þ Nobsi þ Nobpi ðNs Noupi Þ ðmgN l1 Þ
ð142Þ
From the N mass balance, this Ns concentration does not include the N in dissolved form in the waste flow. The soluble TKN concentration in the waste flow is the same as the effluent TKN concentration, Nte, which is soluble N in the form of ammonia (Nae) and unbiodegradable soluble organic N (Nouse). Therefore, from Equation (141), provided nitrifiers are not supported in the AS reactor so that nitrification of ammonia to nitrate does not take place, the effluent TKN concentration Nte is given by
Nte ¼ Nti Ns
organics (Nobsi and Nobpi) is released as FSA when these organics are broken down. This FSA adds to the FSA in the reactor from the influent. Some of the FSA in the reactor is taken up by the OHOs to form new OHO biomass. Some of the OHO biomass in the reactor is lost via the endogenous respiration process. The N associated with the biodegradable part of the OHO biomass is released back to the FSA pool in the reactor but the N in the unbiodegradable endogenous residue part remains as organic N bound in the endogenous residue VSS. Due to these interactions, it is possible that the effluent FSA concentration from a non-nitrifying AS system is higher than the influent FSA concentration – this occurs when the influent TKN comprises a high biodegradable organic N fraction. If the conditions are favorable for nitrification, the net FSA concentration in the reactor is available for the ANOs for growth with the associated generation of nitrate. Unless taken up for OHO growth or nitrified, the FSA remains as such and exits the system with the effluent. So in the absence of nitrification, the effluent ammonia concentration Nae is given by
ð144Þ
The influent TKN comprises ammonia and N bound in organic compounds of a soluble and particulate and biodegradable and unbiodegradable nature. The unbiodegradable organics, some of which contain organic N, are not degraded in the AS system. The influent unbiodegradable soluble organic N (Nousi) exits the system with the effluent (and waste flow) streams. The UPOs are enmeshed with the sludge mass in the reactor and so the organic N associated with these organics exits the system via the daily waste sludge (VSS) harvested from the system. The N bound in the biodegradable
ð145Þ and the effluent TKN (Nte) concentration by
Nte ¼ Nouse þ Nae
ðmgN l1 Þ
ð146Þ
The same approach is applied for the phosphorus (P) requirement for sludge production. Accepting that the P content of the AS in the fully aerobic system without BEPR is 0.025 mgP/mgVSS, the effluent total P (TP) concentration Pte is given by
Pte ¼ Pti Ps
ðmgP l1 Þ
ð147Þ
where
Ps MXv f p Ns ¼ fp ¼ Sti Rs Qi f n Sti
ðmgP l1 influentÞ
ð148Þ
4.14.21.2 N (and P) Removal by Sludge Production A plot of Equations (144) and (148) versus sludge age is given in Figure 29 for fn ¼ 0.10 mgN/mgVSS, fp ¼ 0.025 mgP/mgVSS for the example raw and settled wastewaters. It is evident that higher concentrations of TKN and TP are required for sludge production for raw than for settled wastewaters. This is because greater quantities of sludge are produced per mgCOD organic load on the reactor at the same sludge age when treating raw wastewaters (see Section 4.14.13). Also, the N and P requirements decrease as the sludge age increases because net sludge production decreases as sludge age increases. Generally, for sludge ages greater than about 10 days, the N removal from the wastewater attributable to net sludge production is less than 0.025 mgN/mgCOD load on the reactor. As influent TKN/COD ratios for domestic wastewater are in the approximate range 0.07–0.13 (Figure 29), it is clear that only a minor fraction of the influent TKN (A in Figure 29) is removed by incorporation into sludge mass. Additional N removal (B in Figure 29) is obtained by transferring the N from the dissolved form in the liquid phase to the gas phase
Biological Nutrient Removal Nutrient requirements 0.035 Approximate range of influent TKN/COD and P/COD ratios of municipal wastewaters
0.12 0.10
0.030 0.025
0.08
0.020
0.06
0.015 B
0.04
0.010 Raw
0.02
0.005 A
P requirement (mgP/mgCOD)
N requirement (mgN/mgCOD)
0.14
Settled
0.00
0.000 0
5
10 15 20 Sludge age (days)
25
30
Figure 29 Approximate minimum nutrient N and P requirements as mgN l1 influent TKN and mgP l1 influent total P per mgCOD l1 organic load on the activated sludge reactor vs. sludge age for the example raw and settled wastewaters at 20 1C together with influent TKN and TP to COD concentration ratio ranges for municipal wastewater.
by autotrophic nitrification and heterotrophic denitrification, which transforms the nitrate to nitrogen gas in anoxic (nonaerated) reactor(s). The details of heterotrophic denitrification are presented below. From Figure 29, normal P removal by incorporation into biological sludge mass is limited at about 0.006 and 0.004 mgP/mgCOD for raw and settled wastewaters respectively, effecting a TP removal of about 20–25% from average municipal wastewaters. As transformation of dissolved orthoP to a gaseous form is not possible, to increase the P removal from the liquid phase, additional ortho-P needs to be incorporated into the sludge mass. This can be achieved in two ways: (1) chemically and/or (2) biologically. With chemical P removal, iron or aluminum chlorides or sulfates are dosed to the influent (pre-precipitation), to the AS reactor (simultaneous precipitation) or to the final effluent (post-precipitation). The disadvantage of chemical P removal is that it significantly increases (1) the salinity of treated wastewater, (2) the sludge production due to the inorganic solids formed, and (3) the complexity and cost of the WWTP. With biological P removal, the environmental conditions in the biological reactor are designed in such a way that a specific group of heterotrophic organisms (called PAOs) grow in the AS reactor. With the accumulated polyPs, these organisms have a much higher P content than the OHOs, as high as 0.38 mgP/ mgPAOVSS (Wentzel et al., 1990). The more PAOs that grow in the reactor, the higher will be the mean P content of the VSS sludge mass in the reactor and therefore the higher the P removal via the wasted sludge. With a significant mass of PAOs present, the mean P content of the VSS sludge mass can increase from 0.025 mgP/mgVSS in aerobic systems to 0.10– 0.15 mgP/mgVSS in biological N and P removal systems. The advantage of biological P removal over chemical P removal is that (1) the salinity of the treated wastewater is not increased, (2) sludge production is increased only between 10% and 15%, and (3) the system is less complex and costly to operate.
473
A disadvantage of biological P removal is that, being biological, it is less dependable and more variable than chemical P removal. The biological processes which mediate biological N and P removal in AS systems and the different reactor configurations in which these take place are described in Section 4.14.28.
4.14.22 Nitrification Design Considerations The kinetic equations describing the interactions between the FSA and the organic N are complex and have been developed in terms of the growth–death–regeneration approach in AS simulation models such as ASM1 and ASM2. However, for steady-state conditions assuming (1) all the biodegradable organics are utilized in the reactor and (2) a TKN mass balance over the AS system, a simple steady-state nitrification model can be developed from the nitrification kinetics and the N requirements for sludge production considered above. This model is adequate for steady-state design and from it some general response graphs are developed below for the example raw and settled wastewaters. Dynamic system responses can be determined with the simulation models once (1) the AS system has been designed and sludge age, zone and reactor volumes and recycle flows are known and (2) the steady-state concentrations have been calculated to serve as initial conditions for the simulation. In the nitrifying AS system design, the (1) effluent FSA, TKN, and nitrate concentrations and (2) the nitrification oxygen demand need to be calculated.
4.14.22.1 Effluent TKN The filtered effluent TKN (Nte) comprises the FSA (Nae) and the unbiodegradable soluble organic N (Nouse). Once mAm20, fxt, Rs, and Sf have been selected, the equations for these concentrations are: 1. Effluent FSA (Nae). Nae is given by Equation (132), which applies only if Rs4Rsm, which will be the case for Sf41.0. 2. Effluent soluble biodegradable organic nitrogen concentration (Nobse). The biodegradable organics (both soluble and particulate) are broken down by the OHOs releasing the organically bond N as FSA. In the steady-state model, it is assumed that all the biodegradable organics are utilized. Hence, the effluent soluble biodegradable organic N concentration (Nobse) is zero. 3. Effluent soluble unbiodegradable organic nitrogen concentration (Nouse). Being unbiodegradable, this concentration of organic N flows though the AS system with the result that the effluent concentration (Nouse) is equal to the influent concentration (Nousi), that is,
Nouse ¼ Nousi
ð149Þ
where Nousi is the influent soluble unbiodegradable organic nitrogen, mgOrgN-N l1 ¼ fN0 ous Nti, where fN0 ous is the soluble unbiodegradable organic N fraction of the influent TKN (Nti). The two nonzero effluent TKN concentrations (FSA, Nae and OrgN, Nouse) are soluble and so exit with the effluent (and
474
Biological Nutrient Removal
waste flow). The soluble (filtered) TKN in the effluent (Nte) is given by their sum, that is,
Nte ¼ Nae þ Nousi
ðfiltered TKNÞ
ð150Þ
If the effluent sample is not filtered, the effluent TKN will be higher by the concentration of TKN in the effluent VSS, that is,
Nte ¼ Nae þ Nouse þ f n Xve
ðunfiltered TKNÞ
nitrogen required for sludge production per mgCOD applied (from Equation (144)). The nitrification capacity to influent COD concentration ratio (Nc/Sti) of a system can be estimated approximately by evaluating each of the terms in Equation (153) as follows:
•
ð151Þ
where Xve is the effluent VSS concentration (mgVSS l1) and fn the N content of VSS (B0.1 mgOrgN-N/mgVSS).
•
4.14.22.2 Nitrification Capacity From a TKN mass balance over the AS system and Rs 4 Rsm, the concentration of nitrate generated in the system (Nne) with respect to the influent flow is given by the influent TKN (Nti) minus the soluble effluent TKN (Nte) and the concentration of influent TKN incorporated in the sludge wasted daily from the AS system (Ns), that is,
Nne ¼ Nc ¼ Nti Nte Ns
ð152Þ
The Ns concentration is determined from the mass of N incorporated in the VSS mass harvested from the reactor per day (Equation (142)). The mass of VSS in the reactor (MXv) does not have to include the VSS mass of nitrifiers because this mass, as mentioned earlier, is negligible (o2–4%). In Equation (152), Nc defines the ‘nitrification capacity’ of the AS system. The nitrification capacity (Nc) is the mass of nitrate produced by nitrification per unit average influent flow, that is, mgNO3-N l1. In Equation (150), the effluent TKN concentration (Nte) depends on the efficiency of nitrification. In the calculation for the maximum unaerated sludge mass fraction (fxm) at a selected sludge age, if the factor of safety (Sf) was selected 41.25 to 1.35 at the lowest expected temperature (Tmin), the efficiency of nitrification be high (495%) and Nae generally will be less than 1–2 mgN l1. Also, with Sf 41.25 at Tmin, Nae will be virtually independent of both the system configuration and the subdivision of the sludge mass into aerated and unaerated mass fractions. Consequently, for design, with Sf41.25, Nte will be around 3–4 mgN l1 provided that there is reasonable assurance that the actual mAm20 value will not be less than the value accepted for design and that there is sufficient aeration capacity so that nitrification is not inhibited by an insufficient oxygen supply. Accepting the calculated fxm and selected sludge age (Rs) at the lower temperature, then at higher temperatures the nitrification efficiency and the factor of safety (Sf) both will increase so that at summer temperatures (Tmax), Nte will be lower, approximately 2–3 mgN l1. Dividing Equation (152) by the total influent COD concentration (Sti) yields the nitrification capacity per mgCOD applied to the biological reactor, Nc/Sti, viz.,
Nc =Sti ¼ Nti =Sti Nte =Sti Ns =Sti
ð153Þ
where Nc/Sti is the nitrification capacity per mgCOD applied to the AS system (mgN/mgCOD), Nti/Sti the influent TKN/ COD concentration ratio of the wastewater, and Ns/Sti the
•
Nti/Sti: This ratio is a wastewater characteristic and obtained from the measured influent TKN and COD concentrations – it can range from 0.07 to 0.10 for raw municipal wastewater and 0.10 to 0.14 for settled wastewater. Nte/Sti: Provided the constraint for efficient nitrification is satisfied at the lowest temperature (Tmin), the effluent TKN at Tmin (Nte) will be low at B2–3 mgN l1, that is, for influent COD concentrations (Sti) ranging from 1000 to 500, Nte/Sti will range from 0.005 to 0.010. At Tmax, NteE1– 2 mgN l1 making the Nte/Sti ratio lower. Ns/Sti: Given by Equation (144).
A graphical representation of the relative importance of these three ratios to the nitrification capacity, Nc/Sti, is shown in Figure 30(a) (for 14 1C) and 30(b) (for 22 1C) and were generated by plotting Nc/Sti versus sludge age for selected influent TKN/COD (Nti/Sti) ratios of 0.07, 0.08, and 0.09 for the example raw wastewater and settled wastewater for 40% COD and 15% TKN removal in primary settling, viz., 0.113, 0.127, and 0.141. Also shown are the minimum sludge ages for nitrification at unaerated sludge mass fractions of 0.0, 0.2, 0.4, and 0.6 for the example mAm20 value of 0.45 d1. For a particular unaerated sludge mass fraction, the plotted values of Nc/Sti are valid only at sludge ages longer than the corresponding minimum sludge age. These figures show the relative magnitudes of the three terms that affect the nitrification capacity versus sludge age and temperature. 1. Temperature. To obtain complete nitrification at 14 1C (for a selected fxm), the sludge age required is more than double that at 22 1C. The corresponding nitrification capacities per influent COD at 14 1C show a marginal reduction to those at 22 1C, because sludge production at 14 1C is slightly higher than at 22 1C due to the reduction in endogenous respiration rate of the OHOs. 2. Sludge age. For a selected influent TKN/COD ratio (Nti/Sti), the nitrification capacity (Nc/Sti) increases as the sludge age increases because the N required for sludge production decreases with sludge age, making more FSA available for nitrification. However, the increase is marginal for Rs410 days. 3. Influent TKN/COD ratio (Nti/Sti). Clearly, for both raw and settled wastewater, at any selected sludge age, the nitrification capacity (Nc/Sti) is very sensitive to the influent TKN/COD ratio (Nti/Sti). An increase of 0.01 in Nti/Sti causes equal increase of 0.01 in Nc/Sti. For the same Nti/Sti ratio for raw or settled wastewater, the nitrification capacity (Nc/Sti) for raw wastewater is lower than for settled wastewater because more sludge (VSS) is produced per unit COD load from raw wastewater than from settled wastewater because the unbiodegradable particulate COD fraction (fS’up) in raw water is higher than in settled wastewater. Apart from this difference, an increase in influent TKN/COD ratio will result in an equal increase in nitrate concentration (nitrification capacity) per influent
Biological Nutrient Removal
WW Char fS’us Raw 0.07 Settled 0.12
0.10
0.05 Raw wastewater
0.15
0.0 0.2 0.4 0.6 −1 Unaerated mass fraction 14 °C; UA20 = 45 d bA20 = 0.04 d−1; Sf = 1.25
0.00 0 (a)
fS’up Settled wastewater 0.15 0.141 0.04 0.127 0.113 TKN/COD ratio 0.10 0.09 0.08
Nitrification capacity
Nitrification capacity
0.15
5
20 10 15 Sludge age (days)
25
fs’us 0.07 0.12
fs’up 0.15 0.04
Settled wastewater 0.141 0.127
0.10
0.113 TKN/COD ratio 0.10 0.09 0.08
0.05 Raw wastewater
0.00
30
WW Char Raw Settled
475
0.2 0.6 22 °C; UA20 = 0.45 d−1 0.0 0.4 −1 b Unaerated mass fraction A20 = 0.04 d ; Sf = 1.25 0
5
(b)
10 20 15 Sludge age (days)
25
30
Figure 30 Nitrification capacity per mgCOD applied to the biological reactor vs. sludge age for different influent TKN/COD concentration ratios in the example raw and settled wastewaters at 14 1C (a) and 22 1C (b). Also shown as vertical lines are the minimum sludge ages required to achieve nitrification for Sf ¼ 1.25 for unaerated sludge mass fractions of 0.0, 0.2, 0.4, and 0.6.
COD. This decreases the likelihood, or makes it impossible, to obtain complete denitrification using the wastewater organics as electron donor. This will become clear when denitrification is considered below. Because primary settling increases the influent TKN/COD ratio, N removal via nitrification denitrification is always lower with settled wastewater than with raw wastewater. However, this lower N removal comes with the advantage of a smaller biological reactor and lower oxygen demand resulting significant savings in reactor and oxygenation costs.
4.14.22.3 Mass of Nitrifiers (MXA) and Nitrification Oxygen Demand (FOn) Once nitrification takes place because the sludge age of the system is longer than the minimum required for nitrification, the mass of nitrifiers (MXA, mgVSS) in the reactor is calculated from the flux of nitrate generated (FNne) in the same way as the mass of OHOs (MXBH) was calculated from the flux of biodegradable organics (Equation (86)), viz.,
MXBA ¼ FNne YA Rs =ð1 þ bAT Rs Þ ðmgVSSÞ
ð154Þ
where FNne is the flux of nitrate generated ¼ (Qe þ Qw)Nne ¼ Qi Nne (mgN d1) and Nne is given by Equation (152). The oxygen demand for nitrification is simply 4.57 mgO/ mgN times the flux of nitrate produced, that is,
FOn ¼ 4:57 FNne ¼ OURn Vp
ðmgO d1 Þ
ð155Þ
Table 11 Raw and settled wastewater characteristics required for calculating effluent N concentrations from nitrification AS systems Influent WW characteristic
Sym
Raw
Influent TKN (mgN l1) Influent TKN/COD ratio Influent FSA fraction Unbio sol orgN fraction Unbio partic VSS N content
Nti fns f N0 a fN0 ous fn
60 0.08 0.75 0.03 0.1
Influent pH Influent Alk mg l1 as CaCO3 ANO max spec growth rate Influent flow rate (M l d1)
Alk mAm20 Qi
7.5 200 0.45 15
Seta 51 0.113 0.88 0.034 0.1 7.5 200 0.45 15
a
Settled wastewater (WW) characteristics must be selected/calculated to be consistent with the raw wastewater ones and mass balances over the primary settling tanks, e.g., soluble concentrations must be the same in settled wastewater as in raw wastewater.
organics (COD) removal (see Section 4.14.9.5). The wastewater characteristics for the raw and settled wastewaters for COD removal are listed in Table 7 and the additional characteristics required for nitrification are listed in Table 11. The nitrifier kinetic constants in Table 10, adjusted for wastewater temperatures 14 and 22 1C, were applied. No adjustment to mAm20 for pH was made, that is, an effluent Alkalinity 450 mg l1 as CaCO3 was assumed. Also, it is accepted that all the biodegradable organics are degraded and their N content released as ammonia so the effluent soluble biodegradable organic N concentration (Nobse) is zero.
4.14.23.2 Nitrification Process Behavior
4.14.23 Nitrification Design Example 4.14.23.1 Wastewater Characteristics Design of a nitrification AS system without denitrification is considered below. For the purpose of comparison, the nitrifying AS system is designed for the same wastewater flow and characteristics accepted for the design of the AS system for
From Equation (20a), the unbiodegradable soluble organic nitrogen in the effluent is Nouse ¼ Nousi ¼ 1.8 mgN l1 for raw and settled wastewater The ammonia concentration available for nitrification (Nan) is the influent TKN concentration (Nti) minus the N concentration required for sludge production (Ns) (Equation (142)) and the USO N concentration in the effluent
476
Biological Nutrient Removal
(Nouse), viz.,
Nan ¼ Nti Ns Nouse
ðmgN l1 Þ
ð156Þ
If the sludge age of the system is shorter than the minimum required for nitrification (RsoRsm), no nitrification takes place and the effluent nitrate concentration (Nne) is zero. The effluent ammonia concentration (Nae) is equal to the nitrogen available for nitrification (Nan, Equation (156)). If Rs4Rsm for Sf ¼ 1.0, most of the FSA available for nitrification is nitrified to nitrate and the effluent nitrate concentration (Nne) is the difference between Nan (Equation (156)) and the effluent FSA concentration given by Equation (132). For both RsoRsm and Rs4Rsm, the effluent TKN concentration (Nte) is the sum of effluent ammonia and unbiodegradable soluble organic nitrogen concentrations (Nte ¼ Nae þ Nouse). For RsoRsm, no nitrification takes place so the effluent nitrate concentration (Nne) is zero and the effluent ammonia concentration (Nae) is given by Nan (Equation (156)). The nitrifier sludge mass (MXA) and the nitrification oxygen demand (FOn) are both zero because Nne is zero. With increasing sludge age starting from Rs ¼ 0, Nae from Equation (132) is first negative (which is of course impossible) and then 4Nan (which is also not possible). For a sludge age slightly longer than Rsm, the Nae falls below Nan. From this sludge age, nitrification takes place and for further (even small) increases in sludge age, the Nae rapidly decreases to low values (o4 mgN l1). Hence for Rs4Rsm, the effluent ammonia concentration (Nae) is given by Equation (132), the effluent TKN concentration by Nte ¼ Nae þ Nouse, and the effluent nitrate concentration (Nne) by
Nne ¼ Nan Nae ¼ Nti Ns Nte
ðmgN l1 Þ
ð157Þ
With nitrification, the nitrifier biomass and nitrification oxygen demand are given by Equations (154) and (155). Substituting the influent N concentrations for raw and settled wastewaters and the values of the kinetic constants at 14 1C into the above equations, the results at different sludge ages were calculated. In Figure 31(a), the different effluent N concentrations from the system versus sludge age for raw and settled wastewater at 14 1C are shown. In Figure 31(c) are shown the nitrifier sludge mass (as a % of the reactor VSS mass) and nitrification oxygen demand for raw and settled wastewater at 14 1C. Also shown in Figure 31(c) are the carbonaceous and total oxygen demands for raw and settled wastewater at 14 1C. The calculations were repeated for 22 1C and shown in Figures 31(b) and 31(d). Figures 31(a) and 31(b) show that once the sludge age is approximately 25% longer than the minimum required for nitrification, nitrification is virtually complete (for steady-state conditions) and comparing the results for raw and settled wastewater, there is little difference between the nitrification oxygen demand and the concentrations of ammonia, nitrate, and TKN in the effluent. The reasons for this similar behavior are: (1) the primary settling tank removes only a small fraction of the influent TKN and (2) settled wastewater results in lower sludge production, so that the FSA available for nitrification in raw and settled wastewater is nearly the same. Once
nitrification takes place, temperature has relatively little effect on the different effluent N concentrations. However, a change in temperature causes a significant change in the minimum sludge age for nitrification. Considering Figures 31(a) and 31(b), for RsoRsm, the effluent ammonia concentration (Nae) and hence the effluent TKN concentration (Nte) increase with increasing sludge age up to Rsm because Ns decreases for increases in Rs. For Rs4Rsm, Nae decreases rapidly to o2 mgN l1 so that for Rs41.3Rsm, the effluent TKN concentration is o4 mgN l1. The increase in nitrate concentration (Nne) with an increase in sludge age for Rs41.3Rsm is mainly due to the reduction in N required for sludge production (Ns). This is important for BNR systems – increasing the sludge age increases the nitrification capacity (see Figure 30) so more nitrate has to be denitrified to achieve the same N removal. Figures 31(c) and 31(d) show that the nitrification oxygen demand increases rapidly once Rs4Rsm but for Rs41.3Rsm, further increases are marginal irrespective of the temperature or wastewater type. This nitrification oxygen demand represents an increase of 42% and 65% above the COD for the raw and settled wastewater. However, the total oxygen demand for treating settled wastewater is only 75% of that for treating raw wastewater. In order that nitrification can proceed without inhibition by oxygen limitation, it is important that the aeration equipment is adequately designed to supply the total oxygen demand; generally, heterotrophic organism growth takes precedence over nitrifier growth when oxygen supply (or ammonia) becomes insufficient. This is because heterotrophic organisms can grow adequately with DO concentrations of 0.5–1.0 mgO l1, whereas nitrifiers tend to require higher DO concentrations. Just as the effluent FSA concentration rapidly decreases for Rs4Rsm, the nitrifier sludge mass rapidly increases once Rs4Rsm, is slightly higher at 14 1C than at 22 1C due to the lower endogenous respiration rate. Also, because the concentrations of nitrate produced with raw and settled wastewater are closely similar (B40 mgN l1), the nitrifier sludge mass is approximately the same at the same sludge age (B430 kgVSS at 10d sludge age and B900 kgVSS at 30 day sludge age). Because with raw wastewater so much more sludge mass is produced than with settled wastewater, the nitrifier sludge mass is a much smaller proportion of the VSS mass with raw waste water (B1.4% at 10 day sludge age) than with settled wastewater (B3.3% at 10 day sludge age). Comparing the nitrifier sludge mass to the heterotrophic sludge mass, as in Figures 31(c) and 31(d), the nitrifier sludge mass comprises o4% of VSS mass even at high TKN/COD ratios for settled wastewater and so is ignored in the determination of the VSS concentration in the AS reactor treating domestic wastewater. It is worth repeating that primary sedimentation removes only a minor fraction of the TKN but a significant fraction of COD (15% and 40% in this example). Even though the settled wastewater has a lower TKN concentration than the raw wastewater, the effluent nitrate concentration does not reflect this difference. This is because the N removal for sludge production is lower for settled than for raw wastewater. Consequently, the nitrate concentration for settled wastewater is nearly the same as for raw wastewater – for different
(a)
Nitrogen-FSA, TKN, NO3 0.0
0.2
0.4
0.6
0.8
1.0
1.2
0
14 °C
5
0
FOc
5
Rsm
Ns
%Nit
25
25
30
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
30
Raw WW Settled WW
Raw WW Settled WW
FOn FOn
FOc
FOc
FOt
FOt
15 20 10 Sludge age (days)
%Nit
20 10 15 Sludge age (days)
Nte
10
Nouse = 1.8 mg N l−1 Nte = Nae + Nouse
Nne
Influent TKN
Ns
0
Nte
Rsm
20
30
40
50
60
14 °C
% Nitrifier VSS
(b)
(d)
0.0
0.2
0.4
0.6
0.8
1.0
1.2
0 0
0
5
22 °C Rsm
5
FOn FOn
%Nit
FOt
FOt
25
25
30
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
30
Raw WW Settled WW
22 °C
Raw WW Settled WW
FOc
FOc
10 20 15 Sludge age (days)
%Nit
15 20 10 Sludge age (days)
Ns
10
Nouse = 1.8 mg N l−1 Nte = Nae + Nouse
Nne
Influent TKN
Ns Nte
Rsm
20
30
40
50
60
70
Figure 31 Effluent ammonia (Nae), TKN (Nte), and nitrate (Nne) concentrations and N required for sludge production (Ns) vs. sludge age at 14 1C (a) and 22 1C (b) and nitrification (FOn), carbonaceous (FOc) and total (FOt) oxygen demand in kgO/kgCOD load and % nitrifier VSS mass vs. sludge age at 14 1C (c) and 22 1C (d) for the example raw and settled wastewater.
(c)
Oxygen demand (kgO/kgCOD)
Nitrogen-FSA, TKN, NO3 Oxygen demand (kgO/kgCOD)
70
% Nitrifier VSS
478
Biological Nutrient Removal
wastewater characteristics, it can be higher than raw wastewater. In contrast, the maximum N removal by denitrification using the wastewater organics as electron donor, called the denitrification potential, mainly depends on the influent COD concentration and this concentration is significantly reduced by primary sedimentation. This may result in a situation where it may be possible to obtain near-complete nitrate removal when treating raw wastewater but not when treating settled wastewater. The difference in COD and TKN removal in PSTs therefore has a significant effect on the design of BNR systems.
4.14.24 Biological Denitrification
ages can be in the usual fully aerobic short sludge age range of 3–6 days. Unaerated zones should still be incorporated to derive the benefits of denitrification in the event nitrification does take place. When it does not, the unaerated zone will be anaerobic (no input of DO or nitrate) instead of anoxic, and some BEPR may take place. Because BEPR is not required and therefore not exploited to the full, whether or not it takes place is not important because it does not affect the system behavior very much. With some BEPR, the sludge production will be slightly higher (o5%) per COD load, the VSS/TSS ratio and oxygen demand both somewhat lower (by about 5%). However, BEPR may result in mineral precipitation problems in the sludge treatment facilities if the WAS is anaerobically digested.
4.14.24.1 Interaction between Nitrification and N Removal Nitrification is a prerequisite for denitrification – without it biological N removal is not possible. Once nitrification takes place, N removal by denitrification becomes possible and should be included even when N removal is not required (see Section 4.14.14) by incorporating zones in the reactor that are intentionally unaerated. Because the nitrifiers are obligate aerobes, nitrification does not take place in the unaerated zone(s), so to compensate for this, the system sludge age needs to be increased for situations where nitrification is required. For fully aerobic systems and a wastewater temperature 14 1C, a sludge age of 5–7 days may be sufficient for complete nitrification, taking due consideration of the requirement that the effluent FSA concentration should be low even under cyclic flow and load conditions (Sf41.3). For anoxic – aerobic systems, a sludge age of 15–20 days may be required when a 50% unaerated mass fraction is added (Figure 25). Therefore, for plants where N removal is required, invariably the sludge ages are long due to (1) the uncertainty in the mAm20 value, (2) the need for unaerated zones, and (3) the guarantee of nitrification at the minimum average winter temperature (Tmin). For plants where nitrification is a possibility and not obligatory, uncertainty in the mAm20 value is not important and unaerated zones can be smaller, with the result that sludge Table 12
4.14.24.2 Benefits of Denitrification In the design of fully aerobic systems discussed above, it was suggested that when nitrification is not obligatory but a possibility, unaerated zones should still be incorporated in the system to derive the benefits of denitrification. These benefits include (1) reduction in nitrate concentration which ameliorates the problem of rising sludge from denitrification in the secondary settling tank (Section 4.14.14), (2) recovery of alkalinity (Section 4.14.20.6), and (3) reduction in oxygen demand. With regard to (3), under anoxic conditions, nitrate serves as electron acceptor instead of DO in the degradation of organics (COD) by facultative heterotrophic organisms. The oxygen equivalent of nitrate is 2.86 mgO/mgNO3-N which means that 1 mg NO3-N denitrified to N2 gas has the same electron-accepting capacity as 2.86 mg of oxygen. In nitrification to nitrate, the FSA donates eight electrons (e)/mol, the N changing from an e state of 3 to þ 5. In denitrification to N2, the nitrate accepts 5 e/mol, the N changing from an e state of þ 5 to 0. Because 4.57 mgO/mgFSA-N are required for nitrification, the oxygen equivalent of nitrate in denitrification to N2 is 5/8 4.57 ¼ 2.86 mgO/mgNO3-N (Table 12). Therefore, for every 1 mg NO3-N denitrified to N2 gas in the anoxic zone, during which about 2.86/(1 fcvYH) ¼ 8.6 mgCOD
Comparison of nitrification and denitrification processes in single sludge activated sludge systems Nitrification
Denitrification
Form Function Half-reaction Organisms Environment
Ammonia (NHþ 4) Electron donor Oxidation Autotrophs Aerobic
Nitrate (NO 3) Electron acceptor Reduction Heterotrophs Anoxic
Compound Oxid. no.
NH4 þ 3 Nitrification (oxidation)
N2 0
NO2 þ3
8 e/atom N ¼ 4.57 mgO/mgN Net loss Denitrification (reduction) 5e/atom N ¼ 2.86 mgO/mgN Nitrification: 4.57 mgO/mgNH4-N nitrified to NO3-N Denitrification: 2.86 mgO recovered/mg NO3-N denitrified to N2 gas Therefore, denitrification allows at best 62.5% (5/8 or 2.86/4.57) recovery of the nitrification oxygen demand.
NO3 þ5
Biological Nutrient Removal Effluent TKN liquid phase
Raw wastewater TKN/COD = 0.08 1.0
Oxygen demand (kgO/kgCOD on reactor)
14 °C 22 °C
479
Total incl. nitrification
~5%
N in gas phase ~75%
0.8
N in sludge solid phase ~20%
0.6 Total incl. nitrif. and denit. 0.4
N nitrified (transformed in liquid phase) and possibly denitrified (transferred to gas phase)
Carbonaceous
Possibility of nitrification 0.2
Possibility of denitrification
Figure 33 Exit routes for nitrogen in single sludge nitrification denitrification activated sludge systems.
0.0 0
5
10
15
20
25
30
Sludge age (days) Figure 32 Carbonaceous, total including nitrification and total including nitrification and denitrification oxygen demand per unit COD load on the biological reactor vs. sludge age for the example* raw wastewater. *Note: All the figures in this part which show the behavior of the various activated sludge system configurations were generated from the example raw and settled wastewater characteristics.
is utilized, 2.86 mg less oxygen needs to be supplied to the aerobic zone. Because the oxygen requirement to form the nitrate from ammonia is 4.57 mgO/mgNO3-N, and 2.86 mgO/mgNO3-N is recovered in denitrification to N2 gas, a maximum of 2.86/4.57 (or 5/8ths) ¼ 0.63 of the nitrification oxygen demand can be recovered. A comparison of the nitrification and denitrification reactions is given in Table 12. Under operating conditions, it is not always possible to denitrify all the nitrate formed with the result that the nitrification oxygen recovery by denitrification is about 50% (see Figure 32). Therefore, once the possibility of nitrification exists, it is always worthwhile to consider including intentional denitrification because of the recovery of alkalinity and oxygen. With regard to oxygen, if the oxygen supply is insufficient to meet the combined carbonaceous and nitrification requirement, areas in the aerobic reactor will become anoxic. Under oxygen limited conditions, the aerobic mass fraction in the aerobic reactor will vary depending on the COD and TKN load on the plant over the day. At minimum load, oxygen supply may be adequate so that nitrification may be complete whereas, at peak load, oxygen supply may be insufficient so that nitrification may cease (partially or completely) and denitrification will take place on the accumulated nitrate. This behavior is exploited in the single-reactor ND configurations such as the ditch or Carousel-type systems.
4.14.24.3 N Removal by Denitrification In biological N removal systems, the N is removed by transfer from the liquid phase to the solid and gas phases. About 20% of the influent N is incorporated in the sludge mass (Figure 33) but the bulk of the N (i.e., about 75% when complete denitrification is possible) is removed by transfer to the gas phase via nitrification and denitrification (Figure 33). In the nitrification step, the N remains in the liquid phase because it is transformed from ammonia to nitrate. In the denitrification step, it is transferred from the liquid to the gas phase and escapes to the atmosphere. When complete denitrification is achieved, a relatively small fraction of the influent TKN (B5%) remains in the liquid phase and escapes as total N (TKN þ nitrate) with the effluent. For aerobic conditions, the problem of the designers is to calculate the mass of oxygen electron acceptor required by the OHOs (and ANOs) for the utilization of the known mass of organic electron donors (organics and ammonia) available. For anoxic conditions, the problem is the opposite. Here, the problem is to calculate the mass of electron donors (COD) that are required to denitrify a known mass of electron acceptors nitrate. If sufficient electron donors (COD) are not available then complete denitrification cannot be achieved. The calculation of the nitrogen removal is essentially a reconciliation of electron acceptors (nitrate) and donors (COD) taking due account of (1) the biological kinetics of denitrification and (2) the system operating parameters (such as recycle ratios and anoxic reactor sizes) under which the denitrification is constrained to take place. The electron donors (or COD or energy) for denitrification can come from two sources: (1) internal or (2) external to the AS system. The former are those present in the system itself, that is, those in the incoming wastewater or generated within the biological reactor by the AS itself; the latter are organics imported to the AS system and specifically dosed into the anoxic zone(s) to promote denitrification, (e.g., methanol,
480
Biological Nutrient Removal
acetate, and molasses; Monteith et al., 1980). Here, the focus is on internal COD sources for denitrification, but the principles and procedures are sufficiently general to be adaptable to include external COD (energy) sources also.
4.14.24.4 Denitrification Kinetics There are three internal organics sources, two from the wastewater and one from the AS sludge mass itself. The two in the wastewater are the two main forms of organics (i.e., RBSO) and slowly biodegradable organics (BPO)). The third is slowly biodegradable organics generated by the biomass itself through death and lysis of organism mass (also known as endogenous mass loss/ respiration). This self-generated BPO is utilized in the same way as the wastewater BPO, but is recognized separately because of its different source and rate of supply to that of the influent. The RBSO and BPO (influent or self-generated) are degraded via different mechanisms by the OHOs. The different RBSO and BPO degradation mechanisms lead to different COD utilization rates. The RBSO comprises small simple dissolved organic compounds that can pass directly through the cell wall into the organism, for example, sugars and short-chain fatty acids. Accordingly, the RBSO can be used at a high rate which does not change significantly whether nitrate or oxygen serves as terminal electron acceptor (Ekama et al., 1996a). Simulation models use the Monod equation to model the utilization of RBSO by OHOs under both aerobic and anoxic conditions. The BPO comprises large particulate or colloidal organic compounds, too large to pass into the organism directly. These organics must be broken down (hydrolyzed) in the slime layer surrounding the organism to smaller components, which then can be transferred into the organism and utilized. The extracellular BPO hydrolysis rate is slow and forms the limiting rate in the utilization of BPO (Section 4.14.5.1.3). This hydrolysis rate is much slower under anoxic conditions than under aerobic conditions – only about one-third (Stern and Marais, 1974, van Haandel et al., 1981). This introduces a reduction factor Z in the BPO hydrolysis rate equation for anoxic conditions (Equation (159) below). Research has indicated that the utilization of RBSO is simultaneous with the hydrolysis of BPO. Also the rate of RBSO utilization is considerably faster (7–10 times) than the rate of BPO hydrolysis so the denitrification rate with influent RBSO is much faster than with BPO. Therefore, the influent RBSO is the preferred organic for denitrification and the higher this concentration in the influent with respect to the total COD, the greater the N removal.
4.14.24.5 Denitrification Systems As a result of the different degradation mechanisms and rates of RBSO and BPO utilization, the position of the anoxic zone in the biological reactor significantly affects the denitrification that can be achieved. There are many different configurations of single sludge ND systems but from the point of view of the source of the organics (electron donors), these can be simplified to two basic types of denitrification or combinations of these. The two basic types utilizing internal organics are: (1) post-denitrification, which utilizes self-generated
endogenous organics and (2) pre-denitrification, which utilizes influent wastewater organics. With post-denitrification (Figure 34(a)), the first reactor is aerobic and the second is unaerated. The influent is discharged to the aerobic reactor where aerobic growth of both the heterotrophic and nitrifying organisms takes place. Provided the sludge age is sufficiently long and the aerobic fraction of the system is adequately large, nitrification will be complete in the first reactor. The mixed liquor from the aerobic reactor passes to the anoxic reactor, also called the secondary anoxic reactor, where it is mixed with stirring. The outflow from the anoxic reactor passes through an SST and the underflow is recycled back to the aerobic reactor. The BPO released by the sludge mass via the death of organisms provides the energy source for denitrification in the anoxic reactor. However, the rate of release of energy is low, so that the rate of denitrification is also low. To obtain a meaningful reduction of nitrate, the anoxic mass fraction of the
Anoxic reactor
Aerobic reactor
Waste flow Settler
Influent
Effluent
Sludge recycle
(a)
s
Anoxic Aerobic reactor reactor Mixed liquor Recycle
Waste flow
a
Settler
Influent
Effluent
I
Sludge recycle
(b)
Primary Aerobic anoxic reactor reactor Mixed liquor Recycle a
s
Secondary anoxic reactor Reaeration reactor Waste flow Settler Effluent
Influent
(c)
Sludge recycle
s
Figure 34 (a) The post-denitrification single sludge biological nitrogen removal system. (b) The modified Ludzack–Ettinger single sludge biological nitrogen removal system proposed by Barnard (1973), including the primary anoxic reactor only. (c) The four-stage Bardenpho single sludge biological nitrogen removal system, including primary and secondary anoxic reactors.
Biological Nutrient Removal
system (i.e., the fraction of the mass of sludge in the system that is in the anoxic reactor) must be large and this may cause, depending on the sludge age, cessation of nitrification. Thus, although theoretically the system has the potential to remove all the nitrate, from a practical point this is not possible because the anoxic mass fraction will need to be so large that the conditions for nitrification cannot be satisfied particularly if the temperatures are low (o15 1C). Furthermore, in the anoxic reactor, ammonia is released through organism death and lysis, some of which passes out with the effluent thereby reducing the total nitrogen removal of the system. To minimize the ammonia content of the effluent, a flash or re-aeration reactor sometimes is placed between the anoxic reactor and the SST. In this reactor, N2 gas is stripped from the mixed liquor to avoid possible sludge buoyancy problems in the SST and the ammonia is nitrified to nitrate to assist with compliance of ammonia standards but it reduces the overall efficiency of the nitrate reduction of the system. For these reasons, post-denitrification has not been widely applied in practice.
4.14.24.5.1 The Ludzack–Ettinger system Ludzack and Ettinger (1962) were the first to propose a single sludge ND system utilizing the biodegradable organics in the influent as organics for denitrification. It consisted of two reactors in series, partially separated from each other. The influent was discharged to the first, or primary anoxic reactor which was maintained in an anoxic state by mixing without aeration. The second reactor was aerated and nitrification took place in it. The outflow from the aerobic reactor passed to the SST and the SST underflow was returned to the aerobic (second) reactor. Due to the mixing action in both reactors, an interchange of the nitrified and anoxic liquors was induced. The nitrate which entered the primary anoxic reactor was denitrified to nitrogen gas. Ludzack and Ettinger reported that their system gave variable denitrification results, probably due to the lack of control of the interchange of the contents between the two reactors. In 1973, Barnard proposed an improvement to the Ludzack– Ettinger system by completely separating the anoxic and aerobic reactors, recycling the underflow from the SST to the primary (first) anoxic reactor and providing a mixed liquor recycle from the aerobic to the primary anoxic reactor (Figure 34(b)). These modifications allowed a significant improvement in control over the system N removal performance of the system with the mixed liquor recycle flow. The RBSO and BPO from the influent stimulated high rates of denitrification in the primary anoxic reactor and much higher reductions of nitrate could be achieved than with post-denitrification, even when the pre-denitrification reactor of this system was substantially smaller than the post-denitrification reactor. In this system, called the Modified Ludzack–Ettinger (MLE) system, complete nitrate removal cannot be achieved because a part of the total flow from the aerobic reactor is not recycled to the anoxic reactor but exits the system with the effluent. To reduce the possibility of flotation of sludge in the SST due to denitrification of residual nitrate, the sludge accumulation in the SST needed to be kept to a minimum. This was achieved by having a high underflow recycle ratio from the SST, equal to the mean influent flow (1:1).
481
4.14.24.5.2 The four-stage Bardenpho system In order to overcome the deficiency of incomplete nitrate removal in the MLE system, Barnard (1973) proposed including a secondary anoxic reactor in the system and called it the fourstage Bardenpho system (Figure 34(c)). Barnard considered that the low concentration of nitrate discharged from the aerobic reactor to the secondary anoxic reactor will be denitrified to produce a relatively nitrate-free effluent. He included a flash or re-aeration reactor to strip the nitrogen gas and to nitrify the ammonia released during the denitrification. Although in concept the Bardenpho system has the potential for complete removal of nitrate, in practice this is not possible except when the influent TKN/COD concentration ratio is quite low, o0.09 mgN/mgCOD for normal municipal wastewater at 14 1C. The low denitrification rate and ammonia release (about 20% of the nitrate denitrified) results is an inefficient use of the secondary anoxic sludge mass fraction. As a result of the competition between the aerated and unaerated sludge mass fractions from the requirement to nitrify, (Section 4.14.20.3) usually it is better to exclude the secondary anoxic (and re-aeration) reactor and enlarge the primary anoxic reactor and increase the mixed liquor recycle ratio.
4.14.25 Denitrification Kinetics 4.14.25.1 Denitrification Rates The denitrification behavior in the primary and secondary anoxic zones is best explained by considering these reactors as plug-flow reactors. However, the explanation is equally valid for completely mixed reactors because the denitrification kinetics are essentially zero order with respect to nitrate concentration (van Haandel et al., 1981; Ekama and Wentzel, 1999b). Owing to the two different kinds of biodegradable organics (RBSO and BPO) in the influent wastewater, the denitrification in the primary anoxic reactor follows two phases (Figure 35(a)) – an initial rapid phase where the rate is defined by the simultaneous utilization of RBSO and BPO (K1 þ K2) and a second slower phase where the specific denitrification rate (K2) is defined by the utilization of only BPO originating from the influent and self-generated by the sludge through organism death and lysis. In the secondary anoxic reactor, only a single slow phase of denitrification takes place (Figure 35(b)), the specific rate (K3) being about two-thirds of the slow rate (K2) in the primary anoxic reactor (Stern and Marais, 1974; van Haandel et al., 1981). In the preceding aerobic reactor all the RBSO and most of the BPO of the influent has been utilized with the result that in the secondary anoxic reactor the only biodegradable COD available is BPO from organism death and lysis; the slow rate of supply of this BPO governs the K3 rate and causes this rate to be slower than the K2 rate. The values of the K rates are given in Table 13. A further specific K rate (K4) has been defined for denitrification in intermittently aerated anoxic aerobic digestion of WAS (Warner et al., 1986). This rate is only two-thirds of the K3 rate in the secondary anoxic reactor (Table 13), but sufficiently high to denitrify all the nitrate generated in aerobic digestion of WAS if the 6 h aeration cycle is 50% anoxic and 50% aerobic. Denitrification in anoxic–aerobic digestion adds
482
Biological Nutrient Removal
NO3−N concentration
NO3−N concentration
K1XBH
K2XBH
1st
(a)
K3XBH
Single phase
Second phase
(b)
Time
Time
Figure 35 Nitrate concentration of vs. time profiles in primary anoxic (a) and secondary anoxic (b) plugflow reactors, showing the three phases of denitrification associated with the K1, K2, and K3 rates. In the primary anoxic the initial rapid rate K1 is attributable to the utilization of the influent RBSO and the second slower rate K2 to the utilization of BPO from the influent wastewater and self-generated by organism death and lysis. In the secondary anoxic reactor, the rate K3 is attributable to the utilization of the self-generated BPO only. Table 13
K denitrification rates and their temperature sensitivity
Symbol
20 1 C
y
14 1 C
22 1 C
Equation
K120a K220a K320a K420a
0.72 0.1 0.1 0
1.2 1.08 1.029 1.029
0.241 0.06 0.06 0.04
1.036 0.118 0.08 0.05
158 159 160 161
a
Units – mgNO3-N/(mgOHOVSS d).
the benefits of denitrification to this system, that is, zero alkalinity consumption, oxygen recovery, improved pH control, reduced chemical dosing (Dold et al., 1985), and additionally a nitrogen free dewatering liquor. This last advantage is significant considering the high N content of WAS compared with primary sludge. The constancy of K1, K2, K3 (and K4) specific denitrification rates under constant flow and load conditions can be explained in terms of the kinetics of RBSO and BPO organics utilization included in the AS simulation models such as ASM1 developed later. The utilization of RBSO organics is modeled with the Monod equation and expressing the K1 rate in terms of this yields
K1 ¼
ð1 f cv YH Þf cv mHm Ss 2:86YH Ks þ Ss
where
Ss E 1 ðmgNO3 -N=ðmgOHOVSS dÞÞ Ks þ Ss
ð158Þ
In the plugflow and completely mixed primary anoxic reactor, the Monod term SS/(KS þ SS) remains close to 1 down to low RBSO concentrations because the half-saturation concentration (KS) is low. Accepting YH ¼ 0.45 mgVSS and
fcv ¼ 1.48 mgCOD/mgVSS yields K1 ¼0.26 mH mgNO3-N/ (mgOHOVSS d). So for the measured K1 ¼0.72 mgNO3-N/ (mgOHOVSS d) (Table 13), the mHm must have been about 2.8 d1. This mHm rate is in the range of mHm rates measured in AS systems. In investigating the kinetics of RBSO utilization in aerobic and anoxic selectors, Still et al. (1996) and Ekama et al. (1996a, b) found mHm values ranged between 1.0 d1 in completely mixed reactor systems and 4.5 d1 selector reactor systems, which yields K1 denitrification rates around 0.26 mgNO3-N/(mgOHOVSS d) for completely mixed type systems and 1.17 mgNO3-N/(mgOHOVSS d) for systems in which a selector effect (high mH) has been stimulated in the OHO biomass. The utilization of BPO is expressed in terms of the activesite surface hydrolysis kinetic formulation, which has the form of a Monod equation, except the variable is the adsorbed BPO to active OHO ratio (Xs/XBH), not the bulk liquid BPO concentration. Hence, the K2, K3 (and K4) rates are given by
K2 ¼ K3 ¼ K4 ¼
ð1 f cv YH Þ ZKh ðXs =XBH Þ 2:86f cv YH ½Kx þ ðXs =XBH Þ mgNO3 -N=ðmgOHOVSS dÞ
ð159Þ
where XS/XBH is progressively lower in primary (K2) secondary (K3) and anoxic–aerobic digestion (K4).
Biological Nutrient Removal
In the constant flow and load primary and secondary anoxic plugflow reactors, the (Xs/XBH) ratio changes very little due to the reduced anoxic hydrolysis rate including the Z. The reason for the K2 being higher than K3 arises from different concentrations of adsorbed BPO relative to the active OHO concentration (Xs/XBH) (Figure 36). In the primary anoxic reactor, the ratio is high because adsorbed BPO originates from the influent and OHO death. In the secondary anoxic, the ratio is lower because BPO originates only from OHO death. For the K2 and K3 denitrification rates, there is no simple relationship between the K rates and the ZKh because the adsorbed BPO to OHO ratio (Xs/XBH) is different in the primary and secondary anoxic reactors (and aerobic digester) and varies somewhat with sludge age and unaerated sludge mass fraction. It was concluded that the K1, K2, K3, and K4 denitrification constants have no direct kinetic significance; their constancy is the result of a combination of kinetic reactions which show little variation with sludge age in the range 10–30 days. Temperature does affect the K rates but once these have been adjusted for temperature, again the K rates show little variation at different sludge ages (van Haandel et al., 1981). It can be concluded both from experimental observation and theoretical kinetic points of view that acceptance of constant K2 and K3 rates is acceptable for steady-state design. This is in fact done to estimate the denitrification potential (Dp) of an anoxic reactor under constant flow and load conditions. With regard to K1, this rate can change significantly because the RBSO utilization rate can change appreciably depending on the mixing regime in the anoxic (or aerobic) reactor (Ekama et al., 1986, 1996a, b and Still et al., 1996). However, its variation does not affect ND design significantly because normally primary anoxic reactors are sufficiently large to allow complete utilization of RBSO even when the utilization rate (mHm) is low. In fact, the denitrification design procedure requires that all the RBSO is utilized in the primary anoxic reactor which introduces a minimum primary anoxic sludge mass fraction (fx1 min) and a minimum a-recycle ratio (amin) to
0.12
Specific denit rates (K )
K2 0.10 0.08 0.06
K3 K4
0.04 0.02 0.00 0.0
0.1
0.2
0.3
0.4
XS /XBH ratio (mgCOD/mgCOD) Figure 36 Specific denitrification rate (K) vs. adsorbed SB organics to OHO biomass concentration ratio (XS/XBH), showing the primary anoxic (K2), secondary anoxic (K3), and anoxic–aerobic digestion (K4) specific denitrification rates.
483
ensure this. These concepts can also be used for anoxic selector reactor design (Ekama et al.,1996a). The simulation model was applied also to anoxic–aerobic digestion of WAS. It was found that the model predicted accurately both aerobic and anoxic–aerobic digester behavior under constant and cyclic flow and load conditions and validated the K4 specific denitrification rate (Warner et al., 1986); no significant adjustment to values of the kinetic constants was necessary.
4.14.25.2 Denitrification Potential The concentration of nitrate (per liter influent flow Qi) that an anoxic reactor can denitrify biologically is called that reactor’s denitrification potential. It is called a potential because whether or not it is achieved depends on the nitrate load on the anoxic reactor(s). If too little nitrate is recycled to the anoxic reactor, all the recycled nitrate will be denitrified and the actual removal of nitrate, that is, denitrification performance, will be lower than the potential. In this case the denitrification is system (or recycle) limited. An increase in the system recycle ratios will increase nitrate load on the anoxic reactor and hence also the denitrification. Once the recycle rates are such that the nitrate load on the anoxic reactor(s) equals the denitrification potential of the reactor, then the system denitrification performance is optimal and the recycle ratios are at their optimum values. At this point the anoxic and aerobic reactor nitrate concentrations are just zero and the lowest possible, respectively. Increasing the recycle rates above the optimum increases the nitrate concentration in the anoxic reactor outflow above zero but this does not improve the denitrification performance because the system has now become biological or kinetics limited. The denitrification potential of the anoxic reactor(s) has been achieved and more nitrate cannot be denitrified by the particular anoxic reactors and wastewater. Indeed, increases in the recycle ratios above the optimum values are uneconomical due to unnecessary pumping costs and introduce unnecessary additional DO into the anoxic reactors which causes an undesirable reduction in denitrification performance and increase in effluent nitrate concentration. The principle of denitrification design therefore hinges around (1) calculating the denitrification potential of the anoxic reactor(s); (2) setting the nitrate load imposed on the anoxic reactor(s) equal to the denitrification potential; and (3) calculating the recycle ratios associated with this condition. The recycle ratios so calculated are the optimum values. From the above discussion, it is clear that critical in the design for denitrification is calculation of the nitrate load and denitrification potential. The nitrate load is calculated from the nitrification capacity, which is the concentration of nitrate per liter influent flow (Qi) generated by nitrification (Section 4.14.22.2, Equation (152)). The nitrification capacity (Nc, mg N l1 influent) was shown above to be approximately proportional to the influent TKN concentration (Nti). The denitrification potential is calculated separately for the utilization of the RBSO and BPO. The RBSO gives rise to a rapid denitrification rate so that it can be assumed that it is all utilized in the primary anoxic reactor. This is in fact an objective in the design. Accordingly, the contribution of the RBSO to the denitrification potential is simply the catabolic component of its
484
Biological Nutrient Removal
electron-donating capacity in terms of nitrate as N. Therefore, in the complete utilization of the influent RBSO, a fixed proportion (1 fcvYH) of the RBSO electrons (catabolic component) will be donated to NO3 reducing it to N2. Thus, knowing the influent RBSO concentration and assuming it is all utilized, the denitrification potential of this RBSO is given by
Dp1RBSO ¼ f Sb0 s Sbi ð1 f cv YH Þ=2:86 1
ðmgNO3 -N l
influentÞ
components of the RBSO and BPO yields the total denitrification potential of primary and secondary anoxic reactors, that is,
Dp1 ¼ Dp1RBSO þ Dp1BPO ¼ f Sb0 s Sbi ð1 f cv YH Þ=2:86 þ Sbi K2 f x1 YH Rs =ð1 þ bH Rs Þ ¼ Sbi ff Sb0 s ð1 f cv YH Þ=2:86 þ K2 f x1 YH Rs =ð1 þ bH Rs Þg ðmgN l1 influentÞ
ð163Þ
ð160Þ Dp3 ¼ Dp3RBSO þ Dp3BPO
where Dp1 RBSO is the denitrification potential of the influent RBSO in primary anoxic reactor, Sbi the influent biodeg. COD (mgCOD l1), fSb0 s the RBSO fraction of Sbi, YH the OHO yield coefficient (0.45 mgVSS/mgCOD), and 2.86 the oxygen equivalent of nitrate. For the BPO, this substrate contributes to denitrification in the primary anoxic reactor and the secondary anoxic reactor. The denitrification potentials for the BPO are formulated on the basis of the K2 and K3 specific denitrification rates, respectively. These K rates are a simplification of the kinetic equations describing the utilization of BPO from the influent and/or from organism death and lysis and have a basis in the fundamental biological kinetics incorporated in the AS simulation models such as ASM1 (Henze et al., 1987). The K rates define the denitrification rate as mgNO3 -N denitrified per day per mgOHOVSS mass in the anoxic reactor. To determine the denitrification potential contributed by the BPO, the mass of OHOVSS produced per liter influent flow and the proportion of this mass in the primary and/or secondary anoxic reactors needs to be calculated and multiplied by the K2 or K3 rates. From the steady-state AS model for organics removal (Section 4.14.9.3), the OHO mass in the system (MXBH) is calculated from the biodegradable COD load (Equation (103)). Of this MXBH mass, a fraction fx1 and/or fx3 is continuously present in the primary and/or secondary anoxic reactors, respectively, that is, fx1 and fx3 are the primary and secondary anoxic sludge mass fractions, respectively. The OHOVSS mass in the primary and/or secondary anoxic reactors per liter influent flow is therefore given by
f x1 MXBH =Qi ¼ f x1 Sbi ðYH ÞRs =ð1 þ bH Rs Þ ðmgOHOVSS l1 influent in primary anoxicÞ
f x3 MXBH =Qi ¼ f x3 Sbi ðYH ÞRs =ð1 þ bH Rs Þ ðmgOHOVSS l1 influent in secondary anoxicÞ
Multiplying these masses by the respective K rates gives the primary and secondary anoxic reactor denitrification potentials attributable to BPO (Dp1BPO, Dp3BPO), viz.,
Dp1BPO ¼ K2 f x1 MXBH =Qi ¼ K2 f x1 Sbi YH Rs =ð1 þ bH Rs Þ ð161Þ Dp3BPO ¼ K3 f x3 Sbi YH Rs =ð1 þ bH Rs Þ
ð162Þ
This approach is valid because the K2 and K3 rates are continuous for the entire sludge residence time in the anoxic reactor(s), provided the nitrate concentration does not decrease to zero (Figure 35). Combining the denitrification potential
¼ 0 þ Sbi K3 f x3 YH Rs =ð1 þ bH Rs Þ ðmgN l1 influentÞ ð164Þ In Equations (163) and (164), the K2, K3, and bH rates are temperature sensitive, decreasing as the temperature decreases. The temperature sensitivity of these rates has been measured and is defined in Tables 13 and 6. From Equations (163) and (164), it can be seen that the denitrification potentials are directly proportional to the biodegradable COD concentration of the wastewater (Sbi). This is expected because in the same way that the oxygen demand is directly related to the COD load, so also is the nitrate demand (which is called the denitrification potential) because both oxygen and nitrate act as electron acceptor for the same organic degradation reactions. For the same size anoxic reactor, the primary anoxic has a much larger denitrification potential (by about 2–3 times) than the secondary anoxic because (1) K2 is larger than K3 and (2) more importantly, the RBSO makes a significant contribution to the denitrification potential in the primary anoxic reactor. For this reason the RBSO needs to be accurately specified to ensure reliable estimates of the N removal that can be achieved. For a normal municipal wastewater with an RBSO fraction (fSb0 s) of about 25% (with respect to biodegradable COD), the RBSO contributes about one-third to half of Dp1 depending on the size of the primary anoxic reactor and temperature. In a system where a high degree of N removal is required, between one-fourth and one-third of the carbonaceous oxygen demand is met by denitrification, which reduces the carbonaceous oxygen demand in the aerobic reactor by the same amount. As mentioned earlier, this reduction represents about half of the oxygen that was required to produce the nitrate by nitrification (see Figure 32). From Equation (164), the influent RBSO contribution to the denitrification potential of the secondary anoxic reactor is zero. This is because all the RBSO is utilized either in the preceding primary anoxic and/or in aerobic reactors. However, the Dp3 RBSO term has been included in Equation (164) in the event an external carbon source such as methanol, acetic acid, or high-strength organic wastewater is dosed into the secondary anoxic reactor to improve the denitrification. The sludge mass fraction approach above is valid because the fraction of the VSS (MXv) or TSS (MXt) masses that is OHO mass (MXBH) is constant for specified wastewater characteristics and sludge age and equal to the active fraction (fatOHO or favOHO – Equations (114) and (115)) and very closely the same in the anoxic and aerobic reactors of the system. Therefore, the anoxic and aerobic sludge mass fractions are the same whether calculated from the VSS, TSS, or OHO masses; for example, in an MLE system with anoxic and aerobic reactor volumes
Biological Nutrient Removal of 3000 and 6000 m3, respectively, one notes that nearly one-third of the OHO, VSS, and TSS masses in the system are in the anoxic reactor, and hence the anoxic sludge mass fraction is 0.33.
4.14.25.3 Minimum Primary Anoxic Sludge Mass Fraction In Equation (163), it is assumed that the initial rapid rate of denitrification is always complete, that is, the actual retention time in the primary anoxic reactor is always longer than the time required to utilize all the influent RBSO. This is because in Equation (163), the denitrification attributable to the influent RBSO is stoichiometrically expressed, not kinetically – it gives the concentration of nitrate the K1 rate removes when allowed sufficient time to reach completion. By considering the actual retention time (say t1) required to complete the first phase of denitrification (Figure 35(a)), and noting that t1(a þ s þ 1) is the minimum nominal hydraulic retention time to achieve this, it can be shown that the minimum primary anoxic sludge mass fraction fx1min to remove all the influent RBSO at a rate of K1 mgNO3-N/(mgOHOVSS d) is
f x1min ¼
f Sb0 s ð1 f cv YH Þð1 þ bHT Rs Þ 2:86K1T YH Rs
ð165Þ
Substituting the values of the kinetic constants into Equation (165), yields fx1mino0.08 for Rs 410 days at 14 1C. This value is much lower than most practical primary anoxic reactors so that Equation (163) will be valid in most cases. Equation (165) also applies to sizing anoxic selectors provided K1 (or mH) is appropriately selected (see Section 4.14.25.1, Equation (158); Ekama et al., 1996a).
4.14.25.4 Denitrification – Influence on Reactor Volume and Oxygen Demand From the design approach to nitrification (Equation (136)) and denitrification (Equations (163) and (164)), it can be seen that the design for N removal is done entirely using sludge mass fractions and does not require the volume of the reactor to be known. The volume of the reactor is obtained in the identical fashion as for the fully aerobic system and follows from the choice of the TSS concentration (Xt) for the reactor (Section 4.14.11). The volume of the reactor so obtained is then subdivided in proportion to the calculated primary and/or secondary anoxic and aerobic sludge mass fractions. Consequently, N removal design is grafted directly into the aerobic system design and for the same design reactor TSS concentration and sludge age, a fully aerobic system and an anoxic–aerobic system for N removal will have the same reactor volume. Research has indicated that there are many factors that influence the mass of sludge generated for a given sludge age and daily average COD load, and alternating anoxic–aerobic conditions is one of them. However, relative to the uncertainty in organic (COD) load and unbiodegradable particulate COD fraction and their daily and seasonal variation, these influences are not large enough from a design point of view to be given specific consideration in the design procedure. From a design point of view, the only significant difference between aerobic and anoxic–aerobic conditions is the oxygen demand
485
and this difference needs to be taken into account for economical design (Figure 32).
4.14.26 Development and Demonstration of Design Procedure It was concluded above that the influent wastewater characteristics that need to be accurately known are the influent TKN/COD ratio and RBSO fraction. These have a major influence on the nitrification capacity and denitrification potential, respectively, and hence on the N removal performance and minimum effluent nitrate concentration that can be achieved by biological denitrification. The effect of these two wastewater characteristics on design will be demonstrated below with numerical examples generated from the example raw and settled wastewaters with different influent TKN concentrations and RBSO fractions. The design of biological N removal is developed and demonstrated below by continuing the calculations with the example raw and settled wastewater characteristics listed in Tables 7 and 11. The only additional characteristic required for the denitrification design is the influent RBSO fraction (fSb0 s), which is 0.25 and 0.385 of the biodegradable COD for the raw and settled wastewaters, respectively. The results obtained so far for the COD removal and nitrification calculations for sludge ages 3–30 days are shown in Figures 14, 15, and 31.
4.14.26.1 Review of Calculations For the raw wastewater characteristics (i.e., fS0 up ¼ 0.15 mgCOD/mgCOD, fS0 us ¼ 0.07 mgCOD/mgCOD, Tmin ¼ 14 1C, Sti ¼ 750 mgCOD l1 – see Table 7) and 20 days sludge age, and accepting the nitrogen content of the volatile solids (fn) to be 0.10 mgN/mgVSS, the nitrogen required for sludge production Ns ¼ 17.0 mgN l1 (Equation (144)). From Section 4.14.23.2, the effluent biodegradable and unbiodegradable soluble organic nitrogen concentrations (Nobse and Nouse) are 0.0 and 1.80 mgN l1, respectively. From Equation (132) the effluent ammonia concentration Nae is 2.0 mgN l1. The effluent TKN concentration (Nte) is the sum of Nouse and Nae (Equation (150)) and hence Nte ¼ 3.8 mgN l1. The nitrification capacity (Nc) is found from Equation (152) and for the example raw wastewater (Nti ¼ 60.0 mgN l1; TKN/COD ¼ 0.08 mgN/mgCOD) at 14 1C is
Nc ¼ 60:0 17:0 3:8 ¼ 39:2 mgN l1 The nitrification oxygen demand, FOn is found from Equation (155), that is,
FOn ¼ 4:57Nc Qi ¼ 4:57 39:2 15 106 mgOd1 ¼ 2687 kgOd1 and the mass of nitrifier VSS in the reactor is given by Equation (154), that is,
MXBA ¼ 0:1 20=ð1 þ 0:034 20Þ 39:2 15 ¼ 702 kgVSS
486
Biological Nutrient Removal
In the design, because it is intended to reduce the nitrate concentration as much as possible, the alkalinity change in the wastewater will be minimized; assuming that 80% of the nitrate formed is denitrified, the H2CO3* alk change ¼ 7.14Nc 3.57 (nitrate denitrified) ¼ 7.14 39.2 þ 3.57 0.80 39.2 ¼ 168 mg l1 as CaCO3. With an influent H2CO3* alk of 250 mg l1 as CaCO3 the effluent H2CO3* alk ¼ 250–168 ¼ 82 mg l1 as CaCO3, which, from Figure 27, will maintain a pH above 7 (see Section 4.14.20.6).
4.14.26.2 Allocation of Unaerated Sludge Mass Fraction In nitrogen removal systems, the maximum anoxic sludge mass fraction available for denitrification, fxdm, can be set equal to the maximum unaerated sludge mass fraction fxm at the minimum temperature, that is,
f xdm ¼ f xm
ð166Þ
where fxm is given by Equation (136) for selected Rs, mnmT, and Tmin. This is because for N removal systems, unaerated sludge mass need not be set aside for the anaerobic reactor. In N and P removal systems, some of the unaerated sludge mass (0.12–0.25) needs to be set aside for the anaerobic reactor to stimulate BEPR. This sludge mass fraction, called the anaerobic sludge mass fraction and denoted fxa, cannot be used for denitrification. For BEPR to be as high as possible, no nitrate should be recycled to the anaerobic reactor so that zero denitrification takes place in this reactor. So, for the purposes of this development and demonstration of denitrification behavior, it will be accepted that the maximum unaerated sludge mass fraction available at 20 days sludge age (fxm) is all allocated to anoxic conditions, that is, fxdm ¼ fxm ¼ 0.534.
4.14.26.3 Denitrification Performance of the MLE System 4.14.26.3.1 Optimum recycle ratio a In the MLE system, the anoxic sludge mass fraction is all in the form of a primary anoxic reactor, that is, fx1 ¼ fxdm ¼ fxm. The denitrification potential of the primary anoxic reactor Dp1 is found from Equation (163), that is, for the example raw wastewater at 14 1C and fxm ¼ fxdm ¼ fx1 ¼0.534, Dp1 ¼ 52.5 mgN l1. The only additional wastewater characteristic required to calculate Dp1 is the influent RBSO (Sbsi) concentration or fraction (fSb’s), which for the example raw and settled wastewaters are given in Table 14, that is, 0.25 and 0.385 with respect to the biodegradable COD (Sbi), respectively. In the MLE system, if the nitrate concentration in the outflow of the anoxic reactor is zero, then the nitrate concentration in the aerobic reactor (Nnar) is equal to Nc/ (a þ s þ 1), that is, the nitrification capacity in mgN l1 influent flow diluted by the total (no nitrate containing) flow entering the aerobic reactor which is (a þ s þ 1) times the influent flow, where a and s are the mixed liquor and underflow recycle ratios (with respect to the influent average dry weather flow Qi), respectively. Accepting that there is no denitrification in the secondary settling tank (which needs to be minimized anyway due to the problem of rising sludges), the aerobic reactor and system effluent nitrate concentrations (Nnar and
Table 14 Additional wastewater characteristics required for denitrification (and BEPR) design Wastewater Readily biodegradable soluble organics (RBSO) as.y (1)y.fraction of biodegradable organics (BO, Sbi) COD (fSb0 s) (2)y.fraction of total organics (Sti, COD) (fS0 bs) VFA fraction of biodegradable soluble organics (RBSO), (fSbs0 a)
Raw
Settled
0.25
0.385
0.194
0.324
0.10
0.10
Nne, respectively) are equal and given by
Nne ¼ Nnar ¼ Nc =ða þ s þ 1Þ
ð167Þ
Knowing Nne and Nnar and taking into account DO concentrations in the a and s recycles, that is, Oa and Os mgO l1 respectively, the equivalent nitrate load on the primary anoxic reactor (Nnlp) by the a and s recycles is
Oa Os a þ Nne þ s Nnlp ¼ Nnar 2:86 2:86 The optimum denitrification (i.e., lowest effluent nitrate concentration) is obtained when the equivalent nitrate load on the anoxic reactor is equal to the denitrification potential of the anoxic reactor (i.e., Dp1 ¼ Nnlp), viz.,
Dp1 ¼
Nc Oa Nc Os þ þ aþ s ð168Þ ða þ s þ 1Þ 2:86 ða þ s þ 1Þ 2:86
Solving Equation (168) for a yields the a recycle ratio which exactly loads the primary anoxic reactor to its denitrification potential with nitrate and DO. This a value is the optimum because it results in the lowest Nne, that is,
aopt ¼ ½B þ
pffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi B 2 þ 4AC=ð2AÞ
ð169Þ
where
A ¼ Oa =2:86 B ¼ Nc Dp1 þ fðs þ 1ÞOa þ sOs g=2:86 C ¼ ðs þ 1ÞðDpp sOs =2:86Þ sNc and
Nnemin ¼ Nneaopt ¼ Nc =ðaopt þ s þ 1Þ ðmgN l1 Þ
ð170Þ
For a ¼ aopt, Equation (170) for Nne is valid and will give the minimum Nne attainable. When a raopt Equation (170) is also valid because the assumption on which Equation (169) is based is valid, that is, NnlprDp1 or equivalently, zero nitrate concentration in the outflow of the anoxic reactor. For a 4aopt this assumption is no longer valid and Nne increases as the a recycle ratio increases due to increasing DO flux entering the anoxic reactor. For a4aopt, Nne is given by the difference between the equivalent nitrate load on the anoxic reactor (which is the sum of the nitrification capacity Nc and the nitrate
Biological Nutrient Removal MLE system (settled) Effluent nitrate and a recycle ratio
MLE system (raw) Effluent nitrate vs. a recycle
20
487
20
14 °C s = 0.5 s = 1.0 s = 2.0 22 °C s = 1.0
15
10
Effluent nitrate (mgN I−1)
Effluent nitrate (mgN I−1)
a-opt (14 °C)
a-opt (14 °C) a-opt (22 °C)
5
14 °C s = 0.5 s = 1.0 s = 2.0
15
22 °C s = 1.0
10
N ne min (14 °C)
5
N ne min (14 °C)
0 0
0 5
10 a Recycle ratio
15
0
5
10 a Recycle ratio
15
Figure 37 Effluent nitrate concentration vs. mixed liquor a recycle ratio for the example raw (a) and settled (b) wastewaters for underflow (s) recycle ratio of 1:1 at 14 1C (bold line) and 22 1C (thin line) and for s ¼ 0.5:1 and 2.0:1 at 14 1C (dashed lines).
equivalent of the oxygen concentration with respect to the influent flow) and the denitrification potential Dp1, viz.,
Nne ¼ Nc þ
aOa sOs þ Dp1 2:86 2:86
ðmgN l1 Þ
ð171Þ
As Nc, Dp1, Os, and Oa are constants, the increase in Nne with increasing a above aopt is linear with slope Oa/2.86 mgN l1. At a ¼ aopt, Equations (170) and (171) give the same Nne concentrations. Accepting the design sludge age of 20 days, which allows a maximum unaerated sludge mass fraction fxm of 0.534, the denitrification behavior of the MLE system is demonstrated below for the example raw and settled wastewaters at 14 and 22 1C. In the calculations the DO concentrations in the a and s recycles, Oa and Os are 2 and 1 mgO l1, respectively, and the underflow recycle ratio s is 1:1. This s recycle ratio is usually fixed at a value such that satisfactory settling tank operation is obtained. Details of secondary settling tank theory, design, modeling, and operation are discussed by Ekama et al. (1997) and Ekama and Marais (2004). Substituting the values for the nitrification capacity Nc and denitrification potential Dp1 into Equations (169) and (170), the optimum mixed liquor recycle ratio aopt and minimum effluent nitrate concentration Nneaopt are obtained, for example, for the settled wastewater at 14 1C
A ¼ 2=2:86 ¼ 0:70 B ¼ 39:6 40:1 þ fð1 þ 1Þ2 þ 1 1g=2:86 ¼ þ1:52 C ¼ ð1 þ 1Þð40:1 1 1=2:86Þ 1 39:6 ¼ þ39:61 Hence, aopt ¼ 6.5 and Nnemin ¼ 4.7 mgN l1. The calculations for the example raw and settled wastewater at 14 and 22 1C show that for all four cases aopt exceeds 5. Although the calculations include the discharge of DO to the anoxic reactor, a recycle ratios above 5 to 6 are not cost effective. The small decreases in Nne which are obtained for even large increases in a recycle ratio above about 5:1 do not warrant the additional pumping costs.
This is illustrated in Figure 37 which shows Nne versus a recycle ratio for the example raw (Figure 37(a)) and settled (Figure 37(b)) wastewater at 14 and 22 1C plotted from Equations (170) and (171). For the settled wastewater (Figure 37(b)) at 14 1C and s ¼ 1:1, for aoaopt, the anoxic reactor is underloaded with nitrate and DO and as the a recycle increases up to aopt, the equivalent nitrate load increases toward the anoxic reactor’s denitrification potential. Initially, Nne decreases sharply for increases in a, but as a increases the decrease in Nne becomes smaller. At 14 1C with a ¼ aopt ¼ 6.5, the anoxic reactor is loaded to its denitrification potential by the a and s recycles and a Nnemin ¼ Nneaopt ¼ 4.7 mgN l1 is achieved. At a ¼ aopt ¼ 6.5, the greatest proportion of the anoxic reactor’s denitrification potential is used for denitrification and therefore yields the minimum effluent nitrate concentration (Nneaopt). This is shown in Figures 38(a) and 38(b) for the raw and settled wastewaters at 14 1C. For the settled wastewater at 14 1C (Figure 38(b)) at a ¼ aopt ¼ 6.5, 88% of the equivalent nitrate load (i.e. (a þ s) Nnemin ¼ 35.2 mgN l1 out of a Dp1 ¼ 40.1 mgN l1) is nitrate and therefore 88% of the denitrification potential of the anoxic reactor is utilized for denitrification and 12% for DO removal. The higher the a recycle ratio, the greater the proportion of the denitrification potential is utilized for DO removal. At 14 1C, for a4aopt, the equivalent nitrate load exceeds the denitrification potential and as the a recycle increases so Nne increases due to the increased DO mass flow to the anoxic reactor. From Equation (171), at a ¼ 15, Nne ¼ 10.6 mgN l1 and 27% of the denitrification potential is required to remove DO, leaving only 73% for denitrification (Figures 37(b) and 38(b)). For 14 1C, the plots of Nne versus a at underflow s recycle ratios of 0.5:1 and 2.0:1 are also given in Figure 37 and show that aopt is not significantly different at different s recycle ratios. Also, at low a recycle ratios, changes in s have a significant influence on Nne, but at high a recycle ratios, even significant changes in s do not significantly change Nne. This is because at high a, most of the nitrate is recycled to the anoxic reactor by the a recycle, so that changes in s do not
488
Biological Nutrient Removal MLE system (raw) use of denitrification potential 100
Denitrification potential used for nitrate removal
20
% Denit. potential
% Denit. potential
Denitrification potential used for nitrate removal
0
Denitrification potential used for nitrate removal
Denitrification potential used for nitrate removal
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a-opt (14 °C) a-prac
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Used DO removal
Unused denitrification potential
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MLE system (settled) use of denitrification potential
5
10 a Recycle ratio
15
0 (b)
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10 a Recycle ratio
15
Figure 38 % Denitrification potential unused, used by dissolved oxygen in the recycles and for denitrification vs. a recycle ratio for the example raw (a) and settled (b) wastewaters for underflow (s) recycle ratio of 1:1 at 14 1C.
significantly change the nitrate load on the anoxic reactor. Hence, for the MLE system, decreases in s can be compensated for by increases in a – it makes little difference which recycle brings the nitrate to the anoxic reactor as long as the anoxic reactor is loaded as closely as practically possible to its denitrification potential in order to minimize Nne. For the settled wastewater at 22 1C and s ¼ 1:1 (Figure 37(b)), Nne versus a is similar to that at 14 1C up to a ¼ 6.5. This is because Nc values at 14 and 22 1C for the example raw and settled wastewater are almost the same (i.e., 39.9 and 41.6 mgN l1 at 14 and 22 1C, respectively). However, at 22 1C, the denitrification potential is significantly higher than at 14 1C (40.1 mgN l1 at 14 1C and 52.4 mgN l1 at 22 1C) so that a higher aopt is required (e.g., 17.9) at 22 1C to load the anoxic reactor to its denitrification potential than at 14 1C. Therefore, at 22 1C, as the a recycle increases above 6.5, Nne continues to decrease until aopt ¼ 17.9 is reached. The increase in a from 6.5 to 17.9 reduces Nne from 4.9 to 2.1, that is, only 2.8 mgN l1. This small decrease in Nne is not worth the large increase in pumping costs from 6.5:1 to 17.9:1 required to produce it. Consequently, for economical reasons, the a recycle ratio is limited at a practical maximum (aprac) of say 5:1, which fixes the lowest practical effluent nitrate concentration (Nneprac) from the MLE system between 5 and 10 mgN l1 depending on the influent TKN/ COD ratio. From the design procedure demonstrated so far, it is clear that the procedure hinges around balancing the equivalent nitrate load with the denitrification potential by appropriate choice of the a recycle ratio: for selected system design parameters (sludge age, anoxic mass fraction, underflow recycle ratio, etc.) and wastewater characteristics (temperature, readily biodegradable COD fraction, TKN/COD ratio, etc.), the denitrification potential of the MLE system is fixed. With all the above fixed, the system denitrification performance is controlled by the a recycle ratio, and this performance is optimum when the a recycle ratio is set at the optimum aopt. For aoaopt, the performance will be below optimum because the equivalent nitrate load is less than the denitrification potential (Figure 38); for a ¼ aopt, the performance is optimal because the equivalent
nitrate load equals the denitrification potential; and for a4aopt, the performance is again suboptimal because now the equivalent nitrate load is greater than the denitrification potential and more than necessary DO is recycled to the anoxic reactor which reduces the denitrification (see Figures 37 and 38). If a practical limit on a is set at say aprac ¼ 5:1 and aopt is significantly higher, then a significant proportion of the anoxic reactor’s denitrification potential is not used (Figure 38). There are two options to deal with this unused denitrification potential: (1) change the design, that is, decrease the sludge age (Rs) and/or unaerated sludge mass fraction (fxm) or (2) leave the system as designed (i.e., Rs ¼ 20 days and fxm ¼ 0.534) and keep the unused denitrification potential in reserve as a factor of safety against changes in wastewater characteristics, such as (1) increased organic load, which will require a reduction in sludge age, (2) increased TKN/COD ratio, which will load the anoxic reactor with nitrate at lower a recycle ratios, or (3) decreased RBSO fraction, which decreases the anoxic reactors denitrification potential.
4.14.26.3.2 The balanced MLE system With option (1) the anoxic sludge mass fraction fx1 is decreased to eliminate the unused denitrification potential. The decrease in fx1 increases the aerobic mass fraction and therefore the factor of safety (Sf) on nitrification. To maintain the same Sf, the sludge age of the system can be reduced to that value at which the lower fx1 is equal to the maximum unaerated sludge mass fraction fxm allowed (i.e., fx1 ¼ fxm) for the selected mAm20 and Tmin. An MLE system with a sludge age (Rs) and influent TKN concentration (Nti) such that fx1 ¼ fxm and aopt ¼ aprac (say 5:1), so that this aprac loads the anoxic reactor exactly to its denitrification potential, is called a balanced MLE system. This approach to design of the MLE system was proposed by van Haandel et al. (1982) and gives the most economical AS reactor design, that is, the lowest sludge age, and therefore the smallest reactor volume, and the highest denitrification with the a recycle ratio fixed at some maximum practical limit. The influent TKN/COD ratio, fxm ¼ fx1, fx1 min, Nne, and %N removal (%Nrem) versus sludge age for balanced
Biological Nutrient Removal MLE system (settled, 14 °C) Design at fixed a -opt = 5:1
MLE system (raw, 14 °C) Design at fixed a -opt = 5:1 Balanced sludge age
0.8
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Effluent nitrate (mgN I−1)
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MLE system (settled, 14 °C) Design at fixed a -opt = 5:1
MLE system (raw, 14 °C) Design at fixed a -opt = 5:1
(c)
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Anoxic mass fraction
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489
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Figure 39 Influent TKN/COD ratio (TKN/COD), maximum unaerated (fxm), primary anoxic (fx1), and minimum primary anoxic (fx1 min) sludge mass fractions (a,b) and effluent nitrate concentration and %N removal (c,d) for balanced MLE systems with a 5:1 practical upper limit to the a recycle ratio for the example raw (a,c) and settled (b,d) wastewaters at 14 1C.
MLE systems for the example raw and settled wastewaters at 14 and 22 1C are shown in Figures 39 and 40, respectively. The sludge age which balances the MLE system for given wastewater characteristics and aprac cannot be calculated directly. It is easiest to calculate the influent TKN concentration for a range of sludge ages and choose the sludge age which matches the wastewater TKN concentration (Nti). The procedure for calculating Nti for a balanced MLE system is as follows: from the design mAm20, Tmin, Sf, and a selected sludge age, fxm is calculated from Equation (136). Provided fxm4fx1 min (Equation (165)), fx1 is set equal to fxm. Knowing fx1 and the wastewater characteristics, Dp1 is calculated from Equation (163). This Dp1 and a selected value for aprac are then substituted into Equation (168), which sets the equivalent nitrate load on the anoxic reactor equal to the denitrification potential and hence aopt equals the selected aprac. With Dp1 and a known, Nc is calculated from Equation (168). Once Nc is known, Nti is calculated from Nti ¼ Nte þ Ns þ Nc (Equation (152)), where Nte ¼ Nouse þ Nae (Equation (150)) and Nae is given by Equation (132) because with Sf fixed the Rs fxm relationship is fixed. With Nc and Nti known, the effluent nitrate concentration Nne and % nitrogen removal (%Nrem) are
found from Equation (170) and %Nrem ¼ 100[Nti (Nne þ Nte)]/Nti, respectively. This calculation is repeated for different sludge ages. The shortest sludge age allowed is the one which gives fx1 ¼ fxm ¼ fx1min. In Figure 39, for 14 1C, for the raw wastewater (Figures 39(a) and 39(c)), it can be seen that fx1( ¼ fxm) increases from about 0.09 at 8 days sludge age, at which fxm is just greater than fx1 min, to 0.60 at 26 days sludge age, at which fxm is equal to the upper limit set for it. As fx1 increases so the influent TKN/COD ratio increases from 0.061 at 8 days sludge age to 0.115 at 26 days sludge age. With the increase in TKN/COD ratio, the nitrification capacity Nc increases and hence Nne increases from about 3.2 mgN l1 at 8 days sludge age to 9.3 mgN l1 at 26 days sludge age because the a and s recycle ratios remain at 5:1 and 1:1, respectively (see Equation (170)). The %N removal, which includes the N removed via sludge wastage Ns, decreases marginally from 85% to 82% as the influent TKN/COD ratio and sludge age increase for the balanced MLE system. For the settled wastewater at 14 1C (Figures 39(b) and 39(d)), the influent TKN/COD ratio, fx1 and fx1min results are similar to those for the raw wastewater, that is, for the same
Biological Nutrient Removal MLE system (raw, 22 °C) Design at fixed a-opt = 5:1 1.0
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TKN/COD Influent TKN/COD ratio
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MLE system (settled, 22 °C) Design at fixed a-opt = 5:1
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%N removal 15
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50 Effluent nitrate
5
25
Balanced sludge age
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% Nitrogen removal
490
0 0
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Figure 40 Influent TKN/COD ratio (TKN/COD), maximum unaerated (fxm), primary anoxic (fx1), and minimum primary anoxic (fx1 min) sludge mass fractions (a,b) and effluent nitrate concentration and %N removal (c,d) for balanced MLE systems with a 5:1 practical upper limit to the a recycle ratio for the example raw (a,c) and settled (b,d) wastewaters at 22 1C.
sludge age approximately the same TKN/COD ratio is found for the balanced MLE system. For the settled wastewater, the Nne is slightly lower, increasing from about 3.2 to 6.7 mgN l1 from 8 to 26 days sludge age; also the %N removal is somewhat lower, around 78% mainly due to the lower N removal via sludge wastage Ns. However, it must be remembered that the TKN/COD ratio and RBSO fraction of a settled wastewater are higher than those of the raw wastewater from which it is produced, viz. TKN/COD ratio 0.113 and 0.080 mgN/mgCOD and RBSO fraction (fSb’s) 0.25 and 0.385 for the example settled and raw wastewaters, respectively. Therefore, at 14 1C, while the raw wastewater can be treated in a balanced MLE system at about 11 days sludge age (Figure 39(a)), the sludge age for the settled wastewater balanced MLE system is about 17 days (Figure 39(b)). A comparison of the balanced MLE systems for the example raw and settled wastewaters is given in Table 15. From Table 15 it can be seen that Nne is less than 1 mgN l1 higher for the settled wastewater but the reactor volume and total oxygen demand significantly lower compared with the
Table 15 Comparison of balanced MLE systems treating the example raw and settled wastewaters at 14 1C Parameter
Raw
Settled
Influent TKN/COD ratio Unaerated mass fraction(fxm) Anoxic mass fraction (fx1) Minimum anoxic fraction a Recycle ratio (aprac ¼ aopt) Sludge age (days) Effluent nitrate (Nne, mgN l1) Effluent TKN (Nte, mgN l1) Reactor vol. at 4.5 gTSS l1 (m3) Carb O2 demand (FOc, kgO d1) Nit O2 demand (FOn, kgO d1) O2 recovered (FOd, kgO d1) Tot. O2 demand (FOtd, kgO d1) %N removal Mass TSS wasted (FXt, kg d1) Active fraction wrt TSS (fatOHO)
0.08 0.306 0.306 0.08 5:1 11 5.1 4.3 9484 6156 2492 1327 7321 84.3 3880 0.316
0.113 0.485 0.485 0.108 5:1 17 5.7 4.1 5264 4251 2685 1437 5499 80.9 1394 0.414
Biological Nutrient Removal
raw wastewater. Therefore, from an AS system point of view, treating settled wastewater would be more economical than treating raw wastewater for a comparable effluent quality. Also, both systems require sludge treatment; for the raw wastewater because 11 days sludge age waste sludge is not stable (high active fraction, favOHO) and for the settled wastewater, the primary sludge needs to be stabilized. The 11 days sludge age waste sludge can be stabilized with anoxic aerobic digestion which allows the N released in digestion to be nitrified and denitrified (Warner et al. 1986; Mebrahtu et al., 2010) and primary sludge can be anaerobically digested to benefit from gas generation. The choice of treating raw or settled wastewater therefore does not depend so much on the effluent quality or the economics of the AS system itself, but on the economics of the whole WWTP, including sludge treatment. Because the minimum wastewater temperature (Tmin) governs the AS system (and sludge treatment) design, the balanced MLE system results for 22 1C are not particularly relevant to the temperate climate regions. However, in equatorial and tropical regions, where wastewater treatment is becoming a matter of increasing concern, high wastewater temperatures are encountered. For this reason and for illustrative purposes also, the balanced MLE results for the raw and settled wastewaters are shown in Figure 40. Compared with 14 1C, the upper limit to fxm ¼ 0.60 is reached already at 7 days sludge age and significantly higher influent TKN/COD ratios can be treated at equal sludge ages. These higher TKN/COD ratios result in higher Nne, which for the raw wastewater increases from 3 to 13 mgN l1 and for the settled wastewater from 3 to 9 mgN l1 for increases in sludge age from 4 to 30 days. If Tmin were 22 1C, the example raw and settled wastewaters could be treated at 3 and 4 days sludge age, respectively, yielding Nne of 5 and 6.5 mgN l1, respectively. This reinforces the conclusion in Section 4.14.24.1 that in equatorial and tropical climates it is highly likely that AS plants will nitrify even at very short sludge ages (1–2 days) and therefore to design for denitrification for operational reasons if not for effluent quality reasons.
4.14.26.3.3 Effect of influent TKN/COD ratio When the unused denitrification potential in the anoxic reactor is kept in reserve as a safety factor (option 2), the sludge age and unaerated (anoxic) mass fraction are not changed. For this situation, it is useful to have a sensitivity analysis to see the influence of changing influent TKN/COD ratio and RBSO fraction on the a recycle ratio and effluent nitrate concentration. Continuing with the design for the example raw and settled wastewaters for fixed sludge age at 20 days and unaerated (anoxic) mass fraction at 0.534, a plot of the optimum a recycle ratio aopt and minimum effluent nitrate concentration Nneaopt for underflow recycle ratios s of 0.5, 1.0, and 2.0 versus influent TKN/COD ratio from 0.06 to 0.16 is given in Figure 41 for the raw (a), (c) and settled (b), (d) wastewaters at 14 1C (a), (b) and 22 1C (c), (d). From Figure 41, it can be seen that as the influent TKN/ COD increases, aopt decreases and Nneaopt increases. The aopt–Nneaopt lines in Figure 41 give the system denitrification performance when the denitrification potential of the anoxic reactor is fully used, that is, the system denitrification
491
performance is equal to its denitrification potential and the nitrate concentration is the lowest possible. Also, large increases in the underflow recycle ratio s (i.e., from 0.50:1 to 1.0:1 or 1.0:1 to 2.0:1) decrease aopt but do not change Nneaopt because the DO in the a and s recycles does not differ much in their influence on the anoxic reactor. Therefore, it matters little which recycle flow brings the nitrate load to the anoxic reactor. As long as the anoxic reactor is closely loaded to its denitrification potential, the same minimum effluent nitrate concentration (Nneaopt) will be obtained at aopt. The aopt–Nneaopt lines therefore give the system denitrification performance when the denitrification potential of the anoxic reactor is fully used (Figure 38(b)), that is, the systems denitrification performance is equal to its potential. A better denitrification performance is not possible – the denitrification is kinetics limited and the biomass (and so also the system) does the best it can (for the given K2 denitrification rate). From Equation (170), the system denitrification performance with increasing influent TKN/COD ratio at a fixed practical operating a recycle ratio (aprac) of 5:1 is also shown in Figure 41 as the aprac and Nneaprac lines. It can be seen that Nneaprac increases linearly with increase in influent TKN/COD ratio. For low influent TKN/COD ratios, aprac is considerably lower than aopt and the system denitrification performance is lower than its denitrification potential. This is evident from Nneaprac being greater than Nneaopt. As the TKN/COD ratio increases, aopt decreases until aopt ¼ aprac ¼ 5.0:1. For the raw wastewater at 14 1C (Figure 41(a)), this happens at an influent TKN/COD ratio of 0.104. This is the influent TKN/COD ratio which balances the MLE system for the selected design conditions, namely, 20 days sludge age, fxm ¼ 0.534 and aprac ¼ 5:1 for the example raw wastewater at 14 1C. For influent TKN/ COD ratios 40.104, the a recycle ratio should be set at aopt, which fully uses the anoxic reactor’s denitrification potential and is now lower than aprac ¼ 5:1. Therefore for aprac set at 5:1, only when the influent TKN/COD ratio is 40.104, is the denitrification potential of the anoxic reactor fully used. This same conclusion can be made from Figure 39(a) at 20 days sludge age, that is, fxm ¼ 0.534 and TKN/COD ratio ¼ 0.104. Therefore for influent TKN/COD ratioso0.104, while aprac oaopt, the system denitrification performance is lower than its denitrification potential because not all the denitrification potential of the anoxic reactor is used. Once the TKN/COD ratio increases above that value which balances the MLE system, aoptoaprac and a should be set at aopt to achieve the lowest effluent nitrate concentration (Nneaopt). For these influent TKN/COD ratios, the denitrification potential of the anoxic reactor is fully used and the system denitrification performance is defined by the aopt–Nneaopt lines. Figure 41 is useful because it combines the system denitrification performance (aprac–Nneaprac lines) and the denitrification potential (aopt–Nneaopt lines) in the same diagram as influent TKN/COD ratio increases for a particular wastewater and system design (Rs ¼ 20 days and fxm ¼ 0.534). The intersection point of the straight Nneaprac line and the curved Nneaopt line (i.e., at aopt ¼ aprac ¼ 5:1) gives the influent TKN/ COD ratio for the balanced MLE system for the selected aprac ¼ 5:1. From Figure 41(a), for the raw wastewater at 14 1C, the MLE system (at 20 days sludge age and fxm ¼ 0.534) with a
Biological Nutrient Removal
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Effluent nitrate (mgN I−1)
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Example Raw WW
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MLE system (settled, 14 °C) a Recycle ratio and effluent nitrate
Effluent nitrate (mgN I−1)
MLE system (raw, 14 °C) a Recycle ratio and effluent nitrate
Effluent nitrate (mgN I−1)
492
0.06 (d)
0 0.10 0.12 0.08 0.14 Influent TKN/COD ratio
0.16
Figure 41 Optimum (aopt) and practical upper limit (aprac ¼ 5:1) a recycle ratios (bold lines) and effluent nitrate concentration at aopt (Nneaopt, bold line) and aprac (Nneaprac, dashed line) vs. influent TKN/COD ratio at underflow (s) recycle ratio of 1:1 for the example raw (a,c) and settled (b,d) wastewaters at 14 1C (a,b) and 22 1C (c,d). The optimum a recycle ratio (aopt) values at underflow recycle ratios of 0.5:1 and 2:1 are also shown (thin lines).
recycle ratio 45:1 can maintain effluent nitrate concentrations below 8.1 (total N 12.4) mgN l1 for influent TKN/COD ratios below 0.104 (78.0 mgN l1). With settled wastewater at 14 1C (Figure 41(b)), the MLE system with a 45:1 can maintain effluent nitrate concentrations below 11.3 (total N 14.9) mgN l1 for influent TKN/COD ratios up to 0.132 (59.4 mgN l1). Similarly, from Figures 41(c) and (d), with raw and settled wastewater at 22 1C, the MLE system with a 45:1 can maintain effluent nitrate concentrations below 6.0 and 8.1 mgN l1 (total N 9.9 and 11.1 mgN l1) for influent TKN/COD ratios up to 0.119 (89.3 mgN l1) and 0.148 (66.6 mgN l1). These results show that the MLE system treating settled wastewater delivers lower Nne (by 2–3 mgN l1) than when treating raw wastewater and at influent TKN/ COD ratios significantly higher. However, it should be noted that (1) the influent TKN concentrations (given above) for the raw wastewater are considerably higher than those for the settled wastewater and (2) a settled wastewater with a TKN/ COD ratio of 0.119 (14 1C) or 0.148 (22 1C) would be produced from a raw wastewater with considerably lower influent TKN/COD ratio than 0.104 (14 1C) and 0.132 (22 1C).
4.14.26.3.4 MLE sensivity diagram In Figure 41, the system denitrification performance at a selected aprac ¼ 5 is combined with the system denitrification potential at a ¼ aopt for varying influent TKN/COD ratio and a single influent RBSO fraction value. This influent TKN/COD ratio sensitivity diagram can be extended by adding the Nneaopt lines for other influent RBSO fractions. A sensitivity analysis of the system at the design stage is useful for evaluating the denitrification performance under varying influent TKN/COD ratio and RBSO fractions. These two wastewater characteristics can vary considerably during the life of the plant and have a major impact on the N removal performance of the system. The denitrification potential and system performance are combined for varying influent TKN/COD ratio and RBSO fraction in Figure 42. For the fixed system design parameters (i.e., Rs ¼ 20 days, fxdm ¼ fxm ¼ 0.534, s ¼ 1.0), the curved (bold) lines give Nneaopt when the anoxic reactor is loaded to its denitrification potential, that is, Nne for a ¼ aopt for varying TKN/COD ratio from 0.06 to 0.16 and RBSO fractions from 0.10 to 0.35 for the example raw and settled wastewaters at
Biological Nutrient Removal MLE system (settled, 14 °C) Effluent nitrate vs. TKN/COD ratio
MLE system (raw, 14 °C) Effluent nitrate vs. TKN/COD ratio 40
20 0.5 1
30 RBCOD fraction 20
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MLE system (settled, 22 °C) Effluent nitrate vs. TKN/COD ratio
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0.06 (d)
0.08
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Figure 42 Effluent nitrate concentration vs. influent TKN/COD ratio for influent readily biodegradable (RBSO) fractions (fSb0 s) of 0.10, 0.15, 0.20, 0.25, 0.30, and 0.35 and mixed liquor a recycle ratio from 0 to 10 for the example raw (a,c) and settled (b,d) wastewaters at 14 1C (a,b) and 22 1C (c,d).
14 1C (Figures 42(a) and 42(b)) and 22 1C (Figures 42(c) and 42(d)). The same Nneaopt lines are given in Figures 41(a) and 41(c) for the raw wastewater RBSO fraction (fSb0 s) ¼ 0.25. These Nneaopt lines are calculated from Equations (169) and (170). The straight lines in Figure 42 give Nneaprac for fixed a recycle ratios at indicated values ranging from 0.0:1 to 10:1. These straight Nneaprac lines give the system performance for some selected a recycle ratio and are calculated with the aid of Equation (170) from the nitrification capacity value at the given TKN/COD ratio, fixed s recycle ratio at 1.0:1, and the selected a recycle ratio. The Nneaprac lines for a ¼ aprac ¼ 5:1 are
the same as the dotted lines in Figure 41. At the intersection points of the straight Nneaprac and curved Nneaopt lines, the system performance equals the denitrification potential and represents balanced MLE designs, that is, aopt ¼ aprac. For example, for the raw wastewater at 14 1C, at a ¼ 5.0:1 and fSb0 s ¼ 0.25, the TKN/COD ratio needs to be 0.104 to give an optimal design, that is, aopt ¼ 5:1 and at this TKN/COD ratio, Nne ¼ 8.1 mgN l1. This is the TKN/COD ratio that balances the MLE system at Rs ¼ 20 days and fxm ¼ 0.534 (see Figure 39). For TKN/COD ratioso0.104, aopt increases above 5:1, but if a is maintained at 5:1 (i.e., a ¼ aprac ¼ 5:1), then Nne
494
Biological Nutrient Removal
versus TKN/COD ratio is given by the a ¼ 5:1 straight Nneaprac line. For TKN/COD ratios 40.104, aopt decreases below 5:1, and Nne versus TKN/COD ratio is given by the curved Nneaopt (bold) line. The aopt value at a particular TKN/COD ratio is given by the a recycle ratio value of the intersection point between the vertical influent TKN/COD ratio line and the curved Nneaopt line, for example, for the example raw wastewater (fSb0 s ¼ 0.25) at 14 1C (Figure 42(a)) at a TKN/COD ratio of 0.12, aopt ¼ 2:1, and Nne is 16.0 mgN l1. The usefulness of Figure 42 is that it gives a performance evaluation of an MLE system at a specified sludge age and anoxic mass fraction for varying influent TKN/COD ratio and RBSO fraction taking due account of an upper a recycle ratio limit of aprac. For the example raw wastewater at 22 1C with a RBSO fraction (fSb0 s) of 0.10 (Figure 42(c)), the influent TKN/ COD ratio needs to be greater than 0.113 for a to beo6.0:1. If a is fixed at aprac ¼ 6.0:1 and the TKN/COD is o0.113, then the anoxic reactor is underloaded with nitrate and the denitrification potential is not achieved. The system performance for influent TKN/CODo0.113 is given by the straight Nne line for a ¼ 6:1. At influent TKN/COD ¼ 0.113, the straight Nne line for a ¼ 6:1 cuts the curved Nneaopt line, a ¼ aopt ¼ 6:1 and the system performance equals the denitrification potential. If a is maintained at 6:1 for TKN/COD 40.113, then the anoxic reactor is overloaded with nitrate and optimal denitrification is not achieved due to the unnecessarily high DO load on the anoxic reactor (similar to that shown in Figure 37(b) for a 46.7). The a recycle ratio therefore should be reduced to aopt for influent TKN/COD ratios 40.113, where aopt is given by the a value along the curved Nne line, which represents system performance equal to denitrification potential. For example, if the TKN/COD ratio ¼ 0.120, a ¼ aopt ¼ 4:1 and this a recycle ratio loads the anoxic reactor to its denitrification potential giving Nne of 12.0 mgN l1. Therefore, for TKN/COD ratio 40.113, the system performance and Nne is given by the curved Nne line provided the a recycle ratio is set to aopt, which is given by a recycle ratio line which passes through the intersection point of the vertical TKN/COD ratio line and the curved Nne line. From the above, it can be seen that only on the curved Nne line for the particular RBSO fraction is the system performance equal to the denitrification potential; also the aopt that produces this is given by the a recycle ratio line that passes through the intersection point of the vertical TKN/COD ratio line and the curved Nne line. This curved Nne line (for which a ¼ aopt) marks the boundary between underloaded and overloaded conditions in the anoxic reactor. In the domain above the curved Nne line, the anoxic reactor is underloaded (left of aopt in Figures 37 and 38) and the system performance (Nne) for a particular TKN/COD ratio is given by the intersection point of the vertical TKN/COD ratio line and the straight a recycle ratio line. In the domain below, curved Nne line, the anoxic reactor is overloaded (right of aopt in Figures 37 and 38). The Nne values obtained from this domain are not valid, but if the a recycle ratio is reduced to aopt (i.e., the a value of the intersection point of the vertical TKN/ COD ratio line and the curved Nne line), then the Nne value again is valid. Valid Nne system performance values are therefore given in Figure 42 only on or above the curved Nne boundary line.
From Figure 42, it can be seen that for MLE system at the design Rs ¼ 20 days and fxm ¼ 0.534 and a recycle ratio limited at say 5.0:1 for economical reasons, then the system is best suited to treating high TKN/COD ratios, depending on the RBSO fraction: 40.091 for fSb0 s ¼ 0.10 and 40.117 for fSb0 s ¼ 0.35. This is because with only a primary anoxic reactor, the MLE system cannot produce a low effluent nitrate concentration (o4–6 mgN l1) at a recycle ratio limited at 5.0:1. If obtaining low effluent nitrate concentrations is not required at low TKN/COD ratios, then a balanced MLE design can be selected by reducing the sludge age as demonstrated in Figures 39 and 40. If obtaining low effluent nitrate concentrations is important at low (o0.10) TKN/COD ratios, then this can be achieved at high a recyle ratios (aopt4aprac) in MLE systems or at low a recycle ratios by including a secondary anoxic reactor. Incorporation of a secondary anoxic reactor (and a re-aeration reactor for practical reasons – see Section 4.14.24.5) produces the four-stage Bardenpho system (Figure 34(c)). However, because the K3 denitrification rate is so low and needs to be reduced by at least 20% to account for the ammonia released during endogenous denitrification (which is re-nitrified in the re-aeration reactor), the net additional nitrate removal achieved in a secondary anoxic reactor is very low, too low for secondary anoxic reactors to be included in N removal systems, unless the influent TKN/COD ratio is unusually low.
4.14.27 System Volume and Oxygen Demand 4.14.27.1 System Volume Having determined the subdivision of the sludge mass into anoxic and aerobic fractions to achieve the required N removal, the actual sludge mass in the system needs to be calculated to determine the volumes of the different reactors. The mass of sludge, total (MLSS) or volatile (MLVSS), in the system for selected sludge age and wastewater characteristics for N removal system is the same as for fully aerobic (COD removal) systems. The equations given in Section 4.14.9 therefore apply to N removal systems also. For the example raw and settled wastewaters, the design parameters for the MLE system are listed in Table 16. The MLSS mass values in the system at 20 days sludge age and 14 1C are 68168 and 26 422 kgTSS, respectively. Selecting an MLSS concentration of 4500 mg l1 (4 kg m3) (see Section 4.14.11) means that the volume of the system treating raw wastewater is 15148 m3 and that treating settled wastewater is 5871 m3. Because the sludge mass in the N removal systems usually is uniformly distributed in the system, that is, each reactor of the system has the same MLSS concentration, the volume fraction of each reactor is equal to its sludge mass fraction. For the example raw and settled wastewaters at 14 1C, the volume of the anoxic reactors are 0.534 15148 ¼ 8089 m3 and 0.534 5871 ¼ 3135 m3, respectively. The nominal and actual hydraulic retention times of the anoxic and aerobic reactors are calculated from the reactor volumes divided by the nominal (influent) and total flows passing through them (Equation (59) and Table 16). Note that the reactor nominal retention time is a consequence of the mass of sludge generated from the influent COD flux, the selected MLSS concentration, and the sludge mass fraction – the retention time per se has no significance in
Biological Nutrient Removal Table 16 Design details of MLE systems treating the example raw and settled wastewaters at 14 1C at 20 days sludge and 0.534 unaerated sludge mass fraction Parameter
Raw
Settled
Influent TKN/COD ratio Influent RBCOD fraction (fSb0 s) Unaerated mass fraction(fxm) Anoxic mass fraction (fx1) Minimum anoxic fraction a Recycle ratio (apracoaopt) Sludge age (days) Effluent nitrate (Nne, mgN l1) Effluent TKN (Nte, mgN l1) Effluent total N (Nne þ Nte) System vol at 4.5 gTSS l1 (m3) Anoxic volume (m3) System ret time – nom (h) Aerobic ret time – nom (h) Aerobic ret time – actual (h) Anoxic ret time – nom (h) Anoxic ret time – actual (h) Carb. O2 demand (FOc, kgO d1) Nit O2 demand (FOn, kgO d1) O2 recovered (FOd, kgO d1) Tot. O2 demand (FOtd, kgO d1) %N removal Mass TSS wasted (FXt, kg d1) Active fraction wrt TSS (fatOHO)
0.08 0.25 0.534 0.534 0.07 5:1 20 5.6 3.8 9.4 15148 8089 24.2 11.2 1.6 12.9 1.85 6679 2685 1440 7924 84.4 3408 0.23
0.113 0.385 0.534 0.534 0.105 5:1 20 5.7 3.8 9.5 5871 3135 9.4 4.4 0.63 5 0.72 4311 2719 1458 5572 81.4 1321 0.383
kinetics of and design for nitrification and denitrification (see Section 4.14.9.3).
4.14.27.2 Daily Average Total Oxygen Demand The total oxygen demand in a nitrogen removal system is the sum of that required for organic material (COD) degradation and nitrification, less than recovered by denitrification. The daily average oxygen demand for (1) organic material removal (FOc) is given by Equations (111) and (2) nitrification is given by Equation (155). These oxygen demands in the MLE system at 20 days sludge age for the example raw and settled wastewaters at 14 and 22 1C are 9364 and 7030 kgO d1 (Table 16). The oxygen recovered by denitrification (FOd) is given by 2.86 times the nitrate flux denitrified (Section 4.14.24.2) where nitrate flux denitrified is the product of the daily average influent flow Qi and the nitrate concentration denitrified. The nitrate concentration denitrified is given by the difference in the nitrification capacity Nc and the effluent nitrate concentration. Hence,
FOd ¼ 2:86ðNc Nne ÞQi
ðmgO d1 Þ
495
demand by incorporating ND is only 20% of that required for COD removal only, and (4) the effect of temperature on the total oxygen demand is marginal – less than 3% (see also Figure 32). For the settled wastewater, Table 16 shows that (1) the nitrification oxygen demand is about 63% of that required for COD removal; (2) about 54% of the nitrification oxygen demand can be recovered by denitrification; (3) the additional oxygen demand by incorporating nitrification and denitrification is about 30% of that required for COD removal only, and (4) the effect of temperature on the total oxygen demand is marginal – less than 3% more at the lower temperature. Comparing the total oxygen demand (FOtd) for the raw and settled wastewaters, the total oxygen demand for the latter is about 30% less than that of the former. This saving is possible because primary sedimentation removes 35–45% of the raw wastewater COD. Furthermore, for the settled wastewater, the nitrification oxygen demand is a greater proportion of the total; also, less of the nitrification oxygen demand can be recovered by denitrification compared to the raw wastewater. These effects are due to the higher TKN/COD ratio of the settled wastewater. Knowing the average daily total oxygen demand, (FOtd) the peak total oxygen demand can be roughly estimated by means of a simple design rule (Musvoto et al., 2002). From a large number of simulations with AS model no. 1 (ASM1), it was found that, provided the factor of safety on nitrification (Sf) is greater than 1.25–1.35, the relative amplitude (i.e., (peak average)/average) of the total oxygen demand variation is a fraction 0.33 of the relative amplitude of the TOD of the influent COD and TKN load (i.e., Qi(Sti þ 4.57Nti)). For example, with the raw wastewater case, if the peak influent TOD flux is obtained at a time of day when the influent flow rate, COD and TKN concentrations are 25 M l d1, 1250 mgCOD l1 and 90 mgN l1, respectively – that is, 25(1250 þ 4.57 90) ¼ 41 532 kgTOD d1, and the average influent TOD flux is 15(750 þ 4.57 60) ¼ 15 363 kgTOD d1, the amplitude of the total influent TOD flux is (41 532 15 363)/15 363 ¼1.70; hence, the amplitude of the total oxygen demand is approximately 0.33 1.70 ¼ 0.56; from Table 16 the average daily total oxygen demand (FOtd) is 7924 kgO d1 and hence the peak oxygen demand is (1 þ0.56) 7924 ¼12 378 kgO d1. As with all simplified design rules, the above rule should be used with discretion and caution, and where possible, the peak total oxygen demand is best estimated by means of the AS simulations models.
4.14.28 Biological Excess Phosphorus Removal ð172Þ
From the denitrification performance of the MLE system in Table 16, the oxygen recovered by denitrification for the example raw and settled wastewaters at 14 1C are 1440 and 1458 kgO d1. For the raw wastewater, Table 16 shows that (1) the nitrification oxygen demand (FOn) is about 40% that required for COD removal (FOc), (2) about 55% of FOn can be recovered by incorporating denitrification, (3) the additional oxygen
4.14.28.1 Introduction Phosphorus is the key element in aquatic environments that limits the growth of aquatic plants and algae controls eutrophication. Unlike nitrogen that can be fixed from the atmosphere which contains about 80% nitrogen gas, phosphorus can only come from upstream of aquatic systems (neglecting atmospheric deposition). Diffuse sources of phosphorus, for example, from agricultural fields, are best controlled by proper fertilization plans, while point sources of
Biological Nutrient Removal
phosphorus, for example, from WWTPs, can be removed by chemical or biological processes. Considering the benefit to aquatic environments, strict regulations are being applied for phosphorus removal from wastewaters. Considering the potential benefits of removing phosphorus biologically rather than chemically, along with organic matter and nitrogen from wastewater, BEPR has stimulated much interest in the study of the biochemical mechanisms, the microbiology of the systems, the process engineering and optimization of plants, and in mathematical modeling. Reviews of the development of BEPR have been regularly published over the years (Marais et al., 1983; Arvin, 1985; Wentzel et al., 1991; Jenkins and Tandoi, 1991; van Loosdrecht et al., 1997; Mino et al., 1998; Blackall et al., 2002; Seviour et al., 2003; Oehmen et al., 2007). This section briefly reviews the mechanisms of BEPR, outlines the practical systems to achieve it, summarizes some of the experimental research that led to the development of BEPR models (both steady state and dynamic kinetic), discusses the impact of anoxic zones for denitrification on BEPR, and sets out guidelines for design of NDBEPR systems. In order not to unduly complicate this, the concepts are presented for strictly aerobic phosphorus accumulating organisms (aerobic PAOs) which can use only oxygen as the electron acceptor for energy production. Considering that some denitrifying PAOs (DPAOs) exist and may have a significant impact on the performance of the process, their influence is discussed where appropriate, but is not included in the models described.
4.14.28.2 Principles of BEPR BEPR is the biological uptake and removal of P by AS systems in excess of the amount that is removed by normal completely aerobic AS systems. This is in excess of the normal P requirements for growth of AS. In the completely aerobic AS system, the amount of P typically incorporated in the sludge mass is about 0.02 mgP/mgVSS (0.015 mgP/mgTSS). By the daily wastage of surplus sludge phosphorus is thus effectively removed. This can give a P removal of about 15–25% of the P in many municipal wastewaters. In an BEPR AS system, the amount of P incorporated in the sludge mass is increased from the normal value of 0.02 mgP/mgVSS to values around 0.06–0.15 mgP/ mgVSS (0.05–0.10 mgP/mgTSS). This is achieved by system design or operational modifications that stimulate, in addition to the OHOs present in AS, the growth of organisms that can take up large quantities of P and store them internally in long chains called polyphosphates (polyPs); generically, these organisms are called phosphate accumulating organisms (PAOs). PAOs can incorporate up to 0.38 mgP/mgVSS (0.17 mgP/ mgTSS). In the biological P removal system both the OHOs, which do not remove P in excess, and the PAOs coexist. The larger the proportion of PAOs that can be stimulated to grow in the system, the greater the P content of the AS and, accordingly, the larger the amount of P that can be removed from the influent. Thus, the challenge in design is to increase the amount of the PAOs relative to the OHOs present in the AS as this will increase the capacity for P-accumulation and thereby high phosphorus removal efficiency. The relative proportion of the two organism groups depends, to a large degree, on the fraction
15 Example settled WW % P of VSS (mgP/mgVSS as %)
496
Settled WW 10 Example Raw WW Raw WW 5
P removal =
%P × VSS mass Sludge age × Q i
0 0
10
20
30
40
% Bio COD obtained by PAOs Figure 43 Percentage P (mgP/mgVSS 100) in VSS mass vs. the proportion of biodegradable COD mass (as %) obtained by PAOs.
of the influent wastewater biodegradable COD that each organism group obtains. The greater the fraction of PAOs in the mixed liquor, the greater the %P content of the AS and the greater the BEPR (Figure 43). Design and operational procedures are oriented toward maximizing the growth of PAOs. In an appropriately designed BEPR system, the PAOs can make up about 40% of the active organisms present (or 15% of VSS; 11% of TSS), and this system can usually remove about 10–12 mgP per 500 mg influent COD l1. From the first publications reporting enhanced P removal in some AS systems, there has been some controversy as to whether the mechanism is a precipitation of inorganic compounds, albeit perhaps biologically mediated, or biological through formation and accumulation of P compounds in the organisms. The objective here is not to discuss the evidence that supports the biological nature of enhanced P removal, but to briefly describe the theory of biological P removal and to demonstrate how this theory can be used as an aid for the design of biological P removal AS systems. This does not imply that precipitation of P due to chemical changes resulting from biological action (e.g. alkalinity and pH) does not take place. Although inorganic precipitation of P can certainly take place, it would appear that in the treatment of municipal wastewaters by an appropriately designed AS system, within the normal ranges of pH, alkalinity and calcium concentrations in the influent, enhanced P removal is principally mediated by a biological mechanism (Maurer et al., 1999; de Haas et al., 2000). These mechanisms are described below.
4.14.28.3 Mechanism of BEPR 4.14.28.3.1 Background Historically, several research groups have made a number of important contributions toward elucidating the mechanisms
Biological Nutrient Removal
4.14.28.3.2 Biological P removal microorganisms The basic requirement for BEPR is the presence in the AS system of microorganisms which can accumulate P in excess of normal metabolic requirements, in the form of polyP stored in granules called volutins. In the BEPR models, all organisms in the AS system accumulating polyP in this fashion and exhibiting the classical observed BEPR behavior – anaerobic P release, aerobic P uptake, and associated processes – are lumped together and represented by the generic PAO group. PolyPs can be accumulated by a wide range of bacteria. In general, they are accumulated as a phosphate reserve in relatively low amounts. Only very few types of bacteria seem to be able to harvest the energy that is stored in polyPs to take up VFAs and store them as PHAs under anaerobic conditions (in the absence of an external electron acceptor such as oxygen or nitrate). In the original research on BEPR microbiology conducted with cultivation studies, it was incorrectly considered that PAOs were of the genus Acinetobacter (Fuhs and Chen, 1975; Buchan, 1983; Wentzel et al., 1986), Microlunatus phosphovorus (Nakamura et al., 1995), Lampropedia (Stante et al., 1997), and Tetrasphaera (Maszenan et al., 2000). More recently, cultureindependent methods have shown that Accumulibacter phosphatis, a member of the genus Rhodocyclus (a beta proteobacterium), is a PAO which can be grown in enriched cultures (at up to 90% purity, as shown by fluorescence in situ hybridization (FISH) molecular probes) but not yet in axenic cultures (Wagner et al., 1994; Hesselmann et al., 1999; Crocetti et al., 2000; Martin et al., 2006; Meyer et al., 2006; Oehmen et al., 2007). From a modeling and design perspective, however, the identification of the exact organisms responsible for BEPR is of minor importance, although this may provide information that can be used to refine the models and design procedures; these are not based on the behavior of specific organisms, but rather on the observed behavior of groups of organisms identified by their function, in this case the PAOs.
4.14.28.3.3 Prerequisites To achieve BEPR in AS systems, the growth of organisms that accumulate polyP (PAOs) has to be stimulated. To accomplish this, two conditions are essential: (1) an anaerobic and aerobic (or anoxic) sequence of reactors/conditions and (2) the addition or formation of VFAs in the anaerobic reactor/period.
Glycogen
Concentration
of BEPR, including Shapiro et al. (1967), Fuhs and Chen (1975), Nicholls and Osborn (1979), Rensink et al. (1981), Marais et al. (1983), Lotter (1985), Comeau et al. (1986), Wentzel et al. (1986, 1991), Mino et al. (1987, 1994, 1998), Kuba et al (1993), Smolders et al. (1994a, 1994b, 1995), van Loosdrecht et al. (1997), Maurer et al. (1997), Seviour et al. (2003), Martin et al. (2006), and Oehmen et al. (2007). In this section, an explanation of the basic concepts underlying the more sophisticated mechanistic models for the biological P removal phenomenon is presented. For detailed description of the mechanisms, the reader is referred to the references mentioned earlier in this paragraph.
497
Poly P
PHA VFA Anaerobic
PO4 Aerobic
Figure 44 Schematic diagram showing the changes as a function of time in concentrations of volatile fatty acids (VFAs), P (PO4), polyphosphates (polyPs), polyhydroxyalkanoate (PHA), and glycogen through the anaerobic–aerobic sequence of reactors in a BEPR system.
4.14.28.3.4 Observations With the prerequisites for BEPR present, the following observations have been made at full, pilot, and laboratory scale (Figure 44). Under anaerobic conditions, bulk solution VFAs and intracellular polyP and glycogen decrease and soluble phosphate, Mg2þ, Kþ, and intracellular poly-b-hydroxyalcanoates (PHAs) increase (Randall et al., 1970; Rensink et al., 1981; Hart and Melmed, 1982; Fukase et al., 1982; Watanabe et al., 1984; Arvin, 1985; Hascoe¨t et al., 1985a; Wentzel et al., 1985; Comeau et al., 1986, 1987; Murphy and Lo¨tter, 1986; Gerber et al., 1987; Wentzel et al., 1988; Satoh et al., 1992; Smolders et al., 1994a; Maurer et al., 1997). Under aerobic conditions; intracellular polyP and glycogen increase; soluble phosphate, Mg2þ, Kþ, and intracellular PHA decrease (Fukase et al., 1982; Arvin, 1985; Hascoe¨t et al., 1985a; Comeau et al., 1986; Murphy and Lo¨tter, 1986; Gerber et al., 1987; Wentzel et al., 1988; Satoh et al., 1992; Smolders et al., 1994b; Maurer et al., 1997).
4.14.28.3.5 Biological P removal mechanism In describing the mechanisms of BEPR, a clear distinction is made between the PAOs and OHOs. In the anaerobic/aerobic sequence of reactors, it is considered that VFAs are present in the influent waste stream entering the anaerobic reactor or produced in the anaerobic reactor by fermenting organisms (accepted to be the OHOs in models). In the anaerobic reactor (zero nitrate and oxygen in or entering reactor), the OHOs cannot utilize the VFAs due to the absence of an external electron acceptor, oxygen or nitrate. The PAOs, however, can take up the VFAs from the bulk liquid and store them internally by linking the VFAs together to form complex long-chain carbon molecules of poly-b-hydroxyalkanoates (PHAs). The two common PHAs are poly-b-hydroxybutyrate (PHB: four-carbon compound synthesized from two acetate molecules) and polyhydroxyvalerate (PHV: five-carbon compound from one acetate and one propionate molecules) (Figure 45(a)). Forming PHAs from the VFAs requires energy for three functions: active transport of VFAs across the cell membrane, energization of VFAs into coenzyme A compounds (e.g.,
498
Biological Nutrient Removal Liquid Cell
PHA e−
Glycogen
VFA-CoA Energy
VFA
(PO4)n
VFA
(PO4)n−1 Pi
PAO
Pi
(a)
Liquid CO2
CO2
Catabolism PHA
Cell
e− Glycogen ETC
Anabolism
Energy H2O
New cells
O2 (PO4)n Pi
(b)
(PO4)n −1 PAO
Pi
Figure 45 (a) Simplified biochemical model for PAOs under anaerobic conditions. Anaerobic uptake of volatile fatty acids (VFAs), originating from the influent or from fermentation in the anaerobic reactor, and storage of polyhydroxyalkanoates (PHAs) by the PAOs with associated P release. (b) Simplified biochemical model for PAOs under aerobic conditions. Aerobic utilization of PHAs and growth of PAOs, with P uptake by existing and new PAOs.
acetyl-CoA) and reducing power (NADH) for PHA formation. PolyP degradation is associated with the formation of ADP from AMP, with the phosphokinase enzyme 2 ADP are converted to adenosine triphosphate (ATP) and adenosine monophosphate (AMP) (van Groenestijn et al., 1987). When ATP is used, orthophosphates are released and accumulate in the cell interior together with the counterions of polyP (potassium and magnesium). The efflux of these compounds might be related to building a proton motive force, which either can help in the uptake of acetate or in the generation of a small amount of extra ATP. It is observed (Smolders et al., 1994a) that the energy requirements for acetate uptake increase with increasing pH. This can be associated with the fact that the energy needed for acetate transport increases with pH. ATP is used, notably, for the energization of acetate and propionate into acetyl-CoA and propionyl-CoA. Glycogen degradation also results in ATP formation, NADH production, and intermediates that are
transformed into acetyl-CoA (or propionyl-CoA). Finally, acetyl-CoA and propionyl-CoA are stored as PHA (Comeau et al., 1986; Wentzel et al., 1986; Mino et al., 1998; Smolders et al., 1994b; Martin et al., 2006; Oehmen et al., 2007; Saunders et al., 2007). Thus, the PAOs in the anaerobic reactor have taken up for their exclusive use the VFAs under anaerobic conditions where the OHOs are unable to use these organics. To accomplish this, some of the stored polyP has been consumed and P released to the bulk solution. To stabilize the negative charges on the polyP, the cations Mg2þ, Kþ, and sometimes Ca2þare complexed, which add to the inorganic settleable solids (TSS) in the system (Ekama and Wentzel, 2004). When polyPs are consumed and P is released, mainly Mg2þ and Kþ cations are released in the approximate molar ratio P:Mg2þ:Kþ of 1:0.33:0.33 (Comeau et al., 1987; Brdjanovic et al., 1996; Pattarkine and Randall, 1999). In the subsequent aerobic reactor (presence of DO). In the presence of dissolved oxygen (or of nitrate under anoxic conditions) as an external electron acceptor, the PAOs utilize the stored PHA as a carbon and energy source for energy generation and growth of new cells as well as for regenerating the glycogen consumed in the anaerobic period. The stored PHA is also used as an energy source to take up P from the bulk solution to regenerate the polyP used in the anaerobic reactor, and to synthesize polyP in the new cells that are generated – P uptake (Figure 45(b)). The uptake of P to synthesize polyP in the new cells generated means that more P is taken up than is released in the anaerobic reactor, giving a net removal of P from the liquid phase in the AS system. Accompanying the P uptake, the cations Mg2þ and Kþ also are taken as countercharge for the negatively charged polyP polymer, in the approximate molar ratio P:Mg2þ:Kþ of 1:0.33:0.33. The PAOs, with stored polyP, are removed from the aerobic reactor of the system (where the internally stored polyP concentration in the PAOs is the highest in the system) via the waste sludge stream (wastage from the underflow recycle stream is possible, but not desirable for hydraulic control of sludge age; see Section 4.14.14). At steady state the mass of PAOs wasted per day (with stored polyP) equals the mass of new PAOs generated per day (with stored polyP). Thus, for a fixed sludge age, loading, and system operation, the mass of PAOs in the biological reactors remains constant, so that in the AS system at steady state there is neither a buildup nor a loss of PAOs, and the P/VSS ratio stays approximately constant. The mass of new PAOs formed depends on the mass of stored substrate (PHA) available to the PAOs. Accordingly, the enhanced P removal attained will depend on the mass of PHA stored in the anaerobic reactor.
4.14.28.3.6 Fermentable COD and slowly biodegradable COD As indicated above, under anaerobic conditions, PAOs can take up and store VFAs. However, some wastewaters contained very little VFAs, yet exhibited significant BEPR. This was ascribed to the influent RBOs, (Sbsi) which comprises both VFAs (Sbsai) and fermentable RBO (FBSO, Sbsfi) (Siebritz et al., 1983; Wentzel et al., 1985, 1990; Nicholls et al., 1985; Pitman et al., 1988; Randall et al., 1994). This influent FBSO is
Biological Nutrient Removal Liquid Cell F-RBCOD
F-RBCOD
Energy
VFA OHO
VFA
VFA
PAO
Figure 46 Simplified biochemical model for fermentation of RBSO to VFA by OHOs under anaerobic conditions – VFAs released by OHOs are taken up by PAOs.
fermented to VFAs by the OHOs in the anaerobic reactor, the VFAs becoming available for uptake and storage by the PAOs because the OHOs cannot utilize them due to the absence of an electron acceptor (NO3 or O) (Figure 46). Slowly biodegradable organics (SBO, XS), even though these can be hydrolyzed into RBO under anaerobic conditions, has been shown not to be linked to anaerobic phosphate release. This aspect is of crucial importance as it will influence both the design and operation of BNR systems, such as sizing and determining the number of anaerobic reactors, inclusion of primary sedimentation and maximum BEPR achievable. For the purpose of the BEPR models, the experimental evidence linking BEPR to the RBO is accepted, but a conversion of SBO to RBO is considered to be small enough to be negligible. Accordingly, where VFA production does occur, this will essentially be from the RBO. One exception to this consideration is when primary sludge is fermented in a separate fermentation reactor upstream of the anaerobic reactor – in these dedicated fermenters, some hydrolysis of SBO to RBO and VFAs takes place to augment the influent VFA and RBO concentrations (Lilley et al., 1992).
4.14.28.3.7 Functions of the anaerobic zone From the description of the mechanisms above, with normal domestic wastewater as influent, the anaerobic zone/reactor serves two functions: (1) it stimulates conversion of fermentable organics to VFAs by OHOs, that is, facultative acidogenic fermentation and (2) because it is not possible for the OHOs to metabolize the VFAs (no external electron acceptor), the PAOs take up the released VFAs and store them as PHA. Thereafter, the PAOs do not have to compete for substrate when an external electron acceptor becomes available in the aerobic (or anoxic) zone. Of the above two processes, the former is the slower and determines the size of the anaerobic reactor (Wentzel et al., 1985, 1990). Should primary sludge fermentation be implemented at the treatment plant, the first process would not be needed as much and the size of the anaerobic reactor could be decreased.
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4.14.28.3.8 Influence of recycling oxygen and nitrate to the anaerobic reactor Numerous investigators (e.g., Barnard, 1976; Venter et al., 1978; Siebritz et al., 1980; Hascoe¨t and Florentz, 1985b) noted that the recycling of oxygen and/or nitrate to the anaerobic reactor causes a corresponding decrease in BEPR. In terms of the mechanisms described above, if oxygen and/or nitrate is recycled to the anaerobic reactor, the OHOs are able to utilize the fermentable COD for energy and growth using the oxygen or nitrate as external electron acceptor. For every 1 mgO recycled to the anaerobic reactor 3 mgCOD of fermentable RBO are consumed and for every 1 mgN of nitrate recycled 8.6 mgCOD of fermentable RBO are consumed by the OHOs. The ratio of 3 mgCOD/ mgO consumed comes from the catabolic oxygen requirement in organics utilization (i.e., 1/(1 fcvYH)E3) (Equation (46)). Similarly, considering that 1 mgNO3-N is equivalent to 2.86 mgO (Section 4.14.24.2), a ratio of 2.86/ (1 fcvYH)E8.6 mgCOD consumed by mgNO3-N reduced is obtained. The fermentable RBOs metabolized by the OHOs are not released to the bulk liquid as VFAs. Therefore, the amount of VFAs generated and released to the bulk liquid is reduced by the amount of RBO consumed by the OHOs. Consequently, the mass of VFAs available to the PAOs for storage is reduced, and correspondingly so is the P release, P uptake, and the net P removal. Should the influent RBO already consist of VFAs and oxygen and/or nitrate be recycled, the PAOs and OHOs will compete for the VFAs, the PAOs to take up the VFAs, and the OHOs to metabolize it. Accordingly, even in this situation recycling of oxygen and/or nitrate will reduce the BEPR. Thus, preventing the recycling of oxygen and nitrate to the anaerobic reactor is one of the primary considerations in the design and operation strategy for BEPR systems.
4.14.28.3.9 Denitrification by PAOs The extent of denitrification with associated anoxic P uptake by the PAOs appears to be highly variable (Ekama and Wentzel, 1999b), from near-zero anoxic P uptake (e.g., Wentzel et al., 1989a, Clayton et al., 1989, 1991) to anoxic P uptake dominant over aerobic P uptake (e.g. Sorm et al., 1996; Hu et al., 2000). Experimental evidence tends to suggest that magnitude of anoxic P uptake is influenced by the anoxic mass fraction and the mass of nitrate loaded on the anoxic reactor relative to its denitrification potential (Hu et al., 2002). For the purpose of design it will be considered that anoxic P uptake is not significant. Anoxic P uptake decreases the magnitude of P removal in the system (Ekama and Wentzel, 1999a, 1999b; Hu et al., 2002), and from a design point of view in which maximizing P removal is a priority, anoxic P uptake should be avoided in the system. Hence, in this chapter, anoxic P uptake will not be considered. It must be emphasized, however, that due to the anaerobic conversion of RBO to VFA which are taken up by PAOs, the kinetics of denitrification in the subsequent anoxic reactor change compared with that in the primary anoxic reactor of an MLE system.
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4.14.29 Principles of Maximizing BEPR The principles of maximizing BEPR can be grouped into seven categories. A number of configurations or systems that are based on these principles are identified by specific names (Figure 47). 1. Oxygen entrainment in the anaerobic reactor should be minimized. For this purpose, mixing vortexes, upstream cascades, and screw pumps or air lift pumps should be avoided.
2. Nitrate (and nitrite) entering in the anaerobic reactor should be minimized. A number of named configurations were developed precisely for this purpose (Section 4.14.34). Based on observations a number of laboratory-, pilot- and full-scale systems (Barnard, 1974, 1975a, 1975b; Nicholls, 1975b), to achieve BEPR in the simplest configuration, Barnard (1976) proposed the Phoredox system (Figure 47(a) also known as the A/O process). This system comprises only an anaerobic and aerobic reactor and is intended not to nitrify to avoid nitrate entering the
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Figure 47 System configurations for BEPR: (a) phoredox or A/O; (b) three-stage Bardenpho or (A2/O); (c) five-stage Bardenpho; (d) University of Cape Town (UCT) or Virginia Initiative Process (VIP); (e) modified UCT; (f) Johannesburg (JHB); (g) biological chemical flexible system (BCFS); and (h) phostrip.
Biological Nutrient Removal
anaerobic reactor. Because nitrification can take place even at short sludge ages, particularly in warm climates, one or more anoxic reactors for denitrification are included in the two-reactor anaerobic–aerobic system to protect the anaerobic reactor from nitrate entering it. The position of the anoxic reactor(s) has led to a number of different configurations: (1) one between the anaerobic and aerobic reactors with the return sludge discharged to the anaerobic reactor (three-stage Bardenpho or A2/O systems, Figure 47(b)), (2) anoxic reactors before and after the aerobic reactor with the return sludge discharged to the anaerobic reactor (five-stage Bardenpho, Figure 47(c)), (3) one or two anoxic reactors between the anaerobic and aerobic reactors with the sludge return discharged to the first or only anoxic reactor (UCT, Siebritz et al., 1980 or VIP, Daigger et al., 1987; Figure 47(d) and modified UCT systems; Figure 47(e)), and (4) an anoxic reactor between the anaerobic and aerobic reactors and another in the sludge return flow (JHB system; Figure 47(f)). 3. VFA uptake by PAOs in the anaerobic reactor should be maximized. Primary sludge fermentation is an efficient way to increase the VFA content of the influent even though it also contributes to an increased loading in organic matter and ammonia to the AS system. Sodium acetate or fermentable industrial wastes can be added directly to the anaerobic reactor or industries that produce fermentable organics (e.g., breweries or food processing factories) should not be penalized for discharging their high RBO containing wastewater to the sewer. The sludge mass fraction of the anaerobic reactor can be increased to favor in situ fermentation of the influent or added fermentable organic matter. 4. Effluent particulate phosphorus should be minimized by removing TSSs efficiently. The particulate phosphorus content can reach as high as 18% gP/gTSS for enriched cultures. With a more typically 5–10% P content for municipal wastewater (Figure 43), every 10 mgTSS l1 in the effluent will contribute 0.5 to 1 mgP l1. Thus, efficient secondary clarification, avoiding floating sludge from denitrification in the settling tank, sand filtration, or even ultrafiltration (in a membrane bioreactor) are means of reducing the effluent TSS concentration. 5. Effluent soluble phosphorus should be minimized. Besides optimizing the BEPR process, chemical coagulants such as iron (e.g., FeCl3), aluminum (e.g., alum), or calcium (e.g., lime) salts can be added in the mainstream for pre-, co-, or post-precipitation (in the primary settling tank, in the AS process, downstream of the secondary settling tank, respectively, de Haas et al., 2001). Extracting the supernatant from the anaerobic tank or taking some sludge from the return AS and coagulating them can also lead to lower effluent soluble phosphorus (Sehayek and Marais, 1981; van Loosdrecht et al., 1998; e.g., BCFS process; Figure 47(g)). Sidestream lime precipitation of phosphate released anaerobically from the return sludge can also be done. More efficient phosphate release can be achieved in this sidestream tank by diverting some influent containing readily biodegradable COD (e.g., PhoStrip process, Figure 47(h)). These systems support the biological
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process by phosphate stripping and potential P recovery in the main line, stabilizing the sludge settling properties and optimizing the control of nitrogen removal. In the BCFS system, a third recycle is added from the aerated reactor to the first anoxic reactor in order to maximize denitrification and to be able to aerate the second anoxic reactor during peak flows or cold temperatures. In this way both ammonium and nitrate can be better controlled to low effluent values (ammonium typical below 0.5 gN l1 and nitrate around 5–8 mg N l1). The recycle flows are controlled by a simple redox electrode-based controller (van Loosdrecht et al., 1998). Compartmentalizing the reactors and low effluent ammonia concentration contributes to a stable low SVI – around 120 ml g1 (Kruit et al., 2002; Tsai et al., 2003). Biological phosphorus removal can be supplemented by addition of precipitants to the anaerobic tank. Since phosphate concentrations are high in this tank, the precipitants are used effectively. Dosing chemicals, however, should be done carefully. Too much precipitation will make the phosphate unavailable for PAOs and deteriorate the BEPR efficiency (de Haas et al., 2001). A complicating factor is that the WWTP will respond rapidly to changes in addition of chemicals whereas the biological phosphorus removal process might have a response time of several days if not weeks. In the BCFS process, a small baffle is placed at the end of a plugflow anaerobic tank. The sludge will locally settle back into the anaerobic tank and a clear supernatant can be withdrawn for phosphate precipitation. The phosphorus can then be recovered (Barat and van Loosdrecht, 2006) or the chemical sludge produced can be prevented from accumulating in the AS which would limit the overall capacity of the plant by reducing the sludge age. Should anaerobic or aerobic digestion be performed with the wasted secondary sludge, essentially all of the polyPs will be hydrolyzed to ortho-P and the phosphate released in solution (Jardin and Po¨pel, 1994; Harding et al., 2009; Mebrahtu et al., 2010). Phosphorus recovery in the form of struvite (MgNH4PO4) or hydroxyapatite (Ca10(PO4)6OH2), which can be used as fertilizers, are also means of reducing the loading of soluble phosphate back to the AS process and, eventually, to the effluent. 6. Phosphorus uptake for cell synthesis should be maximized. Although more limited than the other maximization principles in its potential efficiency, maintaining the sludge age as short as possible will result in an increase in phosphorus removal by sludge production (cell synthesis). Although the endogenous respiration rate of the PAOs is low (0.04 d1), another small benefit of reducing the sludge age is that the PAOs degrade to a lower extent their polyP reserves for cell maintenance. 7. Because anoxic P uptake BEPR reduces the P content of the PAOs (Ekama and Wentzel, 1999a, b; Hu et al., 2002), growth of denitrifying PAOs should be avoided to maximize aerobic P uptake BEPR to maximize PAO P content – up to 0.38 mgP/mgPAOVSS (Wentzel et al., 1989b, 1990). For a review of how these developments took place, the reader is referred to Henze et al. (2008). In order to efficiently construct all the tanks in these complex BNR systems, it is possible
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Figure 48 Zandvliet nitrification–denitrification (ND)BEPR WWTP Cape Town, South Africa. By arranging interconnecting recycle flows between the anaerobic (centre), anoxic (inner ring), and aerobic (outer ring) reactors, the system has the flexibility to be operated as a UCT, threestage Bardenpho or JHB system. Photo: GA Ekama.
to shift the construction from rectangular tanks to one round tank with a sloping outside wall divided in rings for the different aerobic/anoxic/anaerobic zones. In this way, the amount of concrete needed is minimized as all the walls require much less strength (Figure 48).
4.14.30 Model Development for BEPR 4.14.30.1 Early Developments When the first mainstream NDBEPR system was proposed (the five-stage Bardenpho, Figure 47(a); Barnard, 1976), initial conceptualization of the phenomenon extended little beyond recognition of (1) the necessity of an anaerobic/aerobic sequence of reactors and (2) the adverse influence of nitrate recycled to the anaerobic zone. With the inclusion of the secondary anoxic reactor, it was believed that nearly complete denitrification of nitrate would be achieved, thereby discharging very low nitrate concentrations to the anaerobic reactor. Design procedures were based on empirically based estimates for sizing denitrification and anaerobic reactors in terms of nominal hydraulic retention time, and sizing of the anaerobic reactor appeared to be linked to depression of the redox potential below some critical value. No rational method for predicting N and P removal was available and for design, removals were estimated largely from experience gained in operating experimental systems similar to the proposed systems (McLaren and Wood, 1976; Simpkins and McLaren, 1978; Osborn and Nicholls, 1978).
4.14.30.2 RBO and Anaerobic Mass Fraction In seeking an explanation for the different P release and enhanced P removal behavioral patterns in lab-scale modified UCT (Figure 47(e)) and MLE (Figure 34(b)) systems, Siebritz et al. (1980, 1983) applied the concept of RBO developed in denitrification and aerobic studies (Dold et al., 1980; van Haandel et al., 1981) to BEPR systems. They noted that the only evident difference between the modified UCT and MLE
systems lay in the concentration of RBO surrounding the organisms in the anaerobic reactor. (They also observed that the UCT and MLE systems with the same anoxic mass fractions yielded approximately the same effluent nitrate concentrations and the ND kinetic models (such as ASM1, Henze et al., 1987 or UCTOLD, Dold et al., 1991) predicted the NDBEPR system response reasonably well even at full scale (Nicholls, 1982). This implied that the anaerobic reactor did not appear to have a detrimental effect on the denitrification (the questions this raises regarding denitrification in NDBEPR systems are discussed in Section 4.14.34.) In the modified UCT system the RBO concentration in the anaerobic reactor is the maximum possible as no nitrate is recycled to the anaerobic reactor; in contrast, in the MLE system sufficient nitrate is recycled to the anoxic reactor to utilize all the RBO. Therefore, the different behavioral patterns of the systems would be consistently described if it is assumed that the concentration of RBO from the influent in the anaerobic reactor surrounding the organisms is a key parameter determining whether or not P release and BEPR take place. (Later it became clear that the parameter influent RBO concentration in the anaerobic reactor surrounding the organisms represented the influent and produced VFAs taken up by the PAOs in the anaerobic reactor.) The validity of this RBO hypothesis was established by Siebritz et al. (1983) at laboratory scale and Nicholls et al. (1985) at full scale, who found that the magnitude of the P release was proportional to the influent RBO concentration. This opened the way for enquiry into other factors affecting the P release and the BEPR and quantifying BEPR. It was concluded that the BEPR depended on two main parameters, viz. (1) influent RBO concentration and (2) the anaerobic sludge mass fraction. Testing the concepts of the parametric model did, in general, demonstrate the utility of the model. At laboratory scale, the concepts were tested in the modified UCT system at different sludge ages, temperatures, anaerobic mass fractions, and influent COD concentrations in which the RBSO fraction of the influent (unsettled municipal sewage) was augmented by the addition of glucose or acetate. Based on the influent RBO concentration and anaerobic mass fraction parameters, the predicted P removal compared quite consistently with the measured P removal. At full scale, evaluation of the Goudkoppies and Northern Works WWTPs with the parametric model provided a consistent explanation when good or poor P removal was obtained (Nicholls et al., 1985; 1986; 1987). Thus, the parametric model allowed some quantitative approach to design of N and P removal plants and provided a basis for evaluating the performance of existing plants (Ekama et al., 1983). This parametric BEPR model, as well the organics removal, ND models presented earlier in this chapter, were published in the NDBEPR system design guide (WRC, 1984). At the time of its publication (1984), the NDBEPR system design approach was criticized and rightly so, primarily because the influent RBO was used twice, once by the PAOs for P removal (uptake in the anaerobic reactor) and again by the OHOs for denitrification in the primary anoxic reactor. This would be possible only if in NDBEPR systems the PAOs utilize all the RBO in the primary anoxic reactor with nitrate as electron acceptor for growth and polyP accumulation in the same fashion as the RBO is completely utilized by the OHOs
Biological Nutrient Removal
in the primary anoxic reactor of the ND system. In this event the major portion of the P uptake and polyP storage by the PAOs should take place in the primary anoxic reactor of the NDBEPR systems. However, P uptake was observed taking place principally in the aerobic zone. This indicated that the denitrification behavior in NDBEPR systems is not the same as that observed ND systems so that the good predictions that had been obtained by the ND models for the NDBEPR systems were fortuitous. Denitrification behavior in NDBEPR systems is discussed in Section 4.14.34 after presenting the BEPR model based on PAO behavior. Essentially up to this time, models of NDBEPR system behavior did not recognize the presence of any specific organism mediating BEPR, only the OHOs for COD removal, denitrification, and RBO fermentation, and the ANOs for nitrification (Table 1). The parametric model in fact considered the active biomass as one group (OHOs) to represent a BEPR sludge with a propensity for P removal; variation in BEPR between different systems was modeled as changes in the propensity for P removal of OHO biomass caused by changes in influent RBSO concentration, anaerobic mass fraction, and/ or nitrate discharge to the anaerobic reactor. However, parallel research in the natural sciences had identified specific organism groups that have the propensity to store large quantities of P in the form of polyP (e.g., Buchan, 1983). This led to a shift in the approach to modeling BEPR in NDBEPR systems, from a representative OHO biomass to a specific organism group mediating BEPR, like the ANOs, the specific organism groups that mediate nitrification. The BEPR organism group became generically termed polyP organisms (Wentzel et al., 1986), bio-P organisms (Comeau et al., 1986), or PAOs (ASM2, Henze et al., 1995).
4.14.30.2.1 NDBEPR system kinetics Wentzel et al. (1988) set out to develop a general model that describes NDBEPR system behavior. They assumed that in an NDBEPR system treating municipal wastewaters, a mixed culture would develop which could be categorized into three groups of organisms: (1) heterotrophic organisms able to accumulate polyP, termed PAO; (2) heterotrophic organisms unable to accumulate polyP, termed OHOs; and (3) autotrophic organisms mediating nitrification, termed ANOs (Table 1). With regard to OHOs and ANOs, they accepted the ND models described in this chapter, viz., the steady-state (WRC, 1984) and general kinetic model (Dold et al., 1980, 1991; van Haandel et al., 1981). These models were extended to incorporate PAO behavior. To achieve this, the kinetic and stoichiometric characteristics of the PAOs in the AS environment needed to be established. From attempts to obtain information on the characteristics of the PAOs using mixed liquor from NDBEPR systems treating municipal wastewaters, Wentzel et al. (1988) noted that the OHO behavior masked the PAO behavior except in its P release, P uptake, and P removal. Accordingly, to isolate the PAO biomass characteristics, they developed enhanced cultures of PAOs in open (nonsterile) AS systems. (Serendipitously, because the UCT laboratory did not have the equipment to develop pure cultures, this was never attempted – in hindsight, this would have been the wrong approach because even today, a pure culture of PAOs has not
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yet been established.) By enhanced culture is meant a culture in which (1) the growth of PAOs is selected into the system to the extent that they become the principal organism group so that their behavior dominates the system in all the measured parameters (OUR, VSS) and (2) growth of competing organisms is selected out but not positively excluded; neither are predation or other interaction effects.
4.14.30.3 Enhanced PAO Cultures 4.14.30.3.1 Enhanced culture development From the biochemical models, Wentzel et al. (1988) were able to identify conditions to be imposed in an NDBEPR AS system to produce an enhanced PAO culture – anaerobic/aerobic sequence with adequate anaerobic mass fraction; influent fed to the anaerobic reactor with acetate as substrate and with adequate macro- and micronutrients, in particular Mg2þ, Kþ, and to a lesser degree Ca2þ, and pH control in the aerobic reactor. Using the UCT and three-stage modified Bardenpho systems, with system sludge ages ranging from 7.5 to 20 days, they developed enhanced cultures of PAOs with greater than 90% of the organisms cultured aerobically being identified as Acinetobacter spp. using the analytical profile index (API) 20NE procedure. (The API 20NE procedure has subsequently been shown to overestimate Acinetobacter spp. numbers due to the testing technique (Lotter et al. 1986; Venter et al. 1989) and selection in culturing (e.g., Wagner et al. 1994). However, for the development of the design and simulation models exact identification of the PAOs in the enhanced cultures has been of minor consequence as the models are based on quantitative experimental observations.) The response of the enhanced culture systems indicated that significant concentrations of PAOs developed. For example, the UCT system (anaerobic mass fraction 15%, sludge age 10 days, and influent of acetate at 500 mgCOD l1) gave phosphate release of 253 mgP l1, phosphate uptake of 314 mgP l1, and phosphate removal of 61 mg l1, all as mgP l1 influent flow. This BEPR behavior was much higher than observed in a mixed culture NDBEPR systems with municipal wastewater influent of 500 mgCOD l1 giving a phosphate release of 45 mg l1, phosphate uptake of 57 mg l1, and phosphate removal of 12 mgP l1. In fact, the behavior of the enhanced culture systems corresponded closely to that of the mixed culture system in terms of the influent RBO/VFA fed – at 100% and 20% influent RBO/VFA respectively for 500 mgCOD l1 feed, the enhanced culture system removed 5 times more P (61 mgP l1) than the mixed culture system. The enhanced culture mixed liquor in the aerobic zone contained 0.25–0.20 mgP/mgVSS and had a VSS/ TSS ratio of 0.46–0.48 as sludge age increased from 7.5 to 20 days, much higher than for mixed culture systems at a P/VSS ratio of 0.1 and a VSS/TSS fraction of 0.78. The low VSS/TSS ratio for the enhanced culture systems is due to the high concentration of polyP with associated counterions in the PAOs, a phenomenon later included in the model by Ekama and Wentzel (2004).
4.14.30.3.2 Enhanced culture kinetic model From experimental observations on the enhanced culture steady-state systems and on a variety of batch tests (anaerobic, anoxic, and aerobic) on mixed liquor harvested from the
Biological Nutrient Removal
4.14.30.3.3 Simplified enhanced culture steady-state model Wentzel et al. (1990) simplified the enhanced culture kinetic model, to develop a steady-state model for the enhanced culture systems under constant flow and load conditions. From an examination of the kinetics of the processes under steady-state conditions, many of the processes were virtually complete so these kinetic relationships no longer serve an important function under steady-state conditions and could be replaced by stoichiometric relationships. The three examples are given as follows: (1) The anaerobic mass fractions provided in the enhanced culture systems were sufficient to ensure that all the acetate substrate was sequestered in the anaerobic zone, that is, the kinetics of acetate storage need not be incorporated. (2) Virtually, all the substrate taken up by the PAOs in the anaerobic zone was utilized in the subsequent aerobic zone, that is, the kinetics of PHA substrate utilization (and polyP storage) did not need to be incorporated. This implied that for the
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steady-state enhanced PAO systems, Wentzel et al. (1989a) elucidated the characteristics and kinetic response of the PAO biomass. Two characteristics of the PAOs in these enhanced cultures were of particular interest: (1) very little propensity to denitrify so that no provision for this process needed to be made in modeling PAO behavior – this has important implications in modeling denitrification in mixed culture NDBEPR systems (see Section 4.14.34) and (2) an extremely low endogenous mass loss rate, 0.04 mgPAOVSS/(mgPAOVSS d) which is much lower than that of OHOs in aerobic AS system at 0.24 mgOHOVSS/(mgOHOVSS d) (Marais and Ekama, 1976). A similar observation had been made by Wentzel et al. (1985) in studies on mixed culture NDBEPR systems treating municipal wastewaters; they noted from plots of phosphate uptake versus phosphate release for various sludge ages that, for a given phosphate release, the phosphate uptake was relatively insensitive to sludge age. In modeling PAO endogenous mass loss, Wentzel et al. (1989a) used the classical endogenous respiration approach (Equation (53)), as distinct from the death-regeneration approach used for the OHOs (Section 4.14.5.4.2), except that provision was made for the situations where no external electron acceptor is available. Taking note of the above, Wentzel et al. (1989a) developed a conceptual model for PAO behavior in the enhanced cultures incorporating the characteristics, processes, and compounds identified as important from the experimental investigation. Using the conceptual model as a basis, Wentzel et al. (1989b) formulated mathematically the process rates and their stoichiometric interactions with the compounds, to develop a kinetic model for the enhanced cultures of PAO. The kinetic and stoichiometric constants of the PAOs in the enhanced cultures were quantified by a variety of experimental procedures (Wentzel et al., 1989b). With these constants, application of the kinetic model to the various batch test responses observed with the enhanced cultures gave good correlation between observations and simulations (Figures 49 and 51). The model was then applied to simulate the steadystate behavior of the enhanced culture UCT and three-stage modified Bardenpho systems, for which good correlation was also obtained. (Wentzel et al., 1989b).
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Figure 49 Experimentally observed and simulated (a) oxygen utilization rate (OUR), (b) total soluble phosphorus (PO4) and nitrate (NO3) concentrations and (c) filtered COD concentrations with time in a batch aerobic digestion test of mixed liquor from an enhanced PAO culture system. Modified from Wentzel MC, Dold PL, Ekama GA, & Marais GR (1989b) Enhanced polyphosphate organism cultures in activatedsludge systems 3. Kinetic model. Water SA 15(2): 89–102.
PAOs, like for the OHOs, the growth process could be accepted as complete so that at steady state, for a given sludge age, a constant relationship exists between the flux of acetate fed to the system and the mass of PAOs formed with stored polyP. (3) P release for anaerobic maintenance energy requirements was small compared with P release for VFA uptake energy requirements, that is, the kinetics of phosphate release for anaerobic
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maintenance energy did not need to be incorporated. However, because the endogenous respiration process is never complete, it had to be retained in the steady-state model and, as for the OHOs, was accepted to take place in all the reactors of the system. Applying these simplifications and assumptions in the steady-state PAO model indicated that the P content of the PAOs was constant with sludge age at 0.38 gP/gPAOVSS, of which 0.03 was biomass P content and 0.35 was polyP content, to account for the observed P removal. What did vary was the relative proportion of PAOs (with stored polyP) in the VSS which accounted for the difference in P removal with sludge age. The resulting steady-state PAO model was identical to the OHO model (Section 4.14.31.1.5), including the value for the PAO yield coefficient (YG ¼ 0.45 mgPAOVSS/mgCOD), but the values for the PAO unbiodegradable residue fraction (fEG) and endogenous respiration rate (bG) were different to those of the OHOs (i.e., 0.25 and 0.04 d1, respectively). The PAO steady-state model provided the means for quantifying the PAO VSS mass and its endogenous residue in mixed culture NDBEPR systems receiving municipal wastewaters as influent.
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200 Simulated Experimental
15
150
100
5
50
0
10
OUR
PO4
0
5
10 15 Time (h)
20
0 25
Carbon oxygen utilization rate (mgO I−1h−1) Carbon oxygen utilization rate (mgO I−1h−1)
70
Soluble P concentration (mgP I−1)
Soluble P concentration (mgp I−1)
80
70
400
505
Carbon oxygen utilization rate (mgO I−1h−1)
80
(a)
(b)
90
Simulated Experimental
Soluble P concentration (mgP I−1)
90
Acetate concentration (mgHAc I−1)
100
100
Acetate concentration (mgHAc I−1)
Soluble P concentration (mgP I−1)
Biological Nutrient Removal
Figure 51 Experimentally observed and simulated total soluble phosphorus (PO4) concentration and carbonaceous oxygen utilization rate (OUR)–time profiles on aeration following anaerobic acetate addition of (a) 0.207 mgCOD/mgVSS (low),(b) 0.363 mgCOD/mgVSS (moderate), and (c) 0.220 mgCOD/mgVSS (high) to mixed liquor drawn from a three-stage Bardenpho enhanced PAO culture system. Modified from Wentzel MC, Dold PL, Ekama GA, and Marais GR (1989a) Enhanced polyphosphate organism cultures in activated sludge systems 2. Experimental behaviour. Water SA 15(2): 71–88.
4.14.30.3.4 Steady-state mixed culture NDBEPR systems Mixed culture steady-state model. Having developed the steadystate model for enhanced culture systems, Wentzel et al. (1990) extended this model to incorporate mixed cultures of PAOs and OHOs present in NDBEPR systems receiving
506
Biological Nutrient Removal
domestic wastewater as influent, to give a steady-state mixed culture model. This extension proved to be possible because (1) enhanced cultures rather than pure cultures were used to establish the kinetic and stoichiometric characteristics of the PAOs. In the enhanced cultures, PAOs present in mixed culture AS were enriched and no single species was artificially selected (as in pure cultures); (2) competing organisms and predators were not artificially excluded (as in pure cultures) so that the PAOs were subjected to the same selective pressures in enhanced as in mixed cultures; (3) the PAOs were also subjected to the same conditions present in mixed culture AS systems (e.g., anaerobic/aerobic sequencing, long SRT 45 days, etc.); and (4) per influent RBO/VFA, the PAOs exhibited the same behavioral patterns in the enhanced cultures as they did in mixed culture AS systems (i.e., P release/uptake, PHA/ polyP accumulation, etc.) – in fact, the similar, though magnified behavior of the PAO enhanced culture compared to the mixed culture systems was one criterion used to establish that the correct enhanced cultures had been established. In extending the model one aspect that emerged was the difference in the endogenous mass loss rate between PAO enhanced culture sludges and the normal aerobic OHO AS. As noted earlier, the high specific endogenous mass loss rate with OHO systems had been attributed to a high rate of predation and regrowth, formulated as death regeneration in the ND kinetic model by Dold et al. (1980). The low specific endogenous mass loss rate with PAOs in the enhanced cultures systems led Wentzel et al. (1989a) to conclude that the PAOs were not predated to the same degree as OHOs, and to adopt an endogenous respiration approach in modeling PAO endogenous mass loss. (From subsequent simulations with the steady-state mixed culture model, it was found that if the PAOs were subjected to a high predation rate, then significant BEPR in the mixed culture NDBEPR system would not be possible – the rate of death of the PAOs would be so high that no significant mass of these organisms could accumulate in the system, and BEPR would be near zero.) The low predation rate on the PAOs, and the fact that the PAOs and OHOs essentially do not compete for the same substrate, implied that PAO and OHO populations act virtually independently of each other in normal mixed culture NDBEPR systems. This allowed modeling the two population groups as essentially separate, except for the fermentation F-RBO to VFA conversion process in the anaerobic reactor, which could be used to quantify the proportion of the biodegradable organics (BO) obtained by the PAOs. This rate of conversion is much slower than the rate of VFA uptake, so that the rate of conversion controls the rate of VFA uptake. Hence, the flux of VFAs that becomes available in the anaerobic reactor to the PAOs is governed by the kinetics of conversion mediated by the OHOs. The work of Me´ganck et al. (1985) and Brodisch (1985) supported this conversion approach, which is also included in the NDBEPR kinetic models (UCTPHO, Wentzel et al., 1992; ASM2, Henze et al., 1995) – they showed that anaerobic/aerobic systems developed organisms which
convert sugars and similar compounds into VFAs in the anaerobic reactor. If nitrate (or oxygen) is recycled to the anaerobic reactor, RBO is utilized preferentially by the OHOs with nitrate (or oxygen) as external electron acceptor, thereby reducing the flux of VFAs available for uptake by the PAOs. A schematic diagram showing the proportion of the influent RBO obtained by the PAOs is shown in Figure 52. OHOs obtain BO that is not obtained by PAOs. From the above, the RBSO is subdivided into two fractions, VFAs (e.g., acetate) and fermentable RBSO (FBSO, e.g., glucose). Both these fractions are measured as RBO in the conventional bioassay (e.g., Ekama et al., 1986; Wentzel et al., 1995, 1999, 2000) and filtration (e.g., Dold et al. 1986; Mamais et al., 1993; Mbewe et al., 1994) tests (see Section 4.14.4.2.2). The rate of VFA uptake by PAOs is so rapid that all influent VFAs will be taken by the PAOs even in very small anaerobic reactors (Figure 50). The F-RBO is converted to VFAs by the OHOs in the anaerobic reactor and the resultant VFAs is available for uptake by the PAOs (Figure 46). The model for this conversion is given by Wentzel et al. (1985) and will be described below. The above model provided Wentzel et al. (1990) with the means for calculating the flux of BO (influent VFA and
Unbiodegradable soluble (effluent)
Influent wastewater COD Biodegradable COD
Unbiodegradable particulate COD
RBCOD F-RBCOD
SBCOD External acid fermentation
Internal acidification
Inert VSS accumulation
Volatile fatty acids (VFAs) P accumulating organisms (PAOs)
Ordinary heterotrophic organisms (OHOs)
Enhanced culture steady-state equations PAO activemass 0.03 mgP/ mg-VSS
O2,NO3−
Usual activated sludge steady-state equations
Usual OHO activemass 0.03 mgP/mg-VSS
PAD endogenous mass 0.03 mgP/mg-VSS
OHO Endogenous mass 0.03 mgP/mg-VSS
Inert mass 0.03 mgP/ mg-VSS
Mixed VSS in system has variable P content (mass P/mass VSS %) Depending on proportion of biodegradable COD obtained by PAOs
Figure 52 Schematic diagram showing the fate of various influent COD fractions in relation to the various OHO and PAO active, endogenous, and inert masses of the sludge.
Biological Nutrient Removal
Predicted P release (mgP I−1 influent)
100 90 80 70 60 50
R S (days) 3 4 5 6 8 10 15 20 21 25 28
40 30 20 10 0
0 (a)
10 20 30 40 50 60 70 80 90 100 Measured P release (mgP I−1 influent)
30
Predicted P removal (mgP I−1)
converted of F-RBO) taken up by the PAOs in the anaerobic reactor. The remainder of the BO flux is obtained by the OHOs. In effect, the conversion model splits the influent BO (COD) into two fractions, one eventually utilized by the PAOs and the other to be utilized by the OHOs. Because of the independent action of these two organism groups, the masses of PAOs (MXBG) and their endogenous residue (MXEG) in the system could be calculated from the enhanced PAO culture steady-state model and the masses of OHOs (MXBH) and their endogenous residue (MXEH) could be calculated from the steady-state OHO model. The mass UPO in the reactor from the influent (XI) could be calculated from the unbiodegradable particulate COD fraction (fS’up) as before (Section 4.14.9.3.2). The five VSS components, each with their P content – 0.38 mgP/mgPAOVSS for the PAOs and 0.025 mgP/mgVSS for the other four components – give the average P content of the VSS. The P removal achieved by the NDBEPR system is the P in sludge mass wasted per day from the system. Wentzel et al. (1990) evaluated the predictive power of the steady-state mixed culture BEPR model against observations made on 30 laboratory-scale NDBEPR systems over a 6-year period. The system configurations were Phoredox, three-stage modified Bardenpho, UCT, MUCT, and JHB with system sludge ages ranging from 3 to 28 days. For the evaluation, the measured nitrate in the recycle to the anaerobic zone was used to estimate the fermentable COD removal in the anaerobic zone by the OHOs with nitrate as external electron acceptor. The fermentable COD remaining was available for conversion in the anaerobic reactor to VFAs and uptake and storage as PHA by the PAOs. Plots of the predicted versus measured P release, P removal, and VSS concentration (Figures 53(a)– 53(c)) show good correlation.
507
25 20 R S (days)
15
3 4 5 6 8 10 15 20 21 25 28
10 5 0 0
(b)
4.14.31 Mixed Culture Steady-State Model 4.14.31.1 Division of Biodegradable Organics between PAOs and OHOs
Sbsi ¼ Sbsai þ Sbsfi
ð173Þ
The VFA in the influent (Sbsai) is directly available to the PAOs for uptake in the anaerobic reactor.
30
4000
Predicted VSS (mg VSS I−1)
4.14.31.1.1 Subdivision of influent RBO From the mechanism for BEPR, only VFAs can be taken up directly by the PAOs in the anaerobic reactor. Accordingly, the influent RBO (Sbsi) is subdivided into two fractions: (1) VFA (Sbsai) and (2) fermentable RBO (FBSO, Sbsfi). Hence,
5 10 15 20 25 Measured P removal (mgP I−1)
3000
R S (days)
2000
3 4 5 6 8 10 15 20 21 25 28
1000
0
4.14.31.1.2 Conversion of FBSO Wentzel et al. (1985) show that the FBSO component (Sbsfi) is converted to VFA in the anaerobic reactor by the OHOs, thereby making additional VFA available to the PAOs for uptake. The rate of conversion is much slower than the rate of VFA uptake, so that the rate of conversion controls the rate of uptake of generated VFA. Wentzel et al. (1985) proposed a
0 (c)
1000
2000
3000
4000
Measured VSS (mgVSS I−1)
Figure 53 Predicted vs. measured P release (a), P removal (b) and VSS concentration (c) in a variety of BEPR systems with various configurations. From Wentzel et al. (1990).
508
Biological Nutrient Removal
first-order conversion rate, viz.,
dSbsf ¼ KCT XBHn Sbsfn dt ðmgCOD l1 h1 Þ
equations for the conversion of FBSO to VFA can be developed. This yields equations for the concentration of FBSO in exiting the nth anaerobic compartment and the mass of OHOs in the entire NDBEPR reactor, MXBH viz.,
ð174Þ
where KCT is the first-order rate constant at temperature T ¼ 0.06 l/(mgOHOVSS d) at 20 1C, and XBHn and Sbsfn the concentrations of OHOs (mgOHOVSS l1) and FRBO (mgCOD l1) exiting the nth anaerobic compartment of the anaerobic reactor.
4.14.31.1.3 Effect of recycling nitrate or oxygen When nitrate or oxygen enter the anaerobic reactor via recycle and influent flows, the OHOs utilize FBSO with these electron acceptors. Hence, the OHOs do not release the VFA generated but completely metabolize the FBSO until the oxygen or nitrate is depleted. In the conversion model this is accommodated by reducing the concentration of FBSO available for conversion, that is,
S0bsfi
¼ Sbsfi 2:86=ð1 f cv YH ÞðrNnr þ Nni Þ 1=ð1 f cv YH ÞðrOr þ Oi Þ
ð175Þ
where S0bsfi is the FBSO available for conversion to VFA (mgCOD l1 influent), Sbsfi the influent FBSO concentration (mgCOD l1), r the recycle ratio to anaerobic reactor relative to the influent flow, Nnr, Or the nitrate and oxygen concentration in the recycle to anaerobic reactor (mgNO3-N l1 and mgO l1, respectively), Nni, Oi the nitrate and oxygen concentrations in the influent to anaerobic reactor (mgNO3-N l1 and mgO l1, respectively), 2.86/(1 fcvYH) ¼ 8.6 the mass of COD utilized per unit nitrate denitrified (mgCOD/mgNO3N), and 1/(1 fcvYH) ¼ 3.0 the mass of COD utilized per unit oxygen utilized (mgCOD/mgO). Kinetics of conversion of FBSO to VFA. The conversion model proposed by Wentzel et al. (1985) assumes that: 1. Only FBSO can be converted to a form suitable for uptake by the PAOs (i.e., VFA); within the timescale of the mixed liquor in the anaerobic reactor, conversion of SBO to VFA is assumed to be negligible. 2. The conversion is mediated by the OHOs in the absence of oxygen and nitrate only. 3. All VFA generated by conversion is immediately taken up by the PAOs. 4. All FBSO not converted to VFA in the anaerobic reactor is utilized subsequently by OHOs. 5. The rate of conversion of FBSO is first order with respect to the FBSO and OHO concentrations in the anaerobic reactor and given by Equation (174). 6. All VFA present in the influent to the anaerobic reactor is immediately taken up by the PAOs.
4.14.31.1.4 Steady-state FBSO conversion equation Applying Equations (174) and (175) within mass balances over the nth anaerobic compartment in a series of N equal volume anaerobic compartments in the anaerobic reactor receiving in a continuous flow NDBEPR system, the steady-state
S0bsfi =ð1 þ rÞ n Sbsfn ¼ f xa MXBH 1 1 þ KCT ð1 þ RÞ N Qi ðmgCOD l1 Þ
ð176Þ
where fxa is the anaerobic mass fraction of the NDBEPR system, N the total number of compartments of equal volume in the anaerobic reactor, n the nth compartment of the series, n ¼ 1,2,yy,N, Sbsfn the concentration of FBSO exiting the nth compartment, MXBH the mass of OHOs in the system (mgOHOVSS), and Qi the influent flow rate (l d1). Equation (176) provides the means to calculate the flux of FBSO converted to VFA in a series of N anaerobic compartments, that is,
FSbCON ¼ Qi ½S0bsfi ð1 þ rÞSbsfN
ðmgCOD d1 Þ
ð177Þ
However, to calculate SbsfN, MXBH/Qi needs to be known. This is calculated from the flux of BO not obtained by the PAOs. All the VFA generated by conversion and all the VFA in the influent are taken up by the PAOs, so the flux of COD taken up by the PAOs, FSbPAO, is given by
FSbPAO ¼ FSbCON þ Qi Sbsai ¼ Qi ½S0bsfi ð1 þ rÞSbsfN þ Qi Sbsai
ðmgCOD d1 Þ
ð178Þ
and the flux of biodegradable COD taken up by the OHOs is given by
FSbOHO ¼ Qi Sbi FSbPAO
ðmgCOD d1 Þ
ð179Þ
Hence, from Equation (103), the mass of OHOs in the NDBEPR system is given by
MXBH ¼
FSbOHO YH Rs ð1 þ bHT Rs Þ
ðmgOHOVSSÞ
ð180Þ
Substituting Equations (179) and (178) into Equation (180) and dividing by Qi yields the MXBH/Qi required in Equation (176), viz.,
MXBH ðSbi ½S0bsfi ð1 þ rÞSbsN þ Sbsai ÞYH Rs ¼ Qi ð1 þ bHT Rs Þ ðmgOHOVSSÞ
ð181Þ
Equations (176) and (181) need to be solved simultaneously to calculate the concentration of FBSO (SbsfN) exiting the last anaerobic compartment (N); the following procedure converges in three to four iterations: (1) Assume SbsfN ¼ 0 mgCOD l1, (2) calculate MXBH/Qi with Equation (181), (3) with MXBH/Qi known, calculate SbsfN with Equation (176), (4) recalculate MXBH/Qi using the new value for SbsfN, (5) repeat steps (3)–(5) until SbsfN and MXBH/Qi are constant. This procedure splits the influent BO between the OHOs and PAOs. Because the growth processes of two organism
Biological Nutrient Removal
groups after the anaerobic reactor are noncompetitive and VFA uptake process and the growth processes on the available organics are complete for both groups, the stoichiometric equations relating the flux of COD utilized and the biomass produced derived earlier (Equation (103)) can be applied to calculate the PAO and OHO masses and their endogenous residue masses.
509
(Supi, XIi) (Equation (67)), viz.,
MXI ¼ FSti f S0 up =f cv Rs
ðmgIVSSÞ
Total VSS in the NDBEPR system is the sum of the five VSS components:
MXv ¼ MXBH þ MXBG þ MXEH þ MXEG þ MXI ðmgVSSÞ
4.14.31.1.5 Mass of VSS in the NDBEPR system
ð188Þ
PAO mass
MXBG ¼ FSbPAO
YG Rs 1 þ bGT Rs
ðmgPAOVSSÞ
ð182Þ
where, YG is the PAO yield coefficient (mgPAOVSS/mgCOD utilized), FSbPAO the flux BO taken up by PAOs in the anaerobic reactor (mgCODd1), and bGT the PAO specific endogenous mass loss rate constant at temperature T (d1).
4.14.31.1.6 PAO P release From the mechanisms of BEPR (Wentzel et al., 1985, 1990), for every mole of VFA taken up, 1 mol of P is released to provide energy to synthesize and store the VFA as PHA. Accordingly, the P release in the anaerobic reactor is given by
FPrel ¼ f prel FSbPAO
ðmgP d1 Þ
ð189aÞ
PAO endogenous mass or
MXEG ¼ f EG bGT MXBG Rs
ðmgVSSÞ
ð183Þ Prel ¼ f prel SbPAO
where fEG is the fraction of PAOs that is unbiodegradable particulate endogenous residue. PAO oxygen demand
FOGc ¼ FOGs ðsynthesisÞ þ FOGe ðendogenous respirationÞ ¼ ð1 f cv YG ÞFSbPAO þ f cv ð1 f EG ÞbGT MXBG YG Rs ¼ FSbPAO ð1 f cv YG Þ þ f cv ð1 f EG ÞbGT 1 þ bGT Rs ðmgO d1 Þ
ð184Þ
OHO mass
MXBH
ðmgP l1 influentÞ
where fprel is the ratio P release/VFA uptake E1.0 molP/mol COD E0.5 mgP/mgCOD and SbPAO the concentration COD taken up by the PAOs per liter influent ¼ FSbPAO/Qi.
4.14.31.1.7 P removal The P removal via the waste sludge is calculated from the individual P content of the five VSS components, viz: By PAOs
MXBG MXEG 1 DPG ¼ f XBGP þ f XEGP Rs Rs Qi ðmgP l1 influentÞ
YH Rs ¼ FSbOHO 1 þ bHT Rs
ðmg OHOVSSÞ
ð185Þ
where YH is the OHO yield coefficient (mgOHOVSS/mgCOD utilized), FSbOHO the flux BO taken up by OHOs in the anaerobic reactor (mgCOD d1), and bHT the OHO specific endogenous mass loss rate constant at temperature T (d1).
ðmgVSSÞ
By OHOs
ð186Þ
FOHc ¼ FOHs ðsynthesisÞ þ FOHe ðendogenous respirationÞ ¼ ð1 f cv YH ÞFSbOHO þ f cv ð1 f EH ÞbGT MXBH YH Rs ¼ FSbOHO ð1 f cv YH Þ þ f cv ð1 f EH ÞbHT bHT Rs ðmgO d Þ
where PG is the P removal by the PAOs (mgP l1influent), fXBGP the P content of PAOs ¼ 0.38 mgP/mgPAOVSS, and fXEGP the P content PAO endogenous mass ¼ 0.03 mgP/ mgEVSS.
MXBH MXEH 1 DPH ¼ f XBHP þ f XEHP Rs Rs Qi
where fEH is the fraction of OHOs that is unbiodegradable particulate endogenous residue. OHO oxygen demand
1
ð190Þ
OHO endogenous mass
MXEH ¼ f EH bHT MXBH Rs
ð189bÞ
ðmgP l1 influentÞ
ð191Þ
where PH is the P removal by the OHOs (mgP l1influent), fXBHP the P content of OHOs ¼ 0.03 mgP/mgOHOVSS, and fXEHP the P content OHO endogenous mass ¼ 0.03 mgP/ mgEVSS. By inert mass
MXI 1 DPI ¼ f XIP Rs Qi
ðmgP l1 influentÞ
ð192Þ
ð187Þ
The same equations derived earlier in Section 4.14.7.1.1 apply for the UPO that accumulate in the reactor from the influent
where PI is the P removal due to inert mass (mgP l1 influent) and fXIP the P content inert VSS mass (mgP/mgIVSS) ¼ 0.025– 0.03 mgP/mgIVSS.
510
Biological Nutrient Removal
The total P removal is given by the sum of the individual P removals, i.e. Total P removal
DPT ¼ DPG þ DPH þ DPI
ðmgP l1 influentÞ
ð193Þ
The effluent P concentration is given by the difference between the influent P and the P removal, i.e. Effluent P concentration
Pte ¼ Pti PT
ðmgP l1 Þ
mgPAOVSS), and the polyP ISS, which is 3.286 mgISS/mgP times the PAO polyP content, which is its total P content (fXBGP) minus its biomass P content (Ekama and Wentzel, 2004). Hence,
ð194Þ
If the P removal is greater than the influent P concentration, then the expectation is that the effluent P concentration will be below 0.5 mgP l1. How far below 0.5 mgP l1 is uncertain because currently this appears to be plant specific. Research is being conducted to investigate what the limits of BEPR technology are and what conditions in the NDBEPR system cause them (Neethling et al., 2009). Revised PAO P content (fXBGP). If the P removal is greater than the influent P concentration, then there is insufficient P in the influent for the PAOs to take up P up to their maximum P content of 0.38 mgP/mgPAOVSS. Their P content (fXBGP) will therefore be limited by the available P. Under these conditions, the PAO P content needs to be revised to match the available P. If this is not done, the reactor ISS concentration, which is strongly influenced by the PAO P content, will be overestimated. In the calculation for the revised PAO P content, it is assumed that the effluent P concentration (Pte) is equal to the P concentration of the unbiodegradable soluble organics (USO; see Section 4.14.4.4.3) and that the P content of the non-PAO VSS components remains unchanged. Unless data are available to indicate a nonzero USO P concentration (Pousi40), it is reasonable to accept it as zero. Clearly, if the wastewater contains USO P, then this will impact achieving the very low effluent P standards that are being set for NDBEPR systems these days. However, it would appear that USO P in municipal wastewaters is effectively zero, or at least masked by the scatter of the difference between membrane filtered effluent TP and OP concentrations. The revised PAO P content (fXBGP) is found by making fXBGP the subject of Equation (193) and the P removal equal to the difference between the influent P and USO P concentrations (Equation (39)), viz.,
f XBGP ¼ ½ðPti Pousi ÞQi Rs f XEGP MXEG f XBHP MXBH f XEHP MXEH f XIP MXI =MXBG ðmgP=mgPAOVSSÞ ð195Þ
4.14.31.2 VSS and TSS Sludge Masses in the Reactor (System) The VSS mass in the NDBEPR reactor is the sum of the five VSS component masses (Equation (188)). The ISS concentration is the sum of the ISS that accumulates in the reactor from the influent (Equation (97)), the OHO ISS, and the PAO ISS. The OHO ISS is 15% of its VSS mass, that is, fiOHO ¼ 0.15 mgISS/ mgPHOVSS (Equation (98)). The PAO ISS is the sum of its biomass ISS, which is the same at the OHO ISS (0.15 mgISS/
XIO ¼ FXIOi Rs þ MXBH þ 3:286ðf XBGP f XBGPBM ÞMXBG ðmgISSÞ ð196Þ where fXBGPBM is the PAO biomass P content ¼ OHO biomass P content ¼ 0.025–0.03 mgP/mgPAOVSS. The TSS mass in the NDBEPR system is the sum of the VSS and ISS masses, that is,
MXt ¼ MXv þ MXIO
ðmgTSSÞ
ð197Þ
This TSS mass is distributed in the various reactors of the NDBEPR system, not necessary at the same TSS concentration in each reactor. The reactor configuration (Figure 47) influences the TSS concentration in the different reactors of the system. Calculating the reactor concentrations from the various mass fraction of the reactors is discussed below.
4.14.31.3 BEPR System Design Considerations 4.14.31.3.1 Process volume requirements An approximate reactor volume, that is, a nonconfigurationspecific volume, can be estimated from a selected average reactor TSS concentration required for the system, that is,
Vp ¼ MXt =Xt
ðm3 Þ
ð198Þ
where Xt is the zone/reactor volume weighed average TSS concentration in the NDBEPR system (mgTSS l1). For all NDBEPR system configurations with SSTs, or with membranes (MBR), at steady-state and average dry weather flow (ADWF) conditions, the concentrations of TSS in the preanoxic (Figure 47(f)) and anaerobic (if present) and anoxic and aerobic zones (Xtpax, Xtana, Xtanx, Xtaer), as fractions of the average system TSS concentration Xt are equal to the ratio of the sludge mass fraction and volume fraction of the zones, that is,
Xtana f mana Xtanx f manx ¼ ; ¼ ; Xt f vana Xt f vanx Xtpax f mpax Xtaer f maer ¼ ; ¼ Xt f vaer Xt f vpax
ð199Þ
where fm, fv are the zone sludge mass and volume fractions respectively, and subscripts ana, anx, aer, and pax are the anaerobic, anoxic, aerobic, and pre-anoxic zones, respectively. For BNR systems with SSTs in which the sludge mass is uniformily distributed, that is, the TSS concentrations are the same in the anaerobic, anoxic, and aerobic zones of the reactor, the sludge mass and volume fractions are equal, such as in the three- and five-stage Bardenpho systems (Figures 47(b) and 47(c)) for N and P removal and the pre- (modified Ludzack–Ettinger, MLE) and post-(Wuhrmann) denitrification and four-stage Bardenpho systems for N removal. For example, if an MLE ND system (Figure 34(b)) requires anoxic and aerobic mass fractions (fmanx, fmaer) of 0.45 and 0.55,
Biological Nutrient Removal
respectively, or a three-stage Bardenpho (Figure 47(b)) system requires anaerobic, anoxic, and aerobic mass fractions (fmana, fmanx, fmaer) of 0.15, 0.35, and 0.50 respectively, the corresponding volume fractions of these zones (fvana, fvanx, fvaer) with respect to the reactor volume (VR) will also be 0.45 and 0.55 for the MLE system and 0.15, 0.35, and 0.50 for the three-stage Bardenpho system. This is because the influent flow dilutes the SST return sludge concentration in the first zone by the same amount as the SST concentrates it after the last zone. This equality of sludge mass and volume fractions does not apply to any multizone BNR system with membrane solid–liquid separation in the aerobic zone, because the aerobic zone concentration is in effect the equivalent of the return sludge concentration from the SST (if there were SSTs). For BNR systems with SSTs, in which the TSS concentrations are not the same in the pre-anoxic, anaerobic, anoxic, or aerobic zones (e.g., in the UCT (Figure 47(d)) or in the JHB (Figure 47(f)) systems), the volume and mass fractions are not equal. For the UCT system, the volume fractions (with respect to Vp) of the anaerobic, anoxic, and aerobic zones (fvana, fvanx, fvaer), and the anaerobic, anoxic, and aerobic TSS concentrations (Xtana, Xtanx, Xtaer) at steady-state and ADWF conditions are related to the anaerobic and aerobic mass fractions (fmana, fmaer), recycle ratio (r) from the anoxic to the anaerobic reactor and system average TSS concentration Xt , as follows:
f mana ðr þ 1Þ rB
ð200aÞ
ð1 f mana f maer Þ B
ð200bÞ
f maer B
ð200cÞ
rB ðr þ 1Þ
ð200dÞ
f vana ¼ f vanx ¼
f vaer ¼
Xtana ¼ Xt
Xtanx ¼ Xtaer ¼ Xt B
ð200eÞ
f mana 1þ r
ð200f Þ
where
B¼
For the JHB system with SSTs, assuming the influent flow to the pre-anoxic zone, which is sometimes included to increase pre-denitrification, is zero, the volume fractions (with respect to Vp) of the pre-anoxic, anaerobic, anoxic, and aerobic zones (fvpax, fvana, fvanx, fvaer), and the pre-anoxic, anaerobic, anoxic, and aerobic TSS concentrations (Xtpax, Xtana, Xtanx, Xtaer) at steady-state and ADWF conditions are related to the pre-anoxic, anaerobic, and aerobic mass fractions (fmpax, fmana, fmaer), underflow recycle ratio (s) from the SST to the pre-anoxic reactor and average TSS concentration Xt , as follows:
f mana C
ð201aÞ
ð1 f mana f maer f mpax Þ C
ð201bÞ
f vana ¼ f vanx ¼
f maer C
ð201cÞ
f mpax s Cðs þ 1Þ
ð201dÞ
f vaer ¼ f vpax ¼
Xtana ¼ Xtanx ¼ Xtaer ¼ Xt C
C¼
ð201eÞ
Cs ð1 þ sÞ
ð201f Þ
f mpax 1 1þs
ð201gÞ
Xtpax ¼ Xt where
511
In BNR systems with membrane solid–liquid separation in the aerobic zone, the sludge mass distributes itself differently in the different zones of the system compared with systems with SSTs. This is because the effluent is withdrawn via the membranes from the aerobic zone which concentrates the sludge in this zone relative to that in the other zones. However, in recycling this concentrated aerobic zone sludge to an upstream zone, it is diluted by the less concentrated incoming sludge stream from the upstream zones. The higher the recycles from downstream zones to upstream zones, the more uniformily the sludge mass is distributed around the system and the closer the sludge concentrations in the different zones. For the UCT system with membranes, the volume fractions (with respect to Vp) of the anaerobic, anoxic, and aerobic zones (fvana, fvanx, fvaer), and the anaerobic, anoxic, and aerobic TSS concentrations (Xtana, Xtanx, Xtaer) at steady-state and ADWF conditions are related to the anaerobic and aerobic mass fractions (fmana, fmaer), recycle ratio (r) from the anoxic to the anaerobic zone, recycle ratio (a) from the aerobic to the anoxic zones, and system average TSS concentration Xt , as follows:
f mana ðr þ 1Þ Dr
ð202aÞ
ð1 f mana f maer Þ D
ð202bÞ
af maer ða þ 1ÞD
ð202cÞ
rD ðr þ 1Þ
ð202dÞ
f vana ¼
f vanx ¼
f vaer ¼
Xtana ¼ Xt
Xtanx ¼ Xt D Xtaer ¼ Xt where
D¼
ða þ 1ÞD a
f mana f maer 1þ r ða þ 1Þ
ð202eÞ ð202f Þ
ð202gÞ
For the JHB system with membranes, the volume fractions (with respect to Vp) of the pre-anoxic, anaerobic, anoxic, and aerobic zones (fvpax, fvana, fvanx, fvaer), and the pre-anoxic, anaerobic, anoxic, and aerobic TSS concentrations (Xtpax, Xtana,
512
Biological Nutrient Removal
Xtanx, Xtaer) at steady-state and ADWF conditions are related to the pre-anoxic, anaerobic, and aerobic mass fractions (fmpax, fmana, fmaer), recycle ratio (s) from the aerobic to the pre-anoxic zones, recycle ratio (a) from the aerobic to the anoxic zones, and average TSS concentration Xt , as follows:
f vana ¼ f vanx ¼
f mana ð1 þ sÞ sE
ð1 f mana f maer f mpax Þða þ s þ 1Þ ða þ sÞE
ð203aÞ
ð203bÞ
f vaer ¼
f maer E
ð203cÞ
f vpax ¼
f mpax E
ð203dÞ
sE ðs þ 1Þ
ð203eÞ
Xtana ¼ Xt
Eða þ sÞ Xtanx ¼ Xt ða þ s þ 1Þ Xtaer ¼ Xtpax ¼ Xt E
ð203f Þ ð203gÞ
ð1 þ sÞ ða þ s þ 1Þ þ f manx E ¼ f mana s ða þ sÞ þf maer þ f mpax
ð203hÞ
Equation (203) applies also to the MLE ND system and the three-stage Bardenpho system with membranes. In Equation (203h) for E, for the MLE system, the anaerobic and pre-anoxic mass fractions are both set to zero, the anoxic mass fraction is 1 minus the aerobic mass fraction (i.e., fmanx ¼ 1 fmaer), and the mixed liquor recycle ratio (a) is also set to zero – only one recycle (s) is required to return nitrate and sludge to the anoxic reactor. For the three-stage Bardenpho system, only the pre-anoxic sludge mass fraction (fmpax) is set to zero. From Equations (200)–(203), the volumes of, and the TSS concentrations in, the various zones of common BNR systems with SST or membrane solid–liquid separation can be calculated for selected anaerobic, aerobic, and pre-anoxic mass fractions (fmaer, fmana, fmpax), and interzone recycle ratios (a, r, and s). In the derivation of these equations, steady-state conditions were assumed and the sludge waste flow rate was ignored – the effect of this is negligible (o2%), especially if the sludge age is long. Generally, a uniform distribution of sludge mass in BNR MBR systems will not occur, even in systems with a single recycle flow from the aerobic to the zone receiving the influent flow. For example, changing an MLE ND system, or a three-stage Bardenpho system with SSTs to membrane solid– liquid separation systems, will change these systems from uniformly distributed sludge mass systems in which the sludge mass and volume fractions are equal to nonuniformly distributed sludge mass systems in which the sludge mass and volume fractions are different, the magnitude of difference depending on the magnitude of the recycle ratios. In multizone BNR systems with membranes in the aerobic reactor and fixed volumes for the anaerobic, anoxic, and
aerobic zones (i.e., fixed volume fractions), the mass fractions can be varied (within a range) by varying the inter-reactor recycle ratios. For example, in a UCT system with anaerobic, anoxic, and aerobic zone volume fractions of 0.25, 0.35, and 0.40 and an r recycle ratio from the anoxic to the anaerobic zones of 1:1, the anaerobic, anoxic, and aerobic zone mass fractions can be varied from 0 to 0.131, 0 to 0.366, and 1 to 0.503, respectively, by changing the a recycle ratio from 0:1 to 5:1. Increasing the a recycle ratio concomitantly increases the nitrate load on the anoxic reactor, thereby increasing the denitrification and N removal as the anoxic mass fraction increases. Increasing the r recycle ratio increases the anaerobic mass fraction (at the expense of the other two zone mass fractions) and increases (not proportionally) the P removal. This zone mass fraction flexibility is a significant advantage of membrane BNR systems over conventional BNR system with SSTs because it allows changing the mass fractions to optimize biological N and P removal in conformity with influent wastewater characteristics and the effluent N and P concentrations required. If required, the performance of membrane BNR systems can be simulated with current BNR AS models such as UCTOLD (for ND, Dold et al., 1991), UCTPHO (for NDBEPR with 490 aerobic P uptake BEPR, Wentzel et al., 1992; Hu et al., 2003), and IWA ASM Nos 1, 2 (ND and BEPR, Henze et al., 1987) by returning the SST underflow into the aerobic zone from which the SST feed flow exits (Parco et al., 2009). However, such simulations require a priori information on the reactor and zone volumes and recycle flows, which would need to be determined with the steady-state procedures set out in this chapter.
4.14.31.3.2 Nitrogen requirements for sludge production The form of the equation for calculating the nitrogen requirement for sludge production (Ns, mgN l1 influent) is the same as set out in Section 4.14.21, Equation (142), that is,
Qi Ns ¼ f n MXv =Rs
ðmgN d1 Þ
However, for the BEPR system the term MXv needs to take account of the changes in VSS components, that is, it must be calculated using Equation (188). Effect of this is to increase Ns because MXv is greater in the NDBEPR system than in the same sludge age ND system receiving the same wastewater. The increase in Ns decreases the nitrification capacity (Nc) (Equation (152)), and hence also the nitrification oxygen demand (FOn, mgO d1). For the rest, the nitrification model calculations remain the same.
4.14.31.3.3 Total oxygen demand The carbonaceous oxygen demand (FOc) is the sum of oxygen demands of PAOs (Equation (184)) and OHOs (Equation (187)):
FOc ¼ FOGc þ FOHc ¼ ð1 f cv YG ÞFSbPAO þ f cv ð1 f EG ÞbGT MXBG þ ð1 f cv YH ÞFSbOHO þ f cv ð1 f EH ÞbHT MXBH ðmgO d1 Þ
ð204Þ
Biological Nutrient Removal
4.14.32 Influence of BEPR on the System 4.14.32.1 Influence on VSS, TSS, and Carbonaceous Oxygen Demand The model for BEPR systems presented above enables the VSS and TSS of the mixed liquor (Equations (188) and (197), respectively) and the carbonaceous oxygen demand (Equation (204)) to be calculated. A comparison of the masses of VSS and TSS in the reactor and the carbonaceous oxygen demand per kg COD load on the bioreactor versus sludge age with and without BEPR are shown in Figures 54(a) and 54(b) for the example raw and settled wastewaters respectively, with influent RBO fractions with respect to the biodegradable COD (fSb’s) of 0.25 and 0.38, respectively (Table 14) and an VFA fraction of 25% of the RBO, that is, influent RBO and VFA concentration of 146 and 36 mgCOD l1 for both wastewaters. The features of the BEPR system are a UCT configuration operated at 20 1C with two equal-sized in-series anaerobic compartments with a total anaerobic mass fraction (fxana) of 0.15, an anoxic to anaerobic (r) recycle of 1:1, and no nitrate recycled to the anaerobic reactor. From Figures 54(a) and 54(b), BEPR in the AS system increases the VSS slightly, by about 5–12% and 15–25% for
Raw wastewater
8
0.8 MLTSS Oxygen demand
0.6
0.4
4 MLVSS 2
0.2
0.0
0 0
5
Although there is only a small difference in VSS production between a BEPR and a non-BEPR system, the constituent
10 15 20 25 Sludge age (days)
30
Settled wastewater
10 Sludge mass − kgVSS/(kgCOD/d) reactor
With BEPR No BEPR 20 °C
6
4.14.32.2 VSS Composition
0.1
Oxygen demand − (kgO/d)/(kgCOD/d)
Sludge mass − kgVSS/(kgCOD/d) reactor
10
raw and settled wastewaters respectively depending on sludge age – the longer the sludge age the greater the difference. This increase in VSS is due to the lower endogenous mass loss/ death rate of the PAOs (0.04 d1 at 20 1C) compared with the OHOs (0.24 d1 at 20 1C). However, the TSS is increased substantially, by about 20–25% and 45–55% for raw and settled wastewaters, respectively, depending on the sludge age. This higher TSS production is due to the large quantities of stored inorganic polyP and the associated inorganic cations necessary to stabilize the polyP chains – principally Mg2þ and Kþ (Fukase et al., 1982; Arvin, 1985; Comeau et al., 1986; Wentzel et al., 1988; Ekama and Wentzel, 2004). The high inorganic content of the PAOs causes the VSS/TSS to be much lower than that of the OHOs, 0.46 mgVSS/mgTSS compared with 0.87 mgVSS/mgTSS (excluding the influent ISS). Thus, the higher the PAO fraction of the mixed liquor, the higher the BEPR, but the lower the VSS/TSS ratio of the mixed liquor. The increase in TSS with the inclusion of BEPR needs to be taken into account in the design of the bioreactor volume (Equation (198)) and daily sludge production. Also, since the inorganic cations that stabilize the polyP are derived from the influent wastewater, there must be sufficient concentrations of these cations in the influent; otherwise, the BEPR may be adversely affected (Wentzel et al., 1989a; Lindrea et al., 1994). Further, because the VSS mass generated per kg COD load is greater with BEPR than without, the oxygen demand with BEPR is correspondingly reduced, by about 5–6% and 8–9% for raw and settled wastewaters, respectively (depending on sludge age, Figures 54(a) and 54(b)).
1.0
With BEPR No BEPR 20 °C 8
0.8 Oxygen demand 0.6
6 MLTSS
0.4
4
MLVSS
2
0.2
Oxygen demand − (kgO/d)/(kgCOD/d)
The total oxygen demand (FOt, mgO d1) is the sum of the carbonaceous and nitrification oxygen demands, taking due account of the change in nitrogen requirements for sludge production (Ns) and nitrification capacity (Nc). For a nonnitrifying BEPR systems, the total oxygen demand FOt is given by FOc. Including nitrification in the BEPR system necessarily means that denitrification must also be included; the effect of nitrification and denitrification on the total oxygen demand will be considered in Section 4.14.34.
513
0.0
0 0
5
10 15 20 25 Sludge age (days)
30
Figure 54 Predicted masses of volatile solids (MXV, MLVSS) and total solids (MXt MLTSS) and daily carbonaceous oxygen demand (FOC) per kg COD load on the biological reactor in ND (thin line) and BEPR (bold line) systems treating raw (a) raw and settled (b) wastewater.
Biological Nutrient Removal
Settled wastewater with BEPR
200
I−1
750 mg COD 25% RBCOD fraction % Of VSS mass (settled WW)
% Composition of VSS mass
Inert 60 OHO endogenous 40 OHO active PAO endog 20
90 80
Additional VSS mass in system treating raw WW, i.e., raw WW produces about 100% more activated sludge VSS mass
140 120
70 60
100
50 Inert
80
40
OHO endogenous
60
30
OHO active
PAO endog
40
20 10
PAO active
0
0 0
5
10
15
20
25
5
% Of VSS mass (settled WW)
Inert mass 60 OHO endogenous mass
40
20
OHO active mass
20
25
30
100
450 mg COD I−1 38% RBCOD fraction
180
80
15
Settled wastewater no BEPR
200
−1
750 mg COD I 25% RBCOD fraction
10
Sludge age (days)
(b)
Raw wastewater no BEPR
100
0 0
30
Sludge age (days)
(a)
% Composition of VSS mass
160
20
PAO active
90 80
160 Additional VSS mass in system treating raw WW. i.e., raw WW produces about 100% more activated sludge VSS mass
140 120
70 60 50
100 Inert
40
80 OHO endogenous
60
30 20
40 OHO active
20
10 0
0
0 0 (c)
450 mg COD 38% RBCOD fraction
180
80
100
I−1
5
10
15
20
25
0
30
Sludge age (days)
(d)
% Of VSS mass (raw WW)
Raw wastewater with BEPR
100
% Of VSS mass (raw WW)
514
5
10
15
20
25
30
Sludge age (days)
Figure 55 Percentage composition of VSS mass for BEPR systems (a, c) and ND systems (No BEPR, b, d) treating raw (a, b) and settled (c, d) wastewater.
sludge fractions for the two systems differ markedly. This can be readily demonstrated by comparing the percentage composition of the VSS mass generated in systems exhibiting BEPR with the ND system that does not. To illustrate, percentage composition of the VSS mass is shown in Figures 55(a) to 55(d) for systems at 20 1C with no BEPR (Figures 55(b)– 55(d)) and with BEPR (Figures 55(a) and 55(c)) respectively treating the example raw (Figures 55(a) and 55(b)) and settled (Figures 55(c) and 55(d)) wastewaters. Note that the BEPR system has a smaller OHO active mass than the no BEPR ND system, but the BEPR system has additionally a significant concentration of PAO biomass.
4.14.32.3 P/VSS ratio A parameter often used to evaluate the BEPR performance of an AS system is the P/VSS (or P/TSS) ratio of the mixed liquor. In Figures 56(a) and 56(b), the calculated P/VSS ratio for a BEPR system with a two-compartment anaerobic reactor and the example raw and settled wastewater characteristics are plotted versus sludge age. A zero discharge of nitrate to the anaerobic reactor is assumed. From Figures 56(a) and 56(b), as the system sludge age increases, the P/VSS ratio increases up to a sludge age of about 10 days. Further increases in sludge age cause a decrease in P/VSS ratio. The initial increase in
Biological Nutrient Removal 15
20
10
0.05 fxana
5
0.0
0
15 0.10 0.05
10
fxana 5 0.0
0 5
0 (a)
10 15 20 Sludge age (days)
25
30
0
5
(b)
10 15 20 Sludge age (days)
25
30
20
15
Settled WW
10
%P in TSS (mgP/mgVSS*100)
Raw WW %P in TSS (mgP/mgVSS*100)
0.25 0.20 0.15
Settled WW
0.25 0.20 0.15 0.10
%P in VSS (mgP/mgVSS*100)
%P in VSS (mgP/mgVSS*100)
Raw WW
0.25 0.20 0.15 0.10
5 fxana
0.05 0.0
15
0.25 0.20 0.15 0.10
10
0.05 fxana
5
0.0
0
0 0 (c)
515
5
10 15 20 Sludge age (days)
25
0
30 (d)
5
10 15 20 Sludge age (days)
25
30
Figure 56 Predicted percentage phosphorus to VSS (P/VSS 100; a, b) and TSS (P/TSS 100; c, d) ratios vs. sludge age for mixed liquor in a BEPR system with various anaerobic mass fractions (fxana) treating the example raw (a, c) and settled (b, d) wastewater.
P/VSS with sludge age is due increasing OHO mass with sludge age, which increase the fermentable RBO to VFAs conversion efficiency in the anaerobic reactor and accordingly yields an increased PAO mass (with associated P content of 0.38 mgP/mgVSS). The decrease in P/VSS can be ascribed to the endogenous respiration effect on PAOs. The P/VSS ratio is therefore a consequence of the selection of the system design parameters sludge age and anaerobic mass fraction and wastewater characteristics. Accordingly, the P/VSS ratio can neither fulfill a function in BEPR plant design nor in BEPR performance assessment between different BEPR plants.
4.14.33 Factors Influencing Magnitude of BEPR The influence of the main design parameters on the magnitude of P removal is demonstrated with the mixed culture steady-state BEPR model. These main parameters are: raw settled wastewater (Sti ¼ 750 and 450 mgCOD l1 respectively, Tables 7, 11, 14), sludge age (SRT ¼ 20 days), anaerobic sludge
mass fraction (fxana ¼ 0.15), influent RBO COD fraction (fSb’s ¼ 0.25 for raw and 0.385 for settled), discharge of nitrate and oxygen to the anaerobic reactor (0 for both) and subdivision of anaerobic reactor into compartments (N ¼ 2). The numbers in brackets are the default wastewater characteristics and system design parameter values.
4.14.33.1 Sludge Age and Anaerobic Mass Fraction Using the characteristics of the example raw and settled wastewater with a total influent COD of 750 and 450 mgCOD l1 respectively, assuming (1) no nitrate and DO enters the anaerobic reactor, (2) a recycle ratio to the anaerobic (r) of 1:1 and the anaerobic reactor is subdivided into two compartments, the P removal (normalized with respect to influent COD, mgPmg1 influent COD) versus sludge age is shown in Figures 57(a) (raw) and 57(b) (settled) for anaerobic mass fractions of 0.00 (no BEPR), 0.05, 0.10, 0.15, 0.20, and 0.25. In the same figures, the actual P removal in mgP l1 is shown on the right-hand axis.
Biological Nutrient Removal
0.02
0.25 0.20 0.15 0.10
22.5
15.0
Influent P 0.05
fxana
0.01
7.5
0.0
5
10 15 20 25 Sludge age (days)
0.04 0.03
Influent P
0.02
0.25 0.20 0.15 0.10
18.0 13.5 9.0
0.05
fxana 4.5
0.01 0.0
0.00
0.0 0
22.5 Settled WW
0.0 0
30 (b)
P removal (mgP I−1)
0.03
0.00 (a)
0.05
30.0 Raw WW
P removal (mgP/mg Infl COD)
P removal (mgP/mg Infl COD)
0.04
P removal (mgP I−1)
516
5
10 15 20 25 Sludge age (days)
30
Figure 57 Predicted P removal vs. sludge age for various anaerobic mass fractions (fxana) for a two-compartment anaerobic reactor BEPR system at 20 days sludge age, treating the example raw (a) and settled (b) wastewater.
The effect of sludge age on P removal is complex. For sludge age o3 days, the P removal increases with increase in sludge age and for sludge age 43 days, P removal decreases with increase in sludge age. The reason for this is that an increase in sludge age causes an increase in the system OHO mass, which in turn causes an increase in fermentable RBO conversion and, therefore, an increase in P release, P uptake, and P removal. However, the increased sludge age also causes a decrease in PAO biomass, its associated P content, due to endogenous respiration which decreases the P removal. At sludge age o3 days, the former effect dominates the P removal, while at sludge age 43 days the latter dominates. The decrease in both PAO biomass with increase in sludge age is crucially affected by the specific endogenous mass loss rate of the PAOs – should the endogenous mass loss rate of the PAOs (0.04 d1) have been the same as that of the OHOs (0.24 d1), virtually no BEPR would have been obtained. The effect of anaerobic mass fraction (fxana) on P removal also is shown in Figures 57(a) and 57(b). For a selected sludge age, an increase in fxana always gives rise to an increase in P removal. This is due to the increased conversion of fermentable RBO with larger anaerobic mass fractions. The improvement in P removal, however, diminishes with each step increase in fxana, due to the first-order nature of the RBO conversion kinetics. From Figures 57(a) and 57(b), it can be seen that for fxana 40.15 only small additional increases in P removal are obtained, which usually are not justified due to the decrease in unaerated sludge mass fraction this causes and the consequent impact on the minimum sludge age for nitrification.
4.14.33.2 Raw and Settled Influent The effect of primary settled wastewater on P removal can be seen by comparing Figures 57(a) and 57(b), which show the P removal for the raw wastewater of original COD of 750 mgCOD l1 and the settled wastewater produced from the raw wastewater with a 450 mgCOD l1. It can be seen that although the P removal per mg influent COD is higher for the settled WW, the P removal in mgP l1 is lower. This decrease is due to the decrease in the flux of biodegradable COD entering
the AS system which causes a reduction in OHO biomass and hence in the fermentable RBO converted and in the mass of PAOs generated. The P removal per influent COD entering the biological reactor is higher for the settled because the fraction of the biodegradable organics that is RBO (Sbsi/Sbi) is higher for settled than for unsettled wastewater because primary settling removes only the settleable organics (although not strictly true, RBO loss or gain in primary settling appears to be small; it is assumed that the RBO is not changed during primary settling).
4.14.33.3 Influence of Influent RBO Fraction Assuming zero discharge of nitrate to the two-compartment anaerobic reactor of mass fraction (fxana) of 0.15, the effect of the influent RBO fraction (fSb’s ¼ Sbsi/Sbi) is illustrated in Figures 58(a) and 58(b) for raw and settled wastewaters, respectively. At any anaerobic mass fraction, the higher the influent RBO fraction, the higher the P removal. In design, one option to improve the P removal is supplementation of influent RBO by, for example, acid fermentation of primary sludge (Pitman et al., 1983; Barnard, 1984; Osborn et al., 1989) or dosing other RBO or VFA into the anaerobic reactor.
4.14.33.4 Influence of Recycling Nitrate and Oxygen to the Anaerobic Reactor The influence of nitrate recycled to the anaerobic reactor is illustrated in Figures 59(a) and 59(b) which show P removal versus nitrate concentration recycled to the anaerobic reactor in a recycle ratio 1:1. Clearly, in agreement with numerous experimental and full-scale NDBEPR systems, recycling nitrate has a markedly deleterious influence on the magnitude of P removal. As the nitrate concentration recycled to the anaerobic reactor increases, the P removal decreases. The same applies to oxygen entering the anaerobic reactor, except that its effect is 1/2.86 times that of nitrate because the oxygen equivalent of nitrate is 2.86 mgO/mgNO3-N. If oxygen and/or nitrate are recycled to the anaerobic reactor, the OHOs no longer convert fermentable RBO to VFAs but instead themselves utilize it for energy and growth with the oxygen or nitrate as external electron acceptor. For every 1
Biological Nutrient Removal
0.02
13.5
9.0
0.05
fxana
0.01
4.5
0.0
0
0.1
0.2
0.3
0.4
0.5
22.5
Influent P
0.02
15.0 0.05
fxana 0.01
0.6
7.5
0.0
0.0 0
Influent RBO fraction (fSb’s)
(a)
0.03
0.00
0.0
0.00
30.0
0.25 0.20 0.15 0.10
Raw WW
P removal (mgP I−1)
Influent P
0.03
0.04
18.0
0.25 0.20 0.15 0.10
P removal (mgP/mg Infl COD)
Settled WW
P removal (mgP I−1)
P removal (mgP/mg Infl COD)
0.04
517
0.1
0.2
0.3
0.4
Influent RBO fraction (fSb’s)
(b)
Figure 58 Predicted P removal vs. readily biodegradable COD (RBCOD, Sbsi) as a fraction of the biodegradable COD (Sbi) fSb’s ¼ Sbsi/Sbi) for various anaerobic mass fractions (fxana) for a two-compartment anaerobic reactor BEPR system at 20 days sludge age, treating the example raw (a) and settled (b) wastewater.
0.25 0.20 0.15 0.10
0.02
22.5 Influent P
15.0
0.05
fxana
0.01
7.5
0.0
0 ( a)
8 2 4 6 Nitrate in recycle (NO3−N I−1)
0.25 0.20 0.15 0.10
0.03
Influent P
13.5
9.0
0.02 0.05
fxana 4.5
0.01 0.0
0.00
0.0
0.00
18.0 Settled WW
10
0.0 0
(b)
P removal (mgP I−1)
0.03
P removal (mgP/mg Infl COD)
Raw WW P removal (mgP I−1)
P removal (mgP/mg Infl COD)
0.04
30.0
0.04
2 4 6 8 Nitrate in recycle (NO3−N I−1)
10
Figure 59 Predicted P removal vs. nitrate concentration in recycle to anaerobic (recycle 1:1) for various anaerobic mass fractions (fxana) for a twocompartment anaerobic reactor BEPR system at 20 days sludge age treating the example raw (a) and settled (b) wastewater.
mgO2 and 1 mgNO3-N recycled to the anaerobic reactor, 3.0 and 8.6 mgCOD, respectively, are utilized (Equation (175)). Consequently, allowing oxygen and/or nitrate to enter the anaerobic reactor reduces the flux of VFAs available to the PAOs for storage, and correspondingly reduces the P release, P uptake, and P removal. From Figures 59(a) and 59(b), when the nitrate concentration in the recycle exceeds about 12 mgN l1, the P removal decreases to 4 (raw) and 2.2 mgP l1 (settled) which is the same as that of an ND system (fxana ¼ 0) with zero BEPR. In this situation, all the influent RBO is denitrified by the OHOs with the result that no VFAs are released and no VFAs are available to the PAOs, and BEPR no longer takes place – the P removal obtained is due to wastage of sludge with normal metabolic P content (0.03 mgP/mgVSS). If the influent RBO concentration increases or decreases, the concentration of recycled nitrate that completely consumes the RBO will increase or decrease correspondingly below about 12 mgN l1 (provided the recycle ratio remains unchanged). Clearly, one of the principal orientations in any design for BEPR is to minimize oxygen entrainment and nitrate recycling to the anaerobic reactor. To achieve this in situations where nitrification is obligatory or unaviodable, a number of different system configurations have been developed (Figure 47).
4.14.33.5 Subdivision of the Anaerobic Reactor into Compartments The effect of subdividing the anaerobic reactor into compartments is shown in Figures 60(a) (raw) and 60(b) (settled). Increasing the anaerobic reactor from a single completely mixed one to two compartments in series significantly improves the P removal, but increasing the number of compartments to greater than 3 yields little additional increase. This increase is due to the increased fermentable RBO conversion with in-series anaerobic reactor operation as a result of the first-order nature of the conversion kinetics. For design, at least two equal-sized in-series anaerobic reactors should be used.
4.14.34 Denitrification in NDBEPR Systems 4.14.34.1 Introduction Because usually N removal is also a requirement for BNR systems, nitrification is included and hence also denitrification to benefit from its advantages (Section 4.14.24). In the steady mixed culture BEPR model, the nitrate recycled to the anaerobic reactor needs to be known considering the adverse
Biological Nutrient Removal
0.25 0.20 0.15
0.03
0.02
0.10 0.05
22.5
15.0
Influent P
fxana
0.01
7.5
0.0
0.00
0.0 1
(a)
0.04
2
4 3 No anaerobic compartments
Settled WW
18.0
0.25 0.20
0.03
13.5 0.15 0.10
0.02
Influent P
9.0
0.05
fxana
0.01
4.5
0.0
0.00
5
0.0 1
(b)
P removal (mgP I−1)
30.0 Raw WW
P removal (mgP/mg Infl COD)
P removal (mgP/mg Infl COD)
0.04
P removal (mgP I−1)
518
2
4 3 No anaerobic compartments
5
Figure 60 Predicted P removal vs. the number of compartments in the anaerobic reactor for various anaerobic mass fractions (fxana) in a BEPR system at 20 days sludge age treating the example raw (a) and seattled (b) wastewater.
influence of recycling nitrate to the anaerobic reactor on P removal. Indeed, one of the principal orientations in the design procedure for P removal is to prevent nitrate recycling. Where nitrification is not required, this can be suppressed in a simple configuration such as the Phoredox (or the A/O) system (Figure 47(a)) but this option may not viable in some countries because nitrification is either obligatory or unavoidable due to high wastewater temperature. Accordingly, reliable and accurate quantification of denitrification in NDBEPR systems is essential not only for security in P removal but also for estimating the N removal by the NDBEPR system. The early approach (1976–85) to quantify denitrification in NDBEPR systems was to use the theory and procedures for ND systems, as set out in Sections 4.14.17–4.14.27 (Nicholls, 1982; Ekama et al., 1983; WRC, 1984). Experimental data indicated that this approach appeared to predict the observed denitrification quite closely. However, from the mechanisms for BEPR, which emerged later, this approach was theoretically inconsistent. The influent RBO was apparently used twice – first in the anaerobic reactor where it is converted to VFAs which are taken up and stored as PHA by the PAOs, and again in the primary anoxic reactor for denitrification via the K1 rate (Section 4.14.25.1). This situation is possible only if the PAOs denitrified significantly using most of their internally stored PHA with nitrate as electron acceptor in the anoxic reactor. This implies that the most of the P uptake should take place in the primary anoxic reactor. However, this was not usually observed. Practically, all (490%) the P uptake took place in the aerobic reactor of the UCT NDBEPR systems operated during the 1980s (Wentzel et al., 1985, 1990).
4.14.34.2 Experimental Basis for Denitrification Kinetics in NDBEPR Systems Clayton et al. (1991) undertook an experimental investigation into the denitrification kinetic behavior in mixed culture NDBEPR systems. A laboratory-scale modified UCT system (Figure 47(e)) was set up and operated for a year and a half. For the first 6 months, the first primary anoxic reactor was a plugflow reactor, thereafter a completely mixed one. The response of the system was monitored daily and profiles on the
plugflow primary anoxic reactor measured periodically. In addition, a variety of anoxic batch tests that reproduced the conditions in primary and secondary anoxic reactors were conducted on mixed liquor harvested from the MUCT system. In the plugflow reactor and batch tests, all the important parameters were measured to delineate the behavior of the OHOs and PAOs. No differences in the concentration time profiles from the plugflow reactor and batch tests were noted. From these tests: (1) Under the steady-state conditions of the MUCT system, the general denitrification formulation for ND systems dNO3/ dt ¼ KXBH applied also to NDBEPR systems. (2) In the primary anoxic reactor, (1) the rapid rate of denitrification associated with RBO in ND systems (K0 1, Figure 35; the K0 rate here for NDBEPR systems is used to distinguish it from the K rate in ND systems) was usually absent or of very short duration, (2) the slower rate of denitrification associated with BPO (K0 2) continued over the entire duration of the plugflow retention time or batch test (as in ND systems) but its rate was approximately 2 1/2 times faster than in the primary anoxic reactor of ND systems, i.e., 0.224 mgNO3-N/(mgOHOVSS d), where the OHOVSS concentration was calculated from the experimental system data with the ND model, not the BEPR model, that is, ignoring the reduction in OHOVSS due to the presence of the PAOs (Clayton et al., 1991). Based on the BEPR model, which yields lower OHOVSS due to the presence of PAOs, the K0 2 rate would be even higher compared with ND systems (Table 17). (3) In the secondary anoxic reactor, the denitrification rate (K0 3) was approximately 1 1/2 times the rate measured in secondary anoxic reactors of ND systems (K3, Figure 35(b)), also based on the ND model (Table 17). Clayton et al. (1991) proposed three possible explanations for the increased denitrification rate constant K0 2 observed in the primary anoxic zone of NDBEPR systems: 1. PAOs can denitrify, thereby contributing to the denitrification rate by utilizing the intracellular PHB acquired in the anaerobic zone. 2. PAOs cannot denitrify and the influent BPO is modified in the anaerobic zone to a more readily hydrolyzable form, thereby inducing a faster denitrification rate by the OHO in the primary anoxic reactor.
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519
Table 17 Specific denitrification rates (K) at 20 1C observed by Clayton et al. (1991) in the MUCT NDBEPR system based on the WRC (1984) and Wentzel et al. (1990) models compared with those in ND systems (WRC, 1984) System
NDBEPR systems based on ND model
NDBEPR systems based on BEPR model
ND systems based on ND model
Units
mgNO3–N/(mgVSS d)
mgNO3–N/(mgAHVSS d)
mgNO3-N/(mgAHVSS d)
Primary anoxic
K10 ¼ 0.61a K20 ¼ 0.224 K30 ¼ 0.100
K10 ¼ 0.70a K20 ¼ 0.255 K30 ¼ 0.114
K1 ¼ 0.720 K2 ¼ 0.101b K3 ¼ 0.072
Secondary anoxic a
Denitrification by this rate contributes negligibly to the N removal of the NDBEPR system. In single-reactor intermittent aeration ND systems Warburton et al. (1991) obtained K2 ¼ 0.128 at 20 1C.
b
Denitrification in N removal systems Primary anoxic reactor
Aerobic reactor
Denitrification in BioP removal systems
Secondary anoxic Reaeration reactor reactor
Mixed liquor recycle a Influent
Waste flow Settler Effluent Sludge recycle s
Influent RBCOD + SBCOD
K1 + K2 K2
Influent SBCOD only
K3
Endogenous SBCOD only
Anaerobic Anoxic reactor reactors
Aerobic reactor
Mixed liquor recycle Influent
Secondary anoxic reactor Waste flow Settler
a
r
Effluent RBCOD acidified and taken up by PAOs No initial rapid K1 rate
Sludge recycle s
K 2′
Influent SBCOD only
K 3′
Endogenous SBCOD only
Figure 61 Comparison of steady-state specific denitrification rates (K, mgNO3-N/(mgOHOVSS d)) in the primary and secondary anoxic reactors of ND (a) and NDBEPR (b) systems. The rates are compared in Table 15.
3. PAOs cannot denitrify and the BPO is not modified in the anaerobic zone but a higher rate of BPO hydrolysis/utilization is stimulated in the OHOs in NDBEPR systems by the anaerobic–anoxic–aerobic sequencing. The PHB concentrations measured in the (1) anaerobic, anoxic, and aerobic zones of the MUCT parent system; (2) anoxic batch tests on mixed liquor harvested from the MUCT system; and (3) anoxic batch tests on mixed liquor harvested from the enhanced PAO cultures of Wentzel et al. (1988, 1989a) demonstrated that PHB did not serve as a substrate source for denitrification (negligible decrease). Therefore, the PAOs did not contribute significantly to the K’2 denitrification rate in the primary anoxic reactor and so cause (1) had to be rejected. This conclusion was supported from the observation that in the mixed and enhanced culture systems and the batch tests, the P uptake was predominantly (490%) aerobic – negligible anoxic P uptake was observed. If the anaerobic reactor pretreats the influent BPO to a more readily utilizable form, then the K denitrification rates should be lower when the NDBEPR sludge is mixed with influent wastewater. Batch tests on sludge from the MUCT system fed the same wastewater as the parent system, yielded the same high K0 2 denitrification rates. Therefore, the anaerobic reactor did not modify the BPO to a more utilizable form. So cause (2) was rejected and default cause (3) had to be accepted. No experimental means was devised to test this third
cause. However, it did at least provide a consistent explanation also for the higher K0 3 in the secondary anoxic reactor – causes (1) and (2) explain only a higher K0 2 rate. A comparison between the denitrification rates in the ND and NDBEPR systems is shown in Figure 61. Because the PAOs did not significantly contribute to the denitrification, the K0 rates had to be recalculated so that the denitrification process in NDBEPR systems is correctly ascribed to the OHO group performing it. The proportion of OHOs in the VSS of NDBEPR systems is smaller than in ND systems (Figure 55) because the PAOs obtain most of the influent RBO. Ekama and Wentzel (1999a) calculated the OHO fraction (favOHO) for NDBEPR systems iteratively with the aid of the steady-state BEPR model (Section 4.14.31) using the measured value for the influent RBO fraction and varying the influent UPOs fraction (UPOCOD, fS’up) until the calculated system VSS mass, now comprising active and endogenous PAO and OHO components and unbiodegradable particulate VSS from the influent, matched that measured. When the correct fS’up had been found, the calculated P removal was matched to that measured by changing the PAO P content (fXBGP) from the enhanced PAO culture value of 0.38 mgP/mgPAOVSS. For MUCT system of Clayton et al. (1991), the recalculated average VSS fractionation results are fS’up ¼ 0.15, favOHO ¼ 0.21, and fXBGP ¼ 0.388 mgP/mgPAOVSS. Because the favOHO based on ND model (Section 4.14.9.4) was 0.24, on average the K0 rates were 0.24/0.21 ¼1.14 or 14%
520
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higher (Table 17). Following this calculation procedure, Ekama and Wentzel (1999a) also calculated the fS’up, favOHO, fXBGP, and K0 2 for four other UCT investigations, viz., Musvoto et al. (1994), Pilson et al. (1995), Sneyders et al. (1997), and Mellin et al. (1997). For the same wastewater source, reasonably consistent fS’up values are expected. For fully aerobic and ND systems, this has been the case in the UCT laboratory. For the Mitchells Plain unsettled wastewater usually fed to the experimental systems, this value was found to be around 0.12 for widely differing aerobic and ND systems, e.g., 0.10870.052 for aerobic systems (Mbewe et al., 1994), and 0.13570.060 (Warburton et al., 1991) and 0.1270.04 (Ubisi et al., 1997a, b) for anoxicaerobic systems. However, for NDBEPR systems, this was not the case. Not only was fS’up higher for NDBEPR systems fed the Mitchells Plain unsettled wastewater, it also varied widely in the different NDBEPR systems, from 0.0470.055 (Sneyders et al., 1997) to 0.29370.063 (Musvoto et al., 1994). Because of the method calculating the fS’up fraction, by reconciling the calculated VSS mass with the measured VSS mass, the variation in fS’up changes the favOHO. This, in turn, affects the K0 2 and K0 3, rates, which are higher for higher fS’up and lower for lower fS’up (Ekama and Wentzel, 1999b). Clearly, there are factors that affect the sludge production per unit COD load in the NDBEPR system that the models do not recognize. Two such factors appear to be the unaerated sludge mass fraction (fxt) and sludge settleability (measured as diluted sludge volume index, DSVI). The higher the fxt, the higher the fS’up, which could be due to an accumulation of undegraded BPO in the system. If this were the only factor, then the method of calculating fS’up and favOHO would be acceptable because undegraded BPO in effect is unbiodegradable particulate organics. However, this is not the only factor because systems with the same fxt yielded different fS’up and favOHO values depending on the DSVI (Musvoto et al., 1994; Casey et al., 1994a, 1994b). As the DSVI and hence AA (low F/M) filament abundance increased, so the system VSS mass decreased and vice versa. The calculated K0 rates varied accordingly, decreasing as the system VSS mass increased and vice versa. No explanation for this variation with DSVI can be advanced. The NDBEPR models, both steady-state (e.g., Wentzel et al., 1990) and dynamic simulation (e.g., ASM2, Henze et al., 1995), are extensions of their predecessors (WRC, 1984, ASM1; Henze et al., 1987) by including the kinetics of BEPR. Relatively few interactions between the ND and BEPR processes take place in these models, the main ones being that (1) the VFAs for the PAOs are generated by the OHOs in the anaerobic zone from the influent RBO, and more importantly, (2) the reduction factor for the BPO hydrolysis/utilization rate, Z, is increased from 0.33 in ASM1 (and UCTOLD, Dold et al., 1991) to 0.60 in ASM2 (and UCTPHO, Wentzel et al., 1992) to account for the increased K0 2 and K0 3 rates observed by Clayton et al. (1991). Insofar as the BEPR kinetics in the ASM2 and UCTPHO models are concerned, P release and uptake occur exclusively in the anaerobic and aerobic reactors respectively, in conformity with the observations of Siebritz et al. (1983), Wentzel et al. (1985, 1989b, 1990), Clayton et al. (1991), and Sneyders et al. (1997). Therefore, for the last two mentioned investigations, given the correct input fS’up and Z values, the NDBEPR models will satisfactorily predict the
performance of the M/UCT systems. However, in three other investigations, viz., Musvoto et al. (1994), Pilson et al. (1995), and Mellin et al. (1997), the P release, P uptake, and P removal behavior were significantly different to that observed on which the models are based. Not only was the excess P removal depressed at about 60% of that expected from the model of Wentzel et al. (1990), but also the P release to removal ratio was decreased. With the depressed P removal, significant P uptake took place in the (second) anoxic reactors of the MUCT systems. This was confirmed in the anoxic batch tests; whereas in the tests of Clayton et al. (1991) and Sneyders et al. (1997) negligible anoxic P uptake took place, in those of Mellin et al. (1997) significant (B40%) P uptake took place. The significant decrease in BEPR with anoxic P uptake BEPR was subsequently confirmed by Vermande et al. (2002) in parallel aerobic P uptake BEPR and anoxic P uptake BEPR systems. It is possible that different species of PAOs find a niche in the systems that can accomplish anoxic P uptake, but which have lower RBCOD to P release, P release to P removal, and fXBGP ratios. Biochemical assays have indicated that some PAOs can denitrify (Lo¨tter, 1985; Lo¨tter et al., 1986) and even anaerobic–anoxic (no aerobic) BEPR systems have been operated successfully (Kuba et al., 1993). Also, in several other studies significant anoxic P uptake has been observed (e.g., Vlekke et al., 1988; Kerrn-Jespersen and Henze, 1993; Bortone et al., 1996; Kuba et al., 1996; Kuba and van Loosdrecht, 1996; Hu et al., 2000, 2007a, 2007b). Denitrification by PAOs is included in the biochemical model of Wentzel et al. (1986, 1991) but is not included in ASM2 (Henze et al., 1995) and UCTPHO (Wentzel et al., 1992) kinetic models. Anoxic P uptake behavior of PAOs has been included in ASM2d (Henze et al., 1999) but it merely allows the P uptake to commence in the anoxic reactor without changing the P uptake, that is, anoxic P uptake is modeled with the same stoichiometry and kinetics as aerobic P uptake, which clearly is not observed experimentally. Realistic anoxic P uptake behavior of PAOs therefore cannot be simulated with current suite of IWA ASM models. Proposals to include PAO denitrification have been made (e.g., Mino et al., 1995; Barker and Dold, 1997; Hu et al., 2007a, 2007b), with varying success. One of the main problems with modeling anoxic P-uptake BEPR is that the triggers that stimulate it are not well understood. Hu et al. (2002) conclude that anoxic P-uptake BEPR is undesirable in NDBEPR systems due to the reduction in P removal per influent RBO it causes. For maximum BEPR with (usually) limited influent RBO, aerobic P-uptake BEPR is required. Large aerobic mass fractions (fxto0.5) and underloaded primary anoxic reactors with nitrate appear to favor aerobic P-uptake BEPR.
4.14.34.3 Denitrification Potential in NDBEPR Systems The denitrification potential is the maximum amount of nitrate per liter influent flow that can be removed by biological means in the anoxic reactors. As the experimental investigation into denitrification kinetics in NDBEPR systems indicated that the formulation
dNO3 =dt ¼ KXBH
Biological Nutrient Removal
developed for ND systems can also be applied to NDBEPR systems, the techniques set out in Section 4.14.25.2 to develop equations for denitrification potential in ND systems can be followed for NDBEPR systems also. For development of these equations, the experimental observation that the PAOs do not denitrify is accepted. Denitrification in the primary anoxic reactor is via utilization of any RBO leaking through the anaerobic reactor, and BPO. Procedures to determine the amount of RBO leaking through the anaerobic reactor to the primary anoxic reactor were set out in Section 4.14.31.1.4, where SbsfN is the concentration of FRBO exiting the last anaerobic compartment. Hence, SbsfN (1 þ recycle ratio) is the concentration FRBO per liter influent flow exiting the anaerobic reactor and available for denitrification in the primary anoxic reactor by OHOs. Accordingly, the denitrification potential in the primary anoxic reactor (Dp1) can be expressed as
Dp1 ¼ SbsfN ð1 þ rÞð1 f cv YH Þ=2:86 þ K02 XBH Rnp
ð205Þ
Following the procedures set out in Section 4.14.26.3, Equation (205) can be modified and simplified to give
Dp1 ¼ SbsfN ð1 þ rÞð1 f cv YH Þ=2:86 þ f x1 K02T ðSbOHO ÞYH Rs =ð1 þ bHT Rs Þ ðmg NO3 -N l1 influentÞ ¼ a0 þ f x1 K02T b0
ð206Þ
where fx1 is the primary anoxic sludge mass fraction and a0 ¼ SbsfN ð1 þ rÞð1 f cv YH Þ=2:86 and b0 ¼ ðSbOHO ÞY H Rs = ð1 þ bHT Rs Þ. In Equation (206), it is assumed that the initial rapid rate of denitrification (K0 1T) on FRBO leaking through the anaerobic reactor, SbsfN(1 þ r) is always complete, that is, the actual retention time in the primary anoxic reactor is longer than the time required to utilize this usually low concentration FRBO. As with ND systems, an equation can be developed to determine the minimum primary anoxic mass fraction f’x1min to deplete this RBO. This minimum will be a very low value (o0.05) which is much smaller than the primary anoxic reactors in NDBEPR systems, so generally Equation (206) is valid. However, Equation (206) is not without complication. To calculate the primary anoxic denitrification potential (Dp1), the concentration of RBO in the outflow from the anaerobic reactor (SbsfN) is required. To calculate SbsfN, the concentration of nitrate recycled to the anaerobic reactor is required which in turn requires Dp1 to be known. This problem is overcome by assuming initially that the nitrate concentration exiting the primary anoxic reactor is zero, as was done for the primary anoxic reactor of the MLE system in Section 4.14.25.2, which for the UCT systems means zero nitrate discharge to the anaerobic reactor. For brevity, other NDBEPR configurations are not considered in this chapter. However, if required, the denitrification potential of the secondary anoxic reactor is found using the principles set out in Section 4.14.25.2, viz.,
Dp3 ¼ f x3 K03T ðSbOHO Þ YH Rs =ð1 þ bHT Rs Þ ¼ f x3 K03T b0
ðmgNO3 -N l1 influentÞ
where fx3 is the secondary anoxic sludge mass fraction.
ð207Þ
521
Equation (207) applies to secondary anoxic reactors situated both in the mainstream (e.g., five-stage Bardenpho) and in the underflow recycle (e.g., JHB system). However, in applying Equation (207) to secondary anoxic reactors situated in the underflow recycle, care must be taken in evaluating fx3, because the mixed liquor concentration is increased by a factor (1 þ s)/s in the underflow anoxic reactor compared with the mainstream reactors.
4.14.34.4 Principles of Denitrification Design Procedures for NDBEPR Systems In NDBEPR systems design is oriented to achieve in a single sludge system COD removal, N removal (ND), and P removal (BEPR). Conflict between the last two objectives may arise, for example, the proportion of the total unaerated sludge mass assigned to the anoxic reactor(s) (N removal) and the anaerobic reactor (P removal). For each design, the priorities for treatment need to be assessed and a compromise reached to optimize the system for the particular effluent quality required. Because P is the element that is the main driver for eutrophication, for most designs of NDBEPR systems the focus is on BEPR with denitrification as a secondary design priority. Accordingly, the principle in denitrification design for NDBEPR systems is to ensure that the anaerobic reactor is protected from recycling of nitrate, which causes a disproportionate decrease in the magnitude of P removal (Figure 59). This principle guides selection of the system configuration (five-stage modified Bardenpho, JHB, and M/ UCT; Figure 47) and provides a starting point for sizing the anoxic reactors. When selecting a system configuration for BEPR, it is necessary to establish whether complete denitrification can be achieved. For the wastewater characteristics (i.e., influent TKN and COD concentrations (Nti and Sti)), maximum specific growth rate of the nitrifiers at 20 1C (mAm20), and the average minimum water temperature, the maximum unaerated sludge mass fraction (fxm) and the nitrification capacity (Nc) can be calculated for a selected sludge age (Rs) (Section 4.14.20.3). This fxm needs to be divided between anaerobic (for BEPR) and anoxic (for denitrification) mass fractions. Consequently, the maximum anoxic sludge mass fraction (fxdm) is the difference between the maximum unaerated mass fraction (fxm) and the selected anaerobic sludge mass fraction (fxa), that is,
f xdm ¼ f xm f xa
ð208Þ
where fxm is given by Equation (136) for a selected Rs, mAm20, Sf, and Tmin. The fxdm then can be subdivided between primary and secondary anoxic sludge mass fractions (fx1 and fx3) and this division fixes the denitrification potential of these two reactors (Dp1 and Dp3) and hence also of the system. If the denitrification potential of the system exceeds the nitrification capacity (i.e., Dp1 þ Dp34Nc), then complete denitrification is possible and the secondary anoxic reactor can be situated in the mainstream, that is, a five-stage Bardenpho system can be selected. If complete denitrification is not possible, then depending on the magnitude of the effluent nitrate
522
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concentration, the underflow (s) recycle cannot be discharged directly to the anaerobic reactor. If the nitrate concentration is low (o3 mgN l1), the secondary anoxic sludge mass fraction (fx3) can be combined with the primary anoxic sludge mass fraction to form a three-stage Bardenpho, which, with a higher a recycle ratio, may produce a lower nitrate concentration in the sludge underflow (and effluent) than the fivestage Bardenpho because the K0 2 rate is higher than the K0 3 (Table 17). If the nitrate concentration is high (43 mgN l1), the secondary anoxic reactor can be moved into underflow recycle to form the JHB system, in which event the denitrification potential of the secondary anoxic reactor (Dp3) must exceed the nitrate and oxygen loads via the underflow s recycle. If this requirement is not met, nitrate will leak through the underflow secondary anoxic reactor to the anaerobic reactor. In this event, since the denitrification potential of the primary anoxic reactor (Dp1) is greater than that of the secondary anoxic reactor (Dp3) for equal anoxic mass fractions, incorporation of a secondary anoxic reactor becomes an inefficient utilization of anoxic mass fraction, and the secondary anoxic mass fraction is added to the primary anoxic reactor, the underflow s recycle needs to be denitrified in the primary anoxic reactor to form the M/UCT system. With each change of configuration, more nitrogen removal is sacrificed (i.e., the effluent nitrate concentration increases) to protect maximum BEPR, that is, zero or very low nitrate discharge to the anaerobic reactor.
4.14.34.5 Analysis of Denitrification in NDBEPR Systems Analysis of the denitrification behavior in the NDBEPR system is essentially the same as for the ND system (Section 4.14.26.3) except: (1) The maximum anoxic mass fraction for denitrification (fxdm) for the NDBEPR system is given by Equation (208), whereas fxdm for the ND system is given by Equation (166). Hence, for the same maximum unaerated sludge mass fraction (fxm), the NDBEPR system has a lower fxdm than the ND system, by an amount equal to fxa. (2) The specific denitrification rates for ND systems (K2 and K3) are substituted with the rates for NDBEPR systems (K0 2 and K0 3, Table 17). (3) The denitrification potentials for the primary and secondary anoxic reactors are modified from Equations (163) and (164) for ND systems to Equations (206) and (207) for the NDBEPR system to take account of the uptake of COD by the PAOs in the anaerobic reactor, the zero denitrification by the PAOs, and the faster OHO denitrification rates in NDBEPR systems. Taking account of the above, denitrification equations can be developed for all the NDBEPR configurations (Figure 47). However, in the interests of brevity, only the UCT configuration will be considered.
4.14.35 Denitrification in the UCT System In the UCT system the denitrification behavior is very similar to that in the MLE system, because the a and s recycles discharge into the primary anoxic reactor, so that taking due account of the effect of incorporating the anaerobic reactor, the design equations and procedures developed for the MLE system can be applied to the UCT system. Since complete
denitrification is not possible in the UCT system (high effluent nitrate concentration), the entire anoxic mass fraction (fxdm) available is used as a primary anoxic reactor (fx1). As in the MLE system, the a and s recycle ratios determine the split of the nitrate generated in the aerobic reactor (nitrification capacity, Nc) between the primary anoxic reactor and the effluent. The a recycle ratio is selected so that the equivalent nitrate load on the primary anoxic reactor via the a and s recycles is equal to its denitrification potential (Dp1). For a selected s recycle ratio, the a recycle ratio that loads the primary anoxic reactor via to its denitrification potential is the optimum a recycle ratio (aopt). The denitrification potential of the (primary) anoxic reactor (Dp1) is found from Equation (206) with fx1 ¼ fxdm. Following the same reasoning as in Section 4.14.26.3, the optimum a recycle ratio (aopt) is given by Equation (169), with the proviso that the Nc and Dp1 are applicable to NDBEPR systems, that is, Nc is lower due to the higher sludge production (Section 4.14.31.5.2) and Dp1 is based on Equation (206) with K0 2. As for the MLE system, at a ¼ aopt, Equation (170) gives the minimum effluent nitrate concentration (Nne) achievable. Equation (170) is valid for all aoaopt because for all aoaopt the assumption on which Equation (169) is based is valid, that is, zero nitrate concentration exiting the primary anoxic reactor. If the system is operated with a4aopt, the equivalent nitrate load on the primary anoxic reactor via the a and s recycles exceeds the denitrification potential and nitrate will also be recycled via the r recycle to the anaerobic reactor, to the detriment of BEPR. Furthermore, if nitrate does leak through the primary anoxic reactor, then the nitrate concentration in the outflow from the primary anoxic reactor no longer is zero, and consequently, Equation (170) for the effluent nitrate concentration (Nne) is no longer valid. Equations for the effluent nitrate concentration for a4aopt can be derived by following the principles applied above for aoaopt, but are not considered because aoaopt is required for zero discharge of nitrate to the anaerobic reactor and maximum BEPR. If Equation (169) yields aopt ¼ 0, then the equivalent nitrate load via the s recycle is sufficient to match the denitrification potential of the primary anoxic reactor; if aopto0, the equivalent nitrate load via the s recycle exceeds Dp1 and nitrate will be recycled via the r recycle to the anaerobic reactor. The implication of this is that the Nc that gives aopt ¼ 0 represents the upper limit (equivalently the maximum influent TKN/ COD concentration ratio) that the UCT system is able to treat and still protect the anaerobic reactor against nitrate entry. All Nc (equivalently influent TKN/COD ratios) above this limit will result in nitrate recycle to the anaerobic reactor, which cannot be controlled in the UCT system (a ¼ 0) except by reducing the s recycle ratio. From the above, the minimum a recycle ratio is a ¼ 0. The maximum a recycle ratio (amax) is determined by some practical upper limit (aprac), usually in the range 5–6, beyond which the higher pumping costs outweigh the small gain in lower effluent nitrate concentration (see Section 4.14.26.3). However, for oxidation ditch type systems, or for systems with ‘‘through the wall’’ a recycles via low head high volume pumps, the a recycle ratio ( ¼ aopt) may be significantly higher than the aprac of 5–6. If aopt 4 aprac and aprac is selected, then the
Biological Nutrient Removal
primary anoxic reactor is oversized. This unused denitrification potential (Figure 38) can be kept (i.e., fx1 not decreased) as a factor of safety (for uncertainty if K0 rate) or the size of the fx1 of primary anoxic reactor reduced to match the its denitrification potential (Dp1) to the equivalent nitrate load, as was done for the balanced MLE system (Section 4.14.26.3.2), which will allow a reduction in sludge age. The procedure for the balance MLE sytem can be followed to determine the new sludge age. All the aspects discussed in Sections 4.14.11 and 4.14.14– 4.14.16 regarding reactor concentration selection, system design and control, selection of sludge age, and treatment of the primary and/or secondary sludge produced also apply to NDBEPR systems and should be referred to there.
4.14.36 Conclusion The ND and NDBEPR models such as IWA ASM1 and 2 (Henze et al., 1987, 1995), UCTOLD (Dold et al., 1991), and UCTPHO (Wentzel et al., 1992) are very helpful for biological nutrient removal process description and simulation. However, models always need to be used with great circumspection and experience of real systems. It would appear that the ND models such as ASM1 and UCTOLD give an acceptably reliable description of the ND AS systems – the model predictions compare favorably with observed results and the wastewater characteristic, stoichiometric, and kinetic constants in the models to achieve this are reasonably consistent. For these models some scientific maturity is apparent, where the default kinetic and stoichiometric constants predict the performance of an ND system with acceptable risk of deviation. For the NDBEPR models, this is not the case. The experiments described in the literature point to three important observations in real NDBEPR systems not recognized in NDBEPR models that model users need to be aware of for prudent and proper application: that is, (1) the large variation in the unbiodegradable particulate COD fraction (fS’up) and hence the OHO active fraction (favOHO) and denitrification rate (K0 2); (2) the large variation in biological P removal behavior and P content of PAOs (fXBGP) with anoxic P-uptake BEPR stimulated in some systems for reasons not well defined yet; and (3) the unaccounted loss of influent COD in NDBEPR systems, in that even in carefully controlled laboratory systems, only 75–85% of the influent COD can be recovered in a COD mass balance (Ekama et al., 1999a,b).
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Pitman AR, Vandalsen L, and Trim BC (1988) Operating experience with biological nutrient removal at the Johannesburg Bushkoppie works. Water Science and Technology 20(4–5): 51--62. Pitman AR, Venter SLV, and Nicholls HA (1983) Practical experience with biological phosphorus removal plants in Johannesburg. Water Science and Technology 15(3/4): 233--259. Poinapen J and Ekama GA (2010) Biological sulphate reduction using primary sewage sludge in a upflow anaerobic sludge bed reactor–Part 5: Development of a steady state model. Water SA 36(2): 193--202. Randall AA, Benefield LD, and Hill WE (1994) The effect of fermentation products on enhanced biological phosphorus removal, polyphosphate storage, and microbial population dynamics. Water Science and Technology 30(6): 213--219. Randall CW, Marshall DW, and King PH (1970) Phosphate release in activated sludge process. Journal of the Sanitary Engineering Division, ASCE 96(SA2): 395--408. Randall EW, Wilkinson A, and Ekama GA (1991) An instrument for the direct determination of oxygen utilization rate. Water SA 17(1): 11--18. Rensink JH, Donker HJGW, and Vries HPD (1981) Biological P-removal in domestic wastewater by the activated sludge process. In: Proceedings of the 5th Europe Sewage and Refuse Symposium, Munich. European Water Association, Hennep, D53773, Germany or International Solid Waste Association (ISWA), Vienna, A-1080, Austria. Richard MG, Jenkins D, Hao O, and Shimizu G (1982) The isolation and characterization of filamentous micro-organisms from activated sludge bulking. Progress Report No. 81-2. Berkeley: SERL, University of California. Riddell MDR, Lee JS, and Wilson TE (1983) Method for estimating the capacity of an activated sludge plant. Journal of the Water Pollution Control Federation 55(4): 360--368. Samson KA and Ekama GA (2000) An assessment of sewage sludge stability with a specific oxygen utilization rate (SOUR) test method. Water Science and Technology 42(9): 37--40. Satoh H, Mino T, and Matsuo T (1992) Uptake of organic substrates and accumulation of polyhydroxyalkanoates linked with glycolysis of intracellular carbohydrates under anaerobic conditions in the biological excess phosphate removal processes. Water Science and Technology 26(5–6): 933--942. Saunders AM, Mabbett AN, McEwan AG, and Blackall LL (2007) Proton motive force generation from stored polymers for the uptake of acetate under anaerobic conditions. FEMS Microbiology Letters 274(2): 245--251. Sehayek L and Marais GVR (1981) Supplementary phosphorus removal by side-line addition of lime in the activated sludge process. Research Report W40. Rondebosch: Department of Civil Engineering, University of Cape Town. Sen D, Mitta P, and Randall CW (1994) Performance of fixed film media integrated in activated sludge reactors to enhanced nitrogen removal. Water Science and Technology 30(11): 13--24. Setter LR, Carpenter WT, and Winslow GC (1945) Practical application of modified sewage aeration. Sewage Works Journal 17(4): 669. Seviour RJ, Mino T, and Onuki M (2003) The microbiology of biological phosphorus removal in activated sludge systems. FEMS Microbiology Reviews 27(1): 99--127. Shapiro J, Levin GV, and Humberto HZ (1967) Anoxically induced release of phosphate in wastewater treatment. Journal of the Water Pollution Control Federation 39: 1810--1818. Siebritz IP, Ekama GA, and Marais GVR (1980) Excess biological phosphorus removal in the activated sludge process at warm temperature climate. In: Proceedings of the Waste Treatment and Utilization 2, pp. 233–251, Toronto: Pergamon. Siebritz IP, Ekama GA, and Marais GVR (1983) A parametric model for biological excess phosphorus removal. Water Science and Technology 15(3/4): 127--152. Simpkins MJ and McLaren AR (1978) Consistent biological phosphate and nitrate removal in an activated sludge plant. Progress in Water Technology 10(5/6): 433--442. Smolders GJF, van der Meij J, van Loosdrecht MCM, and Heijnen JJ (1994a) Stoichiometric model of the aerobic metabolism of the biological phosphorus removal process. Biotechnology and Bioengineering 44(7): 837--848. Smolders GJF, van der Meij J, van Loosdrecht MCM, and Heijnen JJ (1994b) Model of the anaerobic metabolism of the biological phosphorus removal process: Stoichiometry and pH influence. Biotechnology and Bioengineering 43: 461--470. Smolders GJF, Meij Jvd, Loosdrecht MCMv, and Heijnen JJ (1995) A structured metabolic model for anaerobic and aerobic stoichiometry and kinetics of the biological phosphorus removal process. Biotechnology and Bioengineering 47: 277--287. Sneyders MJ, Wentzel MC, and Ekama GA (1997) The effect of unstabilized landfill leachate addition on biological nutrient removal performance in activated sludge
4.15 Biofilms in Water and Wastewater Treatment Z Lewandowski, Montana State University, Bozeman, MT, USA JP Boltz, CH2M HILL, Inc., Tampa, FL, USA & 2011 Elsevier B.V. All rights reserved.
4.15.1 4.15.2 4.15.2.1 4.15.2.2 4.15.2.2.1 4.15.2.2.2 4.15.2.2.3 4.15.2.2.4 4.15.2.2.5 4.15.2.2.6 4.15.2.2.7 4.15.2.2.8 4.15.3 4.15.3.1 4.15.3.1.1 4.15.3.1.2 4.15.3.2 4.15.3.3 4.15.3.4 4.15.3.4.1 4.15.3.4.2 4.15.3.4.3 4.15.3.4.4 4.15.3.4.5 4.15.3.5 4.15.3.5.1 4.15.3.5.2 4.15.3.5.3 4.15.3.5.4 4.15.3.5.5 4.15.3.5.6 4.15.4 4.15.4.1 4.15.4.1.1 4.15.4.1.2 4.15.4.2 4.15.4.3 4.15.4.4 References
Introduction Part I: Biofilm Fundamentals Biofilm Formation and Propagation The Concepts of Biofilms and Biofilm Processes Quantifying microbial activity, hydrodynamics, and mass transport in biofilms Biofilm heterogeneity and its effects Biofilm activity Quantifying local biofilm activity and mass transport in biofilms from microscale measurements Horizontal variability in diffusivity and microbial activity in biofilms Mechanism of mass transfer near biofilm surfaces Biofilm processes at the macroscale and at the microscale Biofilms in conduits Part II: Biofilm Reactors Application of Biofilm Reactors Techniques for evaluating biofilm reactors Graphical procedure Empirical and Semi-Empirical Models Mathematical Biofilm Models for Practice and Research Biofilm Model Features Attachment and detachment process kinetics and rate coefficients Concentration gradients external to the biofilm surface and the mass transfer boundary layer Diffusivity coefficient for the rate-limiting substrate inside the biofilm Parameters: estimation and variable coefficients Calibration protocol Biofilm Reactors in Wastewater Treatment Biofilm reactor compartments Moving bed biofilm reactors Biologically active filters Expanded and fluidized bed biofilm reactors Rotating biological contactors Trickling filters Part III. Undesirable Biofilms: Examples of Biofilm-Related Problems in the Water and Wastewater Industries Biofilms on Metal Surfaces and MIC Differential aeration cells on iron surfaces SRB corrosion Biofilms on Concrete Surfaces: Crown Corrosion of Sewers Biofilms on Filtration Membranes in Drinking Water Treatment Biofilms on Filtration Membranes in Wastewater Treatment
4.15.1 Introduction Fundamental principles describing biofilms exist as a result of focused research. The use of reactors for the treatment of municipal wastewater is a common application of biofilms. Applied research exists that provides a basis for the mechanistic understanding of biofilm reactors. The empirical information derived from such applied research has been used to develop design criteria for biofilm reactors and remains the basis for biofilm reactor design despite the emergence of
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mathematical models as reliable tools for research and practice. Unfortunately, little information exists to bridge the gap between our current understanding of biofilm fundamentals and reactor-scale empirical information. Therefore, there is a clear dichotomy in literature: micro- (biofilm) and macro(reactor) scales. This chapter highlights the division. Part I is dedicated to basic research and communicating the state of the art with respect to understanding biofilms. Part II is practice oriented and describes the use of biofilms for the sanitation of municipal wastewater. A basis for addressing this
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disconnection is presented by (1) describing the fundamental biofilm principles that can be uniformly applied to biofilms in several disciplines extending from medicine to environmental biotechnology and (2) describing a fundamentalbased approach in order to understand and apply biofilms in reactors. The use of mathematical biofilm models is common in both research and practice, but only a cursory presentation of their mathematical description is presented here. Finally, Part III gives examples of undesirable biofilms in water and wastewater industries and describes the attempts to mitigate their effects. Metabolic reactions mediated by microorganisms residing in biofilms promote the biodeterioration of materials, including metals, concrete, and plastics. It is estimated that microbially influenced corrosion (MIC) alone costs the US economy billions of dollars every year.
4.15.2 Part I: Biofilm Fundamentals 4.15.2.1 Biofilm Formation and Propagation Biofilm formation is a process that consists of a sequence of steps. It begins with the adsorption of macromolecules (e.g., proteins, polysaccharides, nucleic acids, and humic acids) and smaller molecules (e.g., fatty acids, lipids, and pollutants such as polyaromatic hydrocarbons and polychlorinated biphenyls) onto surfaces. These adsorbed molecules form conditioning films which may have multiple effects, such as altering the physicochemical characteristics of the surface, acting as a concentrated nutrient source for microorganisms, suppressing or enhancing the release of toxic metal ions from the surface, detoxifying the bulk solution through the adsorption of inhibitory substances, supplying the nutrients and trace elements required for a biofilm, and triggering biofilm sloughing. Once the surface is prepared, cells begin to attach. The initial stages of biofilm formation are well documented, mostly because acquiring images of microorganisms at this stage of biofilm formation is relatively easy. The adherence of bacteria to a surface is followed by the production of slimy adhesive substances, extracellular polymeric substances (EPS). These are predominantly made of polysaccharides and proteins. Although the association of EPS with attached bacteria has been well documented in the literature, there is little evidence to suggest that EPS participates in the initial stages of adhesion. However, EPS definitely assists the formation of mature biofilms by forming a slimy substance called the biofilm matrix. Figure 1 shows the steps in the formation of mature biofilms. The existence of these three phases of biofilm development, as depicted in Figure 1, is generally acknowledged, although the terminology may vary among authors. For example, Notermans et al. (1991) called these phases: (1) adsorption, (2) consolidation, and (3) colonization. Once a mature biofilm has been established on a surface, it actively propagates and eventually covers the entire surface. The mechanisms of propagation in mature biofilms are more complex than those of initial attachment, and several of these mechanisms of biofilm propagation are depicted in Figure 2. Although biofilms can be seen with an unaided eye, imaging their structure, microbial community structure, and
Biofilm formation Attachment
Colonization
Growth
Bulk fluid
Surface Figure 1 Steps in biofilm formation. & 1995 Center for Biofilm Engineering, MSU-BOZEMAN.
distribution of EPS requires the use of several types of microscopy combined with various probes, such as fluorescent in situ hybridization (FISH) probes and fluorescent proteins (FPs) used as reporter genes. The favorite types of microscopy among biofilm researchers are those that allow the examination of living and fully hydrated biofilms. In addition, sophisticated image acquisition devices are often needed that can selectively stimulate and image various probes when more than one type of multicolored probe is used simultaneously. Using these techniques in conjunction with a suitable microscopy, biofilm researchers can detect the presence of the selected physiological groups of microorganisms in the biofilm, their position in the biofilm with respect to other microorganisms and surface, and even their physiological state – dead, injured, or alive. The in vitro FISH techniques, popular in medical diagnostics, require that DNA or RNA be isolated from the sample and separated on a gel, and that the fluorescent probes then be added to the sample. The in situ variety of the hybridization technique, which is extensively used in biofilm research, does not require isolating DNA or RNA prior to the use of the probes; instead, the probes are hybridized to the respective nucleotide sequences inside the cells (Biesterfeld et al., 2001; Delong et al., 1999; Ito et al., 2002; Jang et al., 2005; Manz et al., 1999). In situ hybridization uses fluorescence-labeled complementary DNA or RNA probes, often derived from fragments of DNA that have been isolated, purified, and amplified. In microbial ecology, ribosomal RNA in bacterial cells is targeted by fluorescencelabeled oligonucleotide probes. Figure 3 shows an image of manganese-oxidizing bacteria (MOB) Leptotrix discophora stained with a FISH probe (green) and counterstained with propidium iodide (red). Propidium iodide is a general stain which is quite popular with biofilm researchers (GrayMerod et al., 2005; McNamara et al., 2003; Nancharaiah et al., 2005). In mature biofilms, microorganisms are imbedded in the layer of EPS. Figure 4 shows an image of a mature biofilm acquired using scanning electron microscopy (SEM). It shows microbes embedded in a matrix of EPS attached to a surface, although the EPS in this image were reduced to an entangled network of dry strands because the sample had to be dehydrated before the biofilm was imaged using electron microscopy.
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Streaming Detaching
Seeding dispersal
Rippling
Rolling
Figure 2 Mechanisms of biofilm propagation (MSU-CBE, P.Dirx).
Figure 4 SEM image of a biofilm of Desulfovibrio desulfuricans G20 embedded in EPS (Beyenal et al., 2004).
Figure 3 L. discophora stained with FISH probes and counterstained with propidium iodide. Red indicates cells that were stained with propidium iodide, and green indicates cells that react positively to the fluorescent FISH probe. Yellow indicates green and red overlay. The scale bar is 20 mm (Campbell, 2003).
4.15.2.2 The Concepts of Biofilms and Biofilm Processes It is difficult to offer precise definitions of biofilms and biofilm processes that will satisfy everyone who is interested in studying biofilms and biofilm-based technologies. Several currently used definitions have roots in historical approaches to biofilm studies. These approaches initially referred to biofilms as physical objects – microbial deposits on surfaces – but later expanded the concept to consider biofilms as a mode of
microbial growth, an alternative to microbial growth in suspension. Life scientists often emphasize the definitions that refer to biofilms as a mode of microbial growth. Engineers often find that the definitions that refer to aggregates of microorganisms which are embedded in a matrix composed of microbially excreted EPS and attached to a surface are useful for their applications. Here, we will refer to biofilms as microorganisms and microbial deposits attached to surfaces. We will use the term biofilm processes in reference to all physical, chemical, and biological processes in biofilm systems that affect, or are affected by, the rate of biofilm deposition or the microbial activity in biofilms. Biofilm processes are carried out in biofilm reactors. Colloquially, the terms biofilm reactors and biofilm systems are used interchangeably. However, biofilm systems exist with or without human intervention, while biofilm reactors are produced by our actions.
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When we promote or suppress a biofilm process in a biofilm system, or even when we quantify a biofilm process in a biofilm system without affecting its rate, the biofilm system becomes a biofilm reactor. For example, wetlands can be natural or constructed. However, even natural wetlands become biofilm reactors once we start monitoring biofilm processes in them. We will use the term biofilm system to refer to a group of compartments and their components determining biofilm structure and activity. Biofilm systems are composed of four compartments:
• • • •
the the the the
surface to which the microorganisms are attached; biofilm (the microorganisms and the matrix); solution of nutrients; and gas phase (if present).
Each compartment of a biofilm system can have a number of components. The exact number of components in each compartment may vary, depending on the needs of a particular description. For example, for some analyses it may be convenient to identify two components of the biofilm: (1) the EPS (matrix) and (2) the microorganisms. In another study, it may be convenient to identify three components of the biofilm: (1) the EPS, (2) the microorganisms, and (3) the particular matter trapped in the matrix. Similarly, in some studies it may be convenient to single out two components of the surface – (1) the bulk material and (2) the biomineralized deposits – or, if MIC is studied, it may be convenient to describe the surface by identifying three components: (1) the metal substratum, (2) the corrosion products, and (3) the biomineralized deposits on the surface. The needs of the specific study or analysis dictate the number of components identified in each compartment of the biofilm system. Biofilm studies can be characterized as studies of the relations among the compartments, the properties of one or more compartments, or one or more components of a compartment. Among many factors that are used to quantify biofilm processes, biofilm activity is most often used. Biofilm reactors are often designed and operated to optimize biofilm activity, as are the biofilm reactors used for wastewater treatment discussed later in the text. Typically, biofilm activity is identified with the rate of utilization of the growth-limiting nutrient. In some instances, however, rates other than the rate of substrate utilization or biofilm accumulation are better descriptors of the system dynamics. For example, in studies of MIC, the rate of anodic dissolution of the metal affected by the process may be a more useful descriptor of biofilm activity than the rate at which the growth-limiting substrate is utilized. The choice of the process for evaluating biofilm activity is dictated by the nature of the study, and sometimes by analytical convenience. Monitoring the rate of biofilm accumulation is important in many applications, whether we want to enhance or inhibit the growth of biofilms. The methods employed include optical microscopy (Bakke and Olsson, 1986; Bakke et al., 2001), measuring light intensity reflected from microbially colonized surfaces (Bremer and Geesey, 1991; Cloete and Maluleke, 2005), collecting and analyzing images of biofilm depositions (Milferstedt et al., 2006; Pons et al., 2009), surface sensors based on piezoelectric devices (Nivens et al., 1993; Pereira
et al., 2008), and electrochemical sensors in which stainless steel electrodes change their electrochemical behavior as a result of biofilm deposition (Licina et al., 1992; Borenstein and Licina, 1994).
4.15.2.2.1 Quantifying microbial activity, hydrodynamics, and mass transport in biofilms Microbial activity (biofilm activity), hydrodynamics, and mass transport in biofilms are difficult to discuss separately as they affect each other in many ways. Biofilm activity at the microscale is quantified as the flux, from the bulk solution to the biofilm surface, of the substance selected for evaluating biofilm activity. Since fluxes at the microscale are quantified locally, rather than averaged over the entire surface area as is done when biofilm activity is evaluated at the macroscale, the concentration profiles of the selected substance must be measured with microsensors to assure adequate spatial resolution. The idealized model of hydrodynamics and mass transfer in biofilms shown in Figure 5 is a good starting point for a discussion of biofilm activity at the microscale. In this model the overall flow velocity in the main stream is considered to be the average flow velocity, Cb. This decreases toward the surface of the biofilm, as required by hydrodynamics, and reaches concentration Cs at the biofilm surface. The layer of liquid just above the biofilm surface, where the flow velocity decreases as a result of proximity to the surface, is the hydrodynamic boundary layer, and it is denoted by j. As the flow velocity decreases toward the biofilm surface, the mechanism of mass transport changes from being dominated by convection at locations away from the biofilm, where the flow velocity is high, to being dominated by diffusion at locations near the biofilm surface, where the flow velocity is low. As the microorganisms in the biofilm consume nutrients at the rate at which they are delivered and the mass transport becomes less efficient near the biofilm surface, the nutrient concentration decreases near the surface, forming a nutrient concentration profile within the hydrodynamic boundary layer. The layer of liquid above the biofilm surface where the nutrient concentration decreases is the mass transport boundary layer, and it is denoted by LL and RL is the mass transfer resistance external to the biofilm.
Substrate concentration profile
Flow velocity profile
N = k(Cb − Cs) k =
vb
DW 1 = RL LL Cb C LF
LL
ϕ Biofilm Substratum
Figure 5 Profiles of flow velocity and growth-limiting nutrient concentration near the surface of an idealized biofilm.
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4.15.2.2.2 Biofilm heterogeneity and its effects The term biofilm heterogeneity refers to the extent of the nonuniform distribution of any selected constituent in any of the compartments of the biofilm system, such as the distribution of the biomass, selected nutrients, selected products of microbial metabolism, or selected groups of microorganisms. Since there are many choices for the constituents selected to evaluate biofilm heterogeneity, the term biofilm heterogeneity is usually combined with an adjective referring to the selected constituent, such as structural heterogeneity, chemical heterogeneity, or physiological heterogeneity. The term biofilm heterogeneity was initially used exclusively to refer to the nonuniform distribution of the biomass in a biofilm. As time has passed, more types of heterogeneity have been described, and the term biofilm heterogeneity is not self-explanatory anymore: the specific feature of the biofilm with respect to which the heterogeneity is quantified needs to be specified. Quantifying biofilm heterogeneity is equivalent to quantifying the extent of nonuniform distributions, such as the distribution of biomass in the biofilm. Several tools from the statistical toolbox are available for evaluating the extent of nonuniform distribution; the most popular is the standard deviation. The procedure for estimating the heterogeneity of a selected constituent of a biofilm is identical with the procedure for evaluating the standard deviation of a set of experimental data with one important difference: the deviations from the average are not due to errors in measurement but reflect a feature of the biofilm – heterogeneity. One of the most profound effects of biofilm heterogeneity is that microscale measurements in biofilms deliver different results at different locations. This is an obvious concern as most models referring to microbial growth and activity have been developed for well-mixed reactors, in which the result of a measurement does not depend on the location. Figure 6 shows this effect: three very different profiles of carbon dioxide concentration were measured at three locations in a biofilm. Because of the biofilm heterogeneity, it is impossible to determine a representative location to make the local measurements of biofilm activity that are used to validate models of biofilm processes. To include the effects of biofilm heterogeneity in mathematical models of biofilm processes, the extent of these effects – the spatial variability of the features measured in biofilms – needs to be evaluated experimentally using tools that can take measurements in biofilms to a high spatial resolution. Such tools are routinely used in biofilm research in the form of microelectrodes and various types of microscopy, often enhanced with fluorescent probes. These types of measurements deliver information about selected locations in the biofilm, and their results are referred to as local properties. The most common such measurements are local biofilm activity, local mass transfer coefficient, local diffusivity, and local flow velocity. The definition of the local mass transport coefficient is derived from the measurement procedure: the coefficient of the mass transport of an electroactive species to the tip of an electrically polarized microelectrode. The local mass transport coefficient is measured using an amperometric microelectrode without a membrane operated at the limiting current condition (masstransfer-limited). Local diffusivity is computed from these
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A B C
5 4 3 2 1 0 0
100 200 300 400 Distance from the bottom (µm)
500
Figure 6 Carbon dioxide concentration profiles measured perpendicularly to the bottom (substratum) at three locations in a biofilm microcolony.
measurements by calibrating local mass transport microelectrodes in gels of known diffusivities (Beyenal et al., 1998).
4.15.2.2.3 Biofilm activity Biofilm activity in a biofilm reactor can be evaluated from the mass balance on the growth-limiting nutrient in the reactor:
Biofilm activity ¼
ðCInfluent CEffluent Þ Q A
ð1Þ
where C is the concentration of the growth-limiting nutrient (kg m3), Q the volumetric flow rate in the reactor (m3 s1), and A the surface area covered by the biofilm (m2). Therefore, biofilm activity at the scale of the reactor is the average flux of nutrients across the biofilm surface, which corresponds to the approach delineated in Equations (12) and (13) used in graphical procedure to evaluate pilot-plant observations. Average biofilm activity in a reactor is a useful descriptor of reactor performance. However, when the underlying biofilm processes are to be studied, an image of local biofilm activity is often required. This information can be extracted from growth-limiting substrate concentration profiles measured at
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surface on the oxygen profile coincides with the inflection point of the nutrient concentration profile. It is not easy to determine the exact position of the surface, though. We use a simplified procedure, explained later in Figure 16, to find the approximate position of biofilm surface on concentration profiles measured with microelectrodes. One use of such data is to estimate the local biofilm activity in terms of the flux of the growth-limiting nutrient at the location where the profile was measured. The flux of the nutrient across the biofilm surface, JLF at the location of the measurement is computed as the product of the slope of the concentration profile at the biofilm bulk solution interface by the diffusivity coefficient in water of the substance whose concentration was measured:
Oxygen concentration (mg l−1)
6 5 4 3 2 1
JLF ¼ Dw
0 0
300
600
900
1200
1500
Distance from the bottom (µm) Figure 7 Oxygen concentration profile. The vertical line marks the approximate position of the biofilm surface (Rasmussen and Lewandowski, 1998).
selected locations in the biofilm, as shown in Figure 7. The results from the two scales of observation – (1) the local biofilm activity evaluated from the concentration profiles and (2) the average biofilm activity evaluated from the mass balances around the reactor – provide different types of information. The measurements at the microscale deliver information that cannot be extracted from the measurements at the macroscale. For some biofilm processes, it is important to quantify the extreme values of biofilm activity because the locations in the biofilm where these extreme values occur exhibit extreme properties. For example, in studying MIC, which causes highly localized damage to metal surfaces, it is important to evaluate the extreme values of biofilm activity because the extreme, and highly localized, microbial activity in biofilms determines the extent of microbial corrosion. The average biofilm activity estimated from measurements at the macroscale cannot deliver this information.
4.15.2.2.4 Quantifying local biofilm activity and mass transport in biofilms from microscale measurements The profiles of flow velocity and growth-limiting substrate concentration shown in the conceptual image depicted in Figure 5 can be measured experimentally. Their interpretation leads to a better understanding of the processes occurring in biofilms. Figure 7 shows an oxygen concentration profile measured in a biofilm using an oxygen microelectrode. Nutrient concentration profiles, such as the one shown in Figure 7, are composed of two parts, the part above and the part below the biofilm surface. Different factors shape these parts of the profile: the shape of the profile above the surface is dominated by bulk liquid hydrodynamics, whereas the shape of the profile below the surface is dominated by microbial respiration in the biofilm. These two parts are described by different equations but are connected at the biofilm surface by the requirement of oxygen flux continuity. The position of the
dC dx ðxxs Þ¼0
ð2Þ
where Dw is the diffusivity in water of the substance selected for the evaluation of biofilm activity, usually the growthlimiting nutrient (m2 s1). Diffusivity of this substance in the biofilm is not constant, but instead it varies with distance, as explained below. Early mathematical descriptions of biofilm activity and the shape of the concentration profile within the biofilm were based on the conceptual model of so-called uniform biofilms, depicting biomass uniformly distributed in the space occupied by the biofilm (Atkinson and Davies, 1974; Williamson and McCarty, 1976). Formally, these early mathematical models of microbial activity in biofilms imitated the models of microbial activity in suspension, with the addition of mass transport resistance. They quantified the equilibrium between the rate of utilization of the growth-limiting nutrient and the rate of mass transport in one dimension, toward the surface:
2 qC q C mmax Xf C ; ¼ Df qt f qx2 f Yx=s Ks þ C
0 r x r xs
ð3Þ
At steady state, this equation delivers
Df
d 2C mmax CXf ¼ 2 Yx=s ðKs þ CÞ dx
ð4Þ
Two boundary conditions were generally used to specify the concentrations of oxygen at the bottom and surface of the biofilm:
dC dx
¼ 0;
Cðx¼xs Þ ¼ Cs;
t0
ð5Þ
ðx¼0Þ
where Df is the averaged effective diffusivity of growth-limiting nutrient in the biofilm (m2 s1); x the distance from the bottom (m); xs the distance from the biofilm surface in the new system of coordinates (m); Xf the averaged biofilm density (kg m3); Yx/s the yield coefficient (kg microorganisms/kg nutrient); mmax the maximum specific growth rate (s1); Ks the Monod half-rate constant (kg m3); C the growth-limiting substrate concentration (kg m3); and Cs the growth-limiting substrate concentration at the biofilm surface (kg m3). These early models were subsequently refined by adding additional factors affecting biofilm processes, such as bacterial
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d dC d 2 C dDfl dC mmax CXfl Dfl ¼ Dfx 2 þ ¼ dx dx dx dx dx Yx=s ðKs þ CÞ
ð6Þ
where Dfl is the local effective diffusivity of the growth-limiting nutrient (m2 s1) and Xfl the local biofilm density (kg m3). Accepting that diffusivity and biofilm density are variable introduces two new variables into the equation, and functions describing changes in effective diffusivity and biofilm density need to be quantified before the equations can be solved. Experimental data show that density changes are surprisingly regular in biofilms and can be described as a linear function of biofilm depth. Relative surface-averaged effective diffusivity
0.70 0.65
D*fz = 0.001z + 0.2968
0.60 0.55 D fz
growth and decay in a steady-state biofilm (Rittmann and McCarty, 1980a, 1980b) and then the model was extended to include unsteady states and dual nutrient limitations (Rittmann and Brunner, 1984; Rittmann and Dovantzis, 1983). One of the most popular biofilm models, initially marketed as a software called BIOSIM (Wanner and Gujer, 1986), was later improved to include irregular biofilm structure and renamed AQUASIM (Wanner et al., 1995; Wanner and Reichert, 1996). The growing popularity of the conceptual model of heterogeneous biofilms coincided with the growing popularity of cellular automata (CA) (Wolfram, 1986), and it is not surprising that the heterogeneous biofilm structures were modeled using CA procedures (Wimpenny and Colasanti, 1997a, 1997b). Soon after, Picioreanu et al. (1998a, 1998b) improved this model using more realistic assumptions and used differential equations to describe mass transport with the discrete model describing the structure (Picioreanu et al., 1998a, 1998b). Since its early applications, CA remains the most popular model used to generate biofilm structure. Further improvement of the biofilm model came from Kreft et al. (2001), who developed a two-dimensional (2-D) multinutrient, multi-species model of nitrifying biofilms to predict biofilm structures, that is, surface enlargement, roughness, and diffusion distance. These authors compared the predicted structure of the biofilm with the predictions of the biomass (cells and EPS)-based model developed by Picioreanu et al. (1998a, 1998b), and concluded that the two models had similar solutions. Meanwhile, biofilm researchers urgently needed mathematical description of the biofilm processes that could be used to describe recent progress in understanding biofilm processes. The main problems that needed to be addressed were horizontal and vertical profiles in mass transport and activity in biofilms. These were experimentally verified and the assumption that the effective diffusivity and biofilm density were constant across the biofilm had become difficult to defend. Biofilm diffusivity decreases toward the bottom of the biofilm and biofilm density increases. There have been attempts to include these results in the modeling of biofilm processes but they lead to more complicated mathematical expressions in which diffusivity and biofilm density are functions of distance. To simplify these expressions it is possible to model a biofilm as a stack of layers with constant diffusivity and density, which change from layer to layer rather than continuously. At steady state, this approach delivers the mass transport and activity related to the local properties of the biofilm:
535
*
0.50 0.45 0.40 0.35 0.30 0.25
50
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150
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300
Distance from the bottom, z (µm) Figure 8 The surface-averaged relative effective diffusivity (Dfz*) is multiplied by the diffusivity of the growth-limiting nutrient in the water to calculate the surface-averaged effective diffusivity (Dfz). Since, in the example, the growth-limiting nutrient is oxygen, to calculate the effective diffusivity of oxygen at various distances from the bottom, we must multiply the relative effective diffusivity at various distances from the bottom by the diffusivity of oxygen in water (2.1 105 cm2 s1) (Beyenal and Lewandowski, 2005).
profile, reproduced from Beyenal and Lewandowski (2005), is shown in Figure 8. Assuming that biofilm density varies with depth in a linear fashion, as shown in Figure 8, the diffusivity gradient (x) is constant:
dDfx ¼z dx
ð7Þ
At steady state, this simplifies Equation (5) to the form
Dfl
d 2C dC mmax CXfl ¼ þz 2 dx dx Yx=s ðKs þ CÞ
ð8Þ
Further, it has been demonstrated that in biological aggregates, including biofilms, density is related to effective diffusivity (Fan et al., 1990):
Dfl ¼ 1
0:43X0:92 fl 11:19 þ 0:27X0:99 fl
ð9Þ
Using this equation, we can estimate biofilm density from the variation in local effective diffusivity (Figure 9).
4.15.2.2.5 Horizontal variability in diffusivity and microbial activity in biofilms Concentration profiles of growth-limiting nutrients, such as the one shown in Figure 7, are taken at a specific location in a biofilm. Based on the results, the biofilm activity at that location can be computed. However, when the next profile is taken at another location, even as close as several micrometers from the first location, the two profiles can be significantly different. This is not surprising, considering that biofilms are heterogeneous. However, it brings into question the practice of
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evaluating biofilm activity based on a single measurement at an arbitrarily selected location. For microscale measurements in stratified biofilms, the selected variable, such as local effective diffusivity or local dissolved oxygen concentration, is measured at locations on a grid (Figure 10). Grids are positioned at various distances from the bottom. The results are then presented as maps of the distributions of the selected parameter at the specified distances from the bottom, as shown in Figure 11. One of the main advantages of this approach is that it allows us to average the concentrations of oxygen at the selected distances from the bottom and arrive at a representative profile of oxygen that illustrates its distribution across the biofilm and also shows the deviations from the average due to biofilm heterogeneity. The maps of oxygen distributions shown in Figure 11 served to construct the representative profile of oxygen across this biofilm shown in Figure 12.
100 Pseudomonas aeruginosa (v = 3.2 cm s−1) Mixed culture (v = 1.6 cm s−1) Mixed culture (v = 3.2 cm s−1)
Biofilm density (g l−1)
80
60
40
20
0
0
100 200 300 400 Distance from the bottom, z (μm)
500
Figure 9 Variation in biofilm density with distance from the bottom (Beyenal et al., 1998).
4.15.2.2.6 Mechanism of mass transfer near biofilm surfaces When the local nutrient concentrations measured across a biofilm are plotted versus distance, they form a nutrient concentration profile. It would be expected that the shape of the nutrient concentration profile will follow the shape of the local mass transport coefficient profile when they are measured at the same location. It would also be expected that, at locations where the local mass transport coefficient is high, the local nutrient concentration will be high as well, at least higher than at a location where the local mass transport coefficient is low. Figure 13 shows profiles of oxygen concentration and local mass transport coefficient measured at the same location in a biofilm (Rasmussen and Lewandowski, 1998). As can be seen in Figure 13, the mass transport coefficient profile does not correlate well with the oxygen concentration profile. Approaching the biofilm surface, for example, the oxygen concentration decreases rapidly and reaches quite low levels at the biofilm surface, while the local mass transport coefficient remains quite high at that location. This observation seems difficult to explain: since there is no oxygen consumption in the bulk, the oxygen concentration profile would be expected to follow the shape of the mass transport coefficient profile much closer than it does in Figure 13. However, although these two profiles do not match, each of them is consistent with our knowledge of the system’s behavior. We expect to measure a low concentration of oxygen at the biofilm surface: this result fits the concept of a mass transfer boundary layer of high mass transport resistance above the biofilm surface. Measuring a high mass transport coefficient near the biofilm surface is also not surprising because, as we have estimated, convection is the predominant mass transport mechanism in that zone. The two features cannot coexist: high mass transport resistance and convection. To explain this apparent discrepancy, we need to examine the procedure for measuring flow velocity in biofilms. All available flow velocity measurements in biofilms report only one component of the
Figure 10 Microscale measurements in stratified biofilms. The selected variable, such as the local effective diffusivity or local dissolved oxygen concentration, is measured at the locations where the gridlines intersect. Such grids are positioned at various distances from the bottom (MSU-CBE, P.Dirx).
Biofilms in Water and Wastewater Treatment
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Figure 11 Distribution of oxygen measured in a biofilm at the specified distances from the bottom (Veluchamy, 2006).
flow velocity vector, parallel to the bottom. Based on these results, we estimated that mass transport is controlled by convection near biofilms. However, the convective mass transport rate equals the nutrient concentration times the flow velocity component normal to the reactive surface. The component of the flow velocity parallel to the surface has nothing to do with the convective mass transport toward that surface. Consequently, the estimate of the mass
transport mechanism based on flow velocity holds only in the direction in which the flow velocity was measured. Indeed, when the flow near a surface is laminar, the laminas of liquid slide parallel to the surface, and there is little or no convection across these layers: the mass transport parallel to the surface is convective, while the mass transport perpendicular to the surface remains diffusive. This mechanism is visualized in Figure 14.
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Biofilm 6
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Figure 12 Surface averaged oxygen concentrations (CSA) and standard deviations computed for each data set in Figure 11. The average oxygen concentrations form a representative profile of oxygen concentration, characterizing the area covered with the biofilm, and the envelope of the standard deviation is a measure of the heterogeneity of the measured variable, oxygen concentration in this case (Veluchamy, 2006).
0.2 0.1
k/kmax
0 0
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200 400 600 800 1000 1200 1400 Distance from substratum (µm)
Figure 13 Profiles of oxygen and local mass transfer coefficient through a thin biofilm cluster (’, dissolved oxygen; , local mass transfer coefficient). The vertical line marks the observed thickness of the biofilm. At distances of less than 30 mm, the wall effect caused the local mass transport coefficient to decrease. The biofilm thickness was 70 mm in this location. The value of k/kmax was only slightly affected by the presence of the biofilm up to a distance of less than 30 mm from the substratum (Rasmussen and Lewandowski, 1998).
4.15.2.2.7 Biofilm processes at the macroscale and at the microscale Accurate mathematical models are necessary for advances in biofilm research. Biofilm researchers use mathematical models of biofilm processes not only to predict the outcome of these processes, but also to interpret the results of biofilm studies. In the absence of suitable models, the interpretation of biofilm studies is impaired. Biofilm science and technology are relatively young, and mathematical descriptions of biofilm
processes often lag behind the rapidly expanding knowledge of biofilm processes. On the other hand, most of the experience that was accumulated in modeling biofilm processes in water and wastewater treatment was based on the operating reactors with suspended biomass. Biofilm reactors are different, and some effects common in biofilm reactors are much less usual in reactors with suspended biomass. One effect that is particularly difficult to accommodate in biofilm models is the influence of biofilm heterogeneity on biofilm processes. Biofilm models that describe biofilm processes on the scale of the entire reactor assume that the biofilm is uniformly distributed and its effects do not depend on the location in the reactor. This assumption, which is justified in the case of well-mixed reactors, may or may not be justified in biofilm reactors. With the current sophistication in exploring biofilm processes at the microscale, it is not surprising to observe that the local conditions quantified in biofilms deviate widely from the average conditions described by the biofilm models. One hopes that these deviations from the idealized models cancel each other and that overall, at the macroscale, they do not matter much. One particularly troubling problem is the definition of and the existence of a steady state in biofilm reactors. Defining a steady state in a biofilm reactor may well be the most important question facing biofilm researchers, both those who focus on experiment and those who focus on modeling. The existence of a steady state is obvious in flow reactors, where microbial growth occurs in suspension. In such reactors, the interplay among the microbial growth rates, biomass concentration, and hydraulic and biomass retention times leads to a steady state in which process variables do not change for a long time. In contrast, the reasons for the existence of a steady state in a biofilm reactor are much less clear because an important condition for a steady state is not satisfied in a biofilm reactor: the concentration of biomass in a
Biofilms in Water and Wastewater Treatment
Convection Diffusion Convection and diffusion
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Direction of mass transport Convection Diffusion Direction of measured flow velocity
Figure 14 Alternating zones of convective and diffusive mass transport in heterogeneous biofilms. This hypothetical model of mass transport is consistent with the results in Figure 13. Mass transport in the space occupied by the biofilm is convective, but the amount of nutrient delivered to this space is limited by the diffusive mass transport just above the biofilm surface (MSU-CBE, P.Dirx).
biofilm reactor is not a simple function of retention time and growth rate. Some biofilm technologies actually take advantage of this fact and grow biofilm microorganisms using retention times at which the microorganisms would be washed out from reactors operated with suspended microorganisms. Practically, this problem corresponds to the fact that we are uncertain what function describes detachment in biofilms, and what mechanisms are involved in biofilm detachment, except perhaps for shear stress. The mechanism of biofilm sloughing remains unknown. A steady state for the biomass concentration assumes that the same amount of biomass is generated as is removed by various processes, particularly biofilm detachment. One can argue that if the biofilm reactor is large enough, the microscale biofilm processes will average out on the scale of the reactor, and that this average may be stable even if the components of the average vary over time. This argument, even if it is true, however, does not settle the issue. A question follows: how large does the reactor have to be to ensure that the variations in the microscale biofilm processes average out and the reactor reaches a steady state at the macroscale? There are also difficulties at the microscale. Experimentally measured concentration profiles and flow velocity profiles corroborate the conceptual model shown in Figure 5. However, when it comes to interpreting experimental data, the idealized image of biofilms in Figure 5 is not adequate for many reasons. One reason is shown in Figure 15: the difficulty with locating the position of the biofilm surface. The position of the biofilm surface is important: one of the boundary conditions in the equation describing biofilm activity and mass transport specifies the conditions at the biofilm surface. As can be seen in Figure 15, however, locating it is not trivial. This problem has been addressed experimentally by judiciously locating the surface on a nutrient concentration profile at the location where the profile ends its curvature near the bottom. The rule of locating the biofilm surface at that location has been developed based on the results of studies in which an oxygen electrode and an optical sensor were used to measure the oxygen concentration profile and detect the biofilm surface, where optical density changed (Figure 16). The position of the biofilm surface coincides with the location where the oxygen profile becomes linear. The biofilm surface in Figure 7 was positioned using this principle.
Figure 15 Surface of a biofilm grown at a flow velocity of 0.81 m s1 (Groenenboom, 2000).
4.15.2.2.8 Biofilms in conduits Among the many possible effects that biofilms may have in water conduits, we will discuss two effects in more detail: (1) the effect on flow characteristics – pressure drop in conduits and (2) the effect on material performance – MIC. Flow velocity near the biofilm surface. It is well known that flow velocity affects biofilm processes. Figure 5 shows an example of the effect of flow velocity on mass transport dynamics near the biofilm surface. However, biofilm also affects flow velocity: flow velocity near a wall covered with biofilm is different from that near a wall with no biofilm. Figure 17 shows this effect. The effect of biofilm on flow velocity distribution most certainly influences the dynamics of mass transfer. However, this is not the only effect that biofilm has on hydrodynamics. For example, it is well known that biofilms increase the pressure drop in conduits, but it is not clear what the mechanism of this process is or how to quantify it. To predict pressure drop in pipes the Moody diagram is used, which correlates the Reynolds number and the relative roughness to provide the friction factor, f. This friction factor is then
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Figure 16 Profiles of oxygen concentration and optical density in a biofilm. A combined microsensor – an oxygen microelectrode and an optical density microprobe – permitted locating the biofilm surface at 0.60 mm from the bottom. This distance, when marked on the oxygen concentration profile, indicates that the biofilm surface is at the beginning of the linear part of the oxygen profile within the mass transfer boundary layer; I is the local light intensity, and Io is the maximum light density (Lewandowski et al., 1991).
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Depth (µm) Figure 17 The flow velocity profile near a wall covered with a biofilm is different from the flow velocity profile near the same wall without the biofilm (DeBeer et al., 1994).
plugged into the Darcy–Weisbach equation to calculate the pressure drop:
HL ¼ f
l V2 D 2g
ð10Þ
where HL is the head loss due to friction, l the pipe length, V the average fluid velocity, g the gravitation constant, D the pipe diameter, and f the friction factor provided by the Moody diagram. When the flow velocity increases, the thickness of the boundary layer decreases, and the roughness elements protrude through the boundary layer, further affecting the drag and the pressure drop.
Unfortunately, the Moody diagram is of little help in predicting the pressure drop in conduits covered with biofilms. The pressure drop in such conduits is caused by different factors than the pressure drop in conduits without biofilms because different mechanisms are responsible for the shape of the pressure drop in each of these conduits. These differences sometimes demonstrate themselves in the form of puzzling experimental results, such as decreasing pressure drop resulting from increasing flow velocity, which is a consequence of the elastic and viscoelastic properties of biofilms. Microcolonies are made of bacterial cells embedded in gelatinous EPS that can change shape under stress. At high flow velocities the hydrodynamic boundary layer separates from the
Biofilms in Water and Wastewater Treatment
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microcolonies, causing pressure drag downstream of the microcolony and pulling the material in this direction. The microcolonies slowly flow under the strain, forming elongated shapes that we call streamers. Streamers are often seen when biofilms grow at high flow velocities. The streamers contribute to pressure drop by moving rapidly and dissipating the kinetic energy of the flowing water. Another important consequence of a streamer’s oscillations is that they are transmitted to the underlying microcolonies, which also oscillate rhythmically. This system reacts with turbulent boundary layers much differently than the rigid surface roughness elements of clean pipes do. One way to gain experimental access to the interactions between flowing water and biofilm is to monitor flow velocity profiles. Imaging flow velocity profiles makes it possible to evaluate the effect of biofilm formation on the flow in conduits by quantifying its effect on the entry length in the conduit. The hydrodynamic entry length is defined as the distance needed to develop a steady flow, after the water has passed through the entrance to the reactor. If the presence of biofilm makes the entry length longer, then the biofilm contributes to flow instability, and vice versa. There is a simple relation between the Reynolds number and the entry length: the higher the Reynolds number, the longer the entry length. This effect was used as a base for quantifying the effects of biofilm on the flow in conduits. Flow velocity distribution was measured in a rectangular reactor when the flow velocity was increasing from one measurement to another. As the flow velocity and the Reynolds number increased, the flow stability was monitored in a rectangular conduit using nuclear magnetic resonance (NMR) imaging. The results, shown in Figure 18, demonstrate that the presence of biofilm actually made the flow more stable. The entry length was shorter and the flow reached stability closer to the entrance in the presence of biofilm than in its absence. It is difficult to interpret this result immediately because it is well known that the presence of biofilm increases pressure drop in conduits: traditionally, pressure drop in pipes is related to friction. As pressure drop is larger in biofilm-covered pipes, a natural conclusion was that biofilms must increase friction and therefore the presence of the biofilm should introduce flow instability rather than reduce it. The relation between flowing water and biofilms is determined by two facts: (1) biofilms are made of viscoelastic polymers which actively interact with the oscillations generated by the flow of water and (2) the flow of water affects the biofilm structure. Based on what we now understand, at low flow velocities biofilms can effectively smooth surfaces and stabilize the flow because the oscillating layer of elastic polymeric matrix can effectively damp the vibrations coming from the flowing water. This effect delays the onset of turbulence in conduits covered with biofilm and explains the results shown in Figure 18. However, as the flow velocity increases further, the elastic polymeric matrix must oscillate faster and faster and, eventually, the frequency of its oscillation cannot follow the frequency of the incoming eddies. At that point the biofilm oscillation is out of phase and the biofilm not only fails to damp the flow instabilities but also actively introduces instability by randomly oscillating at a different frequency than the incoming eddies. The pressure drop in the conduit
541
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Figure 18 Flow velocity profiles in a rectangular conduit whose walls were colonized with a biofilm. The increasing flow velocity did not affect the character of the velocity profiles in the reactor with biofilm. On the other hand, the same increase in velocity had a pronounced effect on the reactor without biofilm.
increases rapidly. This effect was, in early biofilm works, mistaken for a similar effect caused by rough surface elements. For example, Picologlou et al. (1980) observed a considerable increase in frictional resistance after the film thickness reached a value approximately equal to the calculated thickness of the hydrodynamic boundary layer for a clean surface. In clean pipes covered with surface roughness elements, when flow velocity increases the boundary layer becomes thinner and at some flow velocity the boundary layer thickness is smaller than the height of the roughness elements. When this happens, the roughness elements protrude through the boundary layer and cause an additional drag, which exhibits itself in a sudden increase of the pressure drop for flow velocities exceeding this critical flow velocity. This model was commonly accepted and was used to explain the pressure drop in conduits covered with biofilms, although even at that time some authors warned that this might not be the true mechanism of the process (Characklis, 1981). Currently, there are no models that can account accurately for pressure drop in conduits covered with biofilm.
4.15.3 Part II: Biofilm Reactors Biological systems treating municipal wastewater require (1) the accumulation of active microorganisms in a bioreactor and (2) the separation of the microorganisms from treated effluent. In suspended growth reactors, such as the activated
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sludge process, microorganisms grow and bioflocculate; the resultant flocs are suspended freely in the bulk phase. Flocculated bacteria are then separated from the bulk liquid by sedimentation or membranes. Clarifier-coupled suspended growth reactors rely on return activated sludge, or underflow, from the coupled clarifier to provide the desired active biomass concentration in the bioreactor. Consequently, clarification unit processes may be limited by the hydraulic loading rate (HLR) or solids loading rate (SLR). Biofilm reactors retain bacterial cells in a biofilm that is attached to the fixed or free moving carriers. The biofilm matrix consists of water and a variety of soluble (C) and particulate (X) components that include soluble microbial products, inert material, and EPS. Without suspended biomass, the bioreactor is decoupled from the liquid–solids separation unit. Active biomass concentrations inside the biofilm are large at 10–60 g of volatile suspended solids (VSS) l1 of biofilm. This biomass range can be compared with the range of concentrations expected for suspended growth reactors, which is typically 3–8 g VSS l1 of reactor volume. The lower value in this range is associated with clarifier-coupled activated sludge processes, and the upper range with membrane bioreactors. In biofilm reactors, bacteria attached to carriers periodically detach from the biofilm matrix and exit the system in the effluent stream. Figure 19 provides a conceptual illustration of different biofilm reactor types. Biofilm reactors can be classified based on the number of phases involved – gas, liquid, solid – according to the biofilm being attached to a fixed or moving carrier within the reactor. They are also classified based on how electron donors or acceptors are applied to seven basic types as listed below (adapted from Harremo¨es and Wilderer (1993)): 1. Three-phase system – fixed biofilm-laden carrier, bulk water, and air. Water trickles over the biofilm surface and
2.
3.
4.
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air moves upward or downward in the third phase (e.g., trickling filter (TF)) (Figure 19(a)). Three-phase system – fixed (or semifixed) biofilm-laden carrier, bulk water, and air. Water flows through the biofilm reactor with gas bubbles (e.g., aerobic biologically active filter (BAF)). Gravel is a fixed media and polystyrene beads are semifixed (Figures 19(b) and 19(c)). Three-phase system – moving biofilm-laden carrier, bulk water, and air. Water flows through the biofilm reactor. Air is introduced with gas bubbles (e.g., aerobic moving bed biofilm reactor (MBBR)) (Figure 19(g)). Two-phase system – moving biofilm-laden carrier and bulk water. Water flows through the biofilm reactor with the electron donor and electron acceptor (e.g., denitrification fluidized bed biofilm reactor (FBBR)) (Figure 19(g)). Two-phase system – fixed biofilm-laden carrier material and bulk water. Water flows through the biofilm reactor with the electron donor and electron acceptor (e.g., denitrification filter) (Figures 19(b) and 19(c)). Three-phase membrane system – a microporous hollowfiber membrane with biofilm and water on one side and gas on the other, diffusing through the membrane to the biofilm (e.g., membrane biofilm reactor (MBfR)) (Figure 19(h)). Two-phase membrane system – a proton exchange membrane separating a compartmentalized biofilm-laden anode from a compartmentalized cathode with water on both sides, but with the electron donor on one side and electron acceptor on the other (e.g., biofilm-based microbial fuel cell (MFC)).
Biofilms are ubiquitous in nature and in engineered systems and can be used beneficially in municipal water and wastewater treatment. Biofilm and suspended growth reactors can meet similar treatment objectives for carbon oxidation,
Air
Air (a)
(b)
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Figure 19 Types of biofilm reactors: (a) trickling filter; (b) submerged fixed bed biofilm reactor operated as up flow or (c) down flow mode; (d) rotating biological contactor; (e) suspended biofilm reactor including airlift reactor; (f) fluidized bed reactor; (g) moving bed biofilm reactor; and (h) membrane attached biofilm reactors. From Morgenroth (2008) Modelling biofilm systems. In: Henze M, van Loosdrecht MCM, Ekama G, and Brdjanovic D (eds.) Biological Wastewater Treatment – Principles, Modelling, and Design, pp. 457–492. London: IWA Publishing.
Biofilms in Water and Wastewater Treatment
nitrification, denitrification, and desulfurization. Biofilm reactors have also been used for the treatment of a variety of oxidized contaminants including perchlorate and bromate. The same microorganisms are responsible for biochemical reactions in both activated sludge and biofilm systems, and respond in the same way to local environmental conditions (i.e., pH, temperature, electron donor, electron acceptor, and macronutrient availability) (Morgenroth, 2008). A key component to be considered by anyone who is evaluating a biofilm reactor is the effect of multiple substrates and biomass fractions and the manner in which the reactor is affected by mass-transport limitations. Substrates typically considered are: 1. soluble compounds, including electron donors (e.g., readily biodegradable chemical oxygen demand (rbCOD), NHþ 4 , NO2 , and H2), electron acceptors (e.g., O2, NO3 , 2 3 NO2 , and SO4 ), and nutrients and buffers (e.g., PO4 , NHþ 4 , and HCO3 ) and 2. particulate compounds, including electron donors (e.g., slowly biodegradable COD (sbCOD)), active biomass fractions (e.g., heterotrophic and autotrophic bacteria), inert biomass, and EPS.
4.15.3.1 Application of Biofilm Reactors This section exists to provide the reader with a general overview of biofilm reactor applications. While general biofilm reactor applicability is described here, several treatment scenarios exist that are not conveniently generalized yet warrant the use of biofilm reactor technology. Water-quality regulations exist to protect human health and the water environment. Organic matter and the nutrients such as nitrogen and phosphorus are major contributors to water-quality impairment. In municipal wastewater-treatment scenarios, biofilm reactors are generally applied for the removal of carbon-based organic matter and/or nitrogenous compounds. Specifically, these biofilm reactors may achieve carbon oxidation, combined carbon oxidation and nitrification, tertiary nitrification, or tertiary denitrification. Biofilm reactors are not commonly used for biological phosphorus removal. Biofilm reactors treating industrial wastewaters have been applied to meet treatment objectives similar to those in municipal wastewater treatment and industrial pretreatment. The objective of pretreatment is to process industrial waste streams until their characteristics are similar to raw sewage (see Metcalf and Eddy (2003) for a description). As a result the industry can then discharge their treated wastewater into municipal sewers where further processing is accomplished at a municipal wastewater-treatment plant. Biofilm reactors are common for industrial applications because the processes are reliable, robust, easy to operate, and resilient to toxic or shock loading.
4.15.3.1.1 Techniques for evaluating biofilm reactors Several approaches exist to evaluate biofilm reactors. The primary objective of a biofilm or biofilm reactor model is to predict soluble substrate flux (J) through the biofilm surface. This flux (M L2 T1) can be used to obtain an estimate of the (1) overall biofilm reactor performance, (2) required biofilm surface area, (3) electron acceptor (e.g., dissolved oxygen), (4)
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external electron donor (e.g., methanol or hydrogen), and (5) biosolids management requirements. This section discusses the relative benefits and limitations to some general methods of evaluating biofilm reactors. The use of mathematical biofilm models is common in both research and practice, but only a cursory presentation of their mathematical description is presented. Excellent resources exist describing aspects of mathematical modeling of biofilms and biofilm reactors (for additional information, see Wanner et al. (2006) and Morgenroth (2008)). The approaches discussed here include a graphical procedure, empirical models, semiempirical models, and mechanistic mathematical models.
4.15.3.1.2 Graphical procedure A graphical procedure can be used to determine the total hydraulic load (THL) required to decrease a substrate concentration, and by definition the biofilm surface area required to provide a desired substrate concentration remaining in the effluent stream. These items can be determined directly. The graphical procedure can be used to determine effluent substrate concentration from any series of continuous flow stirred tank reactors (CFSTRs). A stepwise procedure must be used when a series of CFSTRs will be used. Antoine (1976) and Grady et al. (1999) developed the graphical procedure described here and the approach is valid for any biofilm-based CFSTR. If multiple stages are expected to have different characteristics, then the graphical method requires different flux curves to describe system response in each of the CFSTRs. The procedure requires a graphical representation of substrate flux (J) as a function of bulk-liquid substrate concentration (CB). This relationship between flux and bulk-liquid substrate concentration can be obtained from numerical simulations, full-scale or pilot-plant observations. In practice, this graphical procedure is typically used to extend pilot-plant observations to full-scale biofilm reactor design criteria. The process designer should recognize that the relationship between flux and bulk-liquid substrate concentration is based on the system and location. Therefore, the flux curve required to implement the graphical procedure may not be obtained from or correlate well with values reported in the literature or from different systems. As a result, the process designer should consider carefully the conditions under which the flux curve was developed before applying results. A flux curve representing mass transfer and environmental conditions characteristic of a specific system and operating mode may not be the representative of different biofilm reactor types designed to meet the same treatment objectives. A flux curve generated for the same biofilm reactor type under similar operating conditions, however, may offer some direction in the absence of system-specific numerical simulation or pilot/full-scale observations. When using the graphical procedure to evaluate pilot-plant observations, fluxes should be compared to rates in full-scale systems. Any flux that deviates significantly from those reported for biofilm reactors in published studies should be used only after careful consideration. Pilot or experimental systems may promote a greater flux than expected. The basis for the graphical procedure is a material balance on a
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biofilm-based CFSTR:
0 = Q ⋅ Cin ,i − Q ⋅ C B ,i − J LF ,i ⋅ A − rB ,i ⋅ VB mass per time input
mass per time output
biofilm transformation rate
suspended growth transformation rate
ð11Þ
where Q is the flow rate through the system (m3 d1); Cin,i the influent concentration of soluble substrate i (g m3); CB,i the effluent, or bulk-liquid, concentration of soluble substrate i (g m3); JLF,i the flux of soluble substrate i across the biofilm surface equal to the average biofilm activity in the reactor, as shown in Equation (1) (g m2 d1); A the biofilm surface area (m2); rB,i the rate of substrate i conversion because of suspended biomass (g m2 d1); and VB the bulk-liquid volume (m3). Assuming that transformation occurring in the bulk liquid is negligible, the suspended growth transformation rate (Equation (11)) can be neglected. Rearranging Equation (11) provides the rationale for the graphical procedure:
JLF;i ¼
Q Q Cin;i CB;i A A |fflfflfflffl{zfflfflfflffl} |{z} const:
ð12Þ
slope
The slope, or ( (Q/A)), is referred to as the operating line and represents the total hydraulic load on each stage. Figure 20 illustrates the graphical method. The flux curves have been created based on observations in the first and second stage of a post-denitrification biofilm reactor. The ordinate represents nitrate–nitrogen flux and the abscissa nitrate–nitrogen concentration remaining in the effluent stream. The graphical solution indicates that the
first-stage denitrification biofilm reactor effluent nitrate– nitrogen concentration is approximately 3.9 mg l1. The secondstage effluent nitrate–nitrogen concentration is approximately 1.1 mg l1 with fluxes of approximately 1.6 and 1.1 g m2 d1 in the first and second stage, respectively. The graphical procedure depends on the substrate flux curve(s). The method requires development of multiple flux curves if the performance characteristics of respective stages vary significantly. When using pilot-plant data to generate a flux curve, appropriate scale considerations must be given when designing the pilot unit and experiments.
4.15.3.2 Empirical and Semi-Empirical Models Empirical models can be implemented easily by hand or using a spreadsheet, but they have limited applicability because of their black-box consideration of system parameters. Because environmental conditions and bioreactor configuration affect biofilm reactor performance, a system can respond differently from the description provided by an empirical model. The limited descriptive capacity of empirical models typically results from parameter values and model features based on data that were obtained from few system installations or operating conditions. Therefore, the engineer or scientist should be aware of conditions under which system-specific model parameters have been defined. Significant sources of variability in values include differences in biofilm carrier type and configuration, the extent of concentration gradients external to the biofilm surface, and biofilm composition. Despite their ease of implementation, empirical models can produce results that vary 50–100% of actual system performance.
3.5 Denitrification rate (g m−2 d−1 as NO3−N)
Stage 1 operating line 3.0 Stage 2 operating line
Stage 1 flux response curve
2.5 Stage 2 flux response curve
J LF1
2.0
1.5 −Q/A J LF2
1.0
0.5 CB -stage
0.0 0
1
2
3
CB -stage 1 4
5 6 CNO −N (mg-N l−1) 3
C in 7
8
9
10
Figure 20 Graphical procedure for describing the response of a denitrification moving bed biofilm reactor to defined conditions, including (1) firstand second-stage operating lines and (2) flux curves based on observations at a pilot-scale denitrification moving bed biofilm reactor (Boltz et al., 2010b).
Biofilms in Water and Wastewater Treatment
Coefficient values, and sometimes the empirical models, are typically created to describe system response for the removal of a specific material. The models can be used as an indicator of system viability to meet treatment objectives with respect to the specific process governing transformation. Empirical models are, however, inadequate for describing complex processes such as the explicit evaluation of two-step ammonium oxidation first to nitrite by ammonia-oxidizing bacteria and then to nitrate by nitrite-oxidizing bacteria. Therefore, empirical models have limited application in defining the conditions that either promote or deter complex processes in biological systems. Historically, biofilm reactors have been designed using empirical criteria and models, but this trend is changing. One should recognize that the coefficients in empirical models describing biofilm reactors include system, and many times, location-specific mass-transfer resistances (Grady et al., 1999). For this reason, the values typically differ from apparent or intrinsic values reported in the literature. Once a flux has been determined, Equation (11) can be rearranged, neglecting bulkphase conversion processes, to calculate the material concentration remaining in the effluent:
CB;i ¼ Cin
JLF;i A Q
ð13Þ
If sufficient data exist to allow for the development of parameter values and mathematical relationships capable of describing a complete range of conditions expected when treating municipal wastewater, then empirical models can be used. The addition of model components to account for specific phenomenon encroaches on the premise of mechanistic mathematical model development. For this reason, a distinction is made between empirical and semi-empirical models. Gujer and Boller (1986) and Sen and Randall (2008) provided an example of the latter describing nitrifying TFs, and MBBRs and IFAS systems, respectively.
C
LF
LL
4.15.3.3 Mathematical Biofilm Models for Practice and Research Mathematical modeling can be used to describe certain features of a biofilm or biofilm system (such as a bioreactor) by selecting and solving mathematical expressions. Biofilm reactor research and design commonly involve the use of mathematical biofilm models. These mathematical models are tools that allow the user to efficiently evaluate a variety of complex scenarios. Empirical models fail to provide information that is a concern for biofilm researchers and environmental protection such as biofilm composition and competition among bacteria for multiple substrates and space inside the biofilm, and the influence of individual processes on the interaction between several bacterial types. Mathematical biofilm models have been used as a research tool, but only recently modern biofilm reactors have encouraged the use of biofilm models in engineering practice. Submerged and completely mixed biofilm reactors allow for the application of modern biofilm knowledge, and are conducive to simulation with existing biofilm models (Boltz and Daigger, 2010). As a result, a majority of existing wastewater-treatment plant simulators have been improved to include a biofilm reactor module(s) that is based on the mathematical description of a 1-D biofilm. A user should understand the mathematical biofilm model basis, underlying assumptions, and limitations before applying the model to research or design. A biofilm schematic is shown in Figure 21. The schematic illustrates diffusion and reaction occurring inside a 1-D biofilm. In addition, concentration gradients external to the biofilm surface are illustrated in the manner that they are modeled, namely an external mass transfer resistance represented by a mass transfer boundary layer. The partial differential equation describing molecular diffusion, substrate utilization inside a biofilm, and dynamic accumulation has been presented as Equation (3). It should be emphasized that the basis for a mathematical description of the 1-D biofilm, as described by Equation (3), is simultaneously
C
LF
LL
CB
CB
C LF
C LF
Distance from growth medium
Distance from growth medium
Z Distance from surface, X
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Z Distance from surface, X
Figure 21 Schematic of a 1-D biofilm of thickness LF having an assumed homogeneous (a) and heterogeneous, or layered, (b) biomass distribution. Soluble substrate concentration profile is illustrated with a bulk-liquid concentration (CB) decreasing through a mass transfer boundary layer of thickness LL until reaching the liquid–biofilm interfacial concentration (CLF), and then decreasing through the biofilm.
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occurring molecular diffusion and biochemical reaction. Molecular diffusion is based on Fick’s law. Monod-type kinetics is typically applied to describe the biochemical transformation rate. Analytical solutions to Equation (3) are available only for first- and zero-order rate expressions and assuming steady state. Zero-order kinetics are valid if the bulkliquid substrate concentration is well above the half-saturation concentration (i.e., CB,i 4 Ki), and first-order kinetics is applicable for low substrate concentrations (i.e., CB,ioKi). Solving the second-order differential equations requires constants that can be derived from two boundary conditions described by Equation (5). From the concentration profile (Cf,i(x)) the flux through the biofilm surface (JLF) is calculated as Equation (2). This substrate flux, JLF, is used in biofilm reactor material balances (see Equation (11)). The concentration gradient external to the biofilm surface is not explicitly modeled. Rather, it is modeled as a mass transfer resistance:
JMTBL ¼
1 ðCB;i CLF;i Þ RL
ð14Þ
•
•
Here, JMTBL is the substrate flux in the stagnant liquid layer and RL the mass transfer resistance external to the biofilm. It is helpful to visualize RL by introducing the concept of a mass transfer boundary layer. Defining the thickness of this mass transfer boundary layer provides a more intuitive understanding compared to the mass transfer resistance. Resistance to mass transfer and the mass transfer boundary layer thickness are related according to Equation (15):
RL ¼
LL Dw
ð15Þ
Here, LL is the mass transfer boundary layer thickness and Dw the solute diffusion coefficient in the water phase. The substrate flux through the mass transfer boundary layer (Equation (15)) is linked to the substrate flux across the biofilm surface (Equation (2)). This provides an additional Equation (16) (boundary condition) that is required to calculate the additional unknown value of the substrate concentration at the liquid–biofilm interface (JLF):
JMTBL ¼ JLF
ð16Þ
One of the most difficult aspects of choosing an approach to model biofilms and biofilm reactors is to choose the appropriate level of complexity. An overview of the different model approaches is provided below (after Taka´cs et al., 2010):
•
•
0-D biofilm. One aspect of modeling biofilms is that bacteria are retained in the system and are not washed out with effluent water. The simplest approach for biofilm modeling would be to assume that all biomass in the reactor is exposed to bulk phase concentrations neglecting the effect of mass transport limitations (i.e., 0-D). In wastewater treatment biofilms are relatively thick and are usually masstransfer-limited. Thus, the 0-D modeling approach that neglects mass transfer limitations is not useful except in special cases. 1-D homogeneous biofilm (single limiting substrate). This approach takes into account mass transfer limitations into
•
the biofilm and the corresponding effects on concentration profiles and substrate flux into the biofilm. It is assumed that active bacteria are homogeneously distributed over the thickness of the biofilm. The approach is valid only if calculations are performed for the limiting substrate which has to be determined a priori by the user as described in Morgenroth (2008). The flux of the nonlimiting substrates can be calculated based on reaction stoichiometry. 1-D homogeneous biofilm (multiple substrates and multiple biomass components). One key aspect of modeling biofilms is to evaluate the competition and coexistence of different groups of bacteria and local environmental conditions. Local process conditions can be accurately determined by calculating penetration depths for different soluble substrates. Based on the fluxes the growth of individual groups of bacteria can be determined. To simplify calculations it can be assumed that all bacterial groups are homogeneously distributed over the thickness of the biofilm (Rauch et al., 1999; Boltz et al., 2009a). 1-D heterogeneous biofilm. Different groups of bacteria are competing in a biofilm not only for substrate but also for space where bacteria toward the surface are less influenced by mass transport limitations. Bacteria growing toward the base of the biofilm are often rate limited by substrate availability resulting from mass transfer limitations. On the other hand, these bacteria are better protected from detachment. These 1-D heterogeneous biofilm models must keep track of local growth and decay of the different bacterial groups and of detachment to calculate biomass distributions over the biofilm thickness. 2-D and 3-D biofilm models. Practically, biofilms are not as smooth and flat as is assumed in 1-D biofilm models. Mathematical models have been developed that predict the development of biofilms in two or three dimensions, the influence of the heterogeneous structure on fluid flow, and ultimately the combination of fluid flow and biofilm structure on substrate availability and removal inside the biofilm. For most questions related to practical biofilm reactor studies, such multi-dimensional models are not necessary. However, it is important for model users to recognize that biofilm structure influences local fluid dynamics and external mass transport, which are simultaneously affected by biofilm reactor appurtenances and mode of operation. Such interactions are not accounted for in existing 1-D biofilm models due to a rigid segregation of the bulk phase, mass transfer boundary layer, and biofilm (which is assumed to have a uniform thickness and smooth surface). Multi-dimensional biofilm models have been used to quantify the influence of biofilm structure on local fluid dynamics and external mass transport (Eberl et al., 2000).
Different scales of heterogeneity are relevant for biofilm reactors. The length scale of the biofilm thickness, which is on the order of 100–1000 mm, is taken into account in 1-D and multi-dimensional biofilm models. Substrate fluxes from these simulations can then be integrated into models describing overall reactor performance where the length scale is on the order of 1 m. However, heterogeneities can also be observed in biofilm reactors in between these scales where, in some cases, patchy biofilms are observed and where certain
Biofilms in Water and Wastewater Treatment
parts of the biofilm support medium is bare while at other areas dense biofilms develop (B1–10 cm). These heterogeneities in between the small and the large scale are typically not considered in biofilm models and it is not clear to what extent they are relevant (Taka´cs et al., 2010). No simple and general recommendations can be given on what approach is the most appropriate for describing biofilm reactors. Wanner et al. (2006) provided a detailed description of different modeling approaches and a discussion on how the modeling approaches compare for different modeling scenarios. Many commercially available wastewater-treatment plant simulators used for biofilm reactor design and evaluation takes into account multiple substrates and biomass fractions in either a heterogeneous or a homogeneous 1-D biofilm. Examples of software, and references to the biofilm model that constitutes the biofilm reactor module, that is applied to design, optimize, and evaluate, typically pilot- or full-scale biofilm reactors are summarized in Table 1.
4.15.3.4 Biofilm Model Features Excellent guides exist that describe the mathematical modeling of biofilms (see Wanner et al., 2006; Morgenroth, 2008). However, the state of biofilm modeling is subject to several uncertainties. In the context of this chapter, Boltz et al. (2010a) summarized the following items which cause uncertainty when using 1-D biofilm models to describe biofilm reactors: (1) the fate of particulate substrates, (2) biofilm distribution in the reactor and the effect biofilms have on reactor components, (3) dynamics and fate of biofilm detachment, (4) quantifying concentration gradients external to the biofilm surface, and (5) a lack of generally accepted biofilm reactor Table 1
model calibration protocol. Parameter estimation and model calibration are serious concerns for process engineers who apply biofilm models in engineering practice. Therefore, parameters that are critical components of biofilm reactor models (that use a 1-D mathematical biofilm model) are introduced, including: attachment (kat) and detachment (kdet) coefficients, the mass transfer boundary layer, rate-limiting substrate diffusivity coefficient inside the biofilm (Df,ratelimiting), and the biokinetic parameters maximum growth rate (m) and the ratelimiting substrate half-saturation coefficient (Ki,ratelimiting) (Boltz et al., 2010b).
4.15.3.4.1 Attachment and detachment process kinetics and rate coefficients An accurate mathematical description of particle attachment and detachment processes is a critical component of biofilm (reactor) models. Unfortunately, attachment/detachment process mechanics are poorly understood. Conceptually, particles suspended in the bulk liquid are hydrodynamically transported to the vicinity of the biofilm. From the bulk phase, particles are subjected to concentration gradients external to the biofilm surface. Particles enter the biofilm matrix through channels, crevasses, and other structural irregularities where they attach to the biofilm surface (see Reichert and Wanner (1997) for a description of particle transport within the biofilm matrix). Once entrapped, the particles can be hydrolyzed by extracellular polymeric enzymes resulting in soluble substrate that diffuses into the biofilm. Then, the soluble substrate is subject to well-known biochemical transformation processes that yield biomass. Alternatively, particles that have attached to the biofilm surface from the bulk phase remain unaltered and exit the system after detaching from the biofilm
Biofilm models used in practice (Boltz et al. 2010b)
Software
Company
Biofilm model type and biomass distribution
Reference
AQUASIMTM
EAWAG, Swiss Federal Institute of Aquatic Science and Technology, Du¨bendorf, Switzerland (www.eawag.ch/index_EN) Aquaregen, Mountain View, California (www.aquifas.com) EnviroSim Associates Ltd., Flamborough, Canada (www.envirosim.com)
1-D, DY, N; heterogeneous
Wanner and Reichert (1996) (modified)
1-D, DY, SE and N, heterogeneous 1-D, DY, N, heterogeneous
Sen and Randall (2008)
Hydromantis Inc., Hamilton, Canada (www.hydromantis.com) CH2M HILL Inc., Englewood, Colorado (www.ch2m.com/corporate) ifak GmbH, Magdeburg, Germany (www.ifak-system.com) WRc, Wiltshire, England (www.wateronline.com/ storefronts/wrcgroup.html) MOSTforWATER, Kortrijk, Belgium (www.mostforwater.com)
1-D, DY, N, heterogeneous
AQUIFASTM BioWinTM
GPS-XTM Pro2DTM SimbaTM STOATTM WESTTM
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1-D, SS, N(A), homogeneous (constant Lf) 1-D, DY, N, heterogeneous 1-D, DY, N, heterogeneous 1-D, DY, N(A)a, Nb, homogeneousa, heterogeneousb
Wanner and Reichert (1996) (modified), Taka´cs et al. (2007) Hydromantis (2006) Boltz et al. (2009a; 2009b) Wanner and Reichert (1996) (modified) Wanner and Reichert (1996) (modified) Rauch et al. (1999)a, Wanner and Reichert (1996) (modified)b
a
Rauch et al. (1999) is linked with the definition ’N(A)’ and ’homogeneous’. Wanner and Reichert (1996) (modified) is linked with the definition ’N’ and ’heterogeneous’.
b
1-D, one dimensional; DY, dynamic; N, numerical; N(A), numerical solution using analytical flux expressions; SE, semi-empirical; SS, steady-state. Hydromantis, Inc. (2006) Attached growth models. In: GPS-X Technical Reference, pp. 157–185 (unpublished).
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matrix. Most of the heterogeneous 1-D biofilm models listed in Table 1 describe the rate of particle attachment (rat) as a first-order process ðrat ¼ kat XTSS;bulk Þ depending on an attachment rate coefficient (kat) and the bulk-liquid particle concentration. Boltz and La Motta (2007) presented a model describing variability in this parameter with influent particle concentrations. The researchers postulated that increasing particle concentrations ultimately reduced the biofilm surface area available for particle attachment; thereby, the particle attachment coefficient decreases until reaching a plateau. The plateau was considered commensurate with a condition in which a minimum biofilm area was consistently available as a result of continuously detaching biofilm fragments (during steady operating conditions – variable hydrodynamics can influence biofilm structure). Given the current state of the science, experimental data are required to develop/validate or evaluate existing approaches for simulating the fate of particles in biofilm reactors. Steady-state biofilm models have assumed a constant biofilm thickness in which case biofilm growth is balanced by internal loss (e.g., decay and hydrolysis, or endogenous respiration) and/or detachment. This approach has been successfully applied to simulate biofilm reactors at steady state, but their dynamic simulation requires that a detachment model is included despite rather limited mechanistic understanding. The rate (Morgenroth and Wilderer, 2000; Boltz et al., 2010a) and category (i.e., abrasion, erosion, sloughing, and predator grazing) of detachment can have a significant influence on biofilm reactor simulation and performance (Morgenroth, 2003). Kissel et al. (1984) stated that problems inherent to biofilm detachment modeling include a poor understanding of fundamental (biofilm detachment) process mechanics and the inability to predict exactly at what location inside the biofilm that detachment will occur. Detachment location is important when taking into account a heterogeneous biofilm distribution throughout the reactor either by combining multiple 1-D simulations or by 2- or 3-D modeling (Morgenroth et al., 2000). Unlike attachment, Boltz et al. (2010a) described eight different biofilm detachment rate expressions (rdet) for the heterogeneous 1-D biofilm models listed in Table 1. Detachment rate equations can be categorized based on the aspect controlling detachment: biofilm thickness (LF), shear, or growth/activity. Mixed-culture biofilms, such as those growing in a combined carbon oxidation and nitrification MBBR, are subject to competition for substrate between fast-growing heterotrophic and slow-growing autotrophic organisms (primarily for dissolved oxygen). Morgenroth and Wilderer (2000) performed a modeling study that demonstrated ammonium flux was significantly influenced by the mode of simulated detachment. Essentially, biofilm (thickness) dynamics influenced competition for substrate between heterotrophic and autotrophic organisms; high variations in biofilm thickness dynamics favored the faster growing heterotrophic organisms.
4.15.3.4.2 Concentration gradients external to the biofilm surface and the mass transfer boundary layer Biofilms growing virtually in all full-scale biofilm reactors are subject to some degree of substrate concentration gradients
external to the biofilm surface. Concentration gradients external to the biofilm surface are not explicitly simulated in 1-D biofilm models. Rather, the reduction in concentration of any substrate is modeled as a mass-transfer resistance, RL ( ¼ LL/Dw). Based on the observation that the external masstransfer resistance, RL, is more dependent on biofilm reactor bulk-liquid hydrodynamics than biofilm thickness or surface heterogeneity, the impact of RL can be accounted for by empirical correlations (Boltz et al., 2010a). However, a realistic description of hydrodynamic effects ultimately depends on an accurate estimate of the mass-transfer boundary layer thickness LL. Therefore, the mass-transfer boundary layer thickness is an important facet of biofilm-reactor models that use a 1-D biofilm model. Despite the potential significant impact the mass-transfer boundary layer thickness may have on biofilmreactor model results and process design, factors influencing the interface between the biofilm model and reactor scale is one important feature of biofilm-reactor models that is not well understood.
4.15.3.4.3 Diffusivity coefficient for the rate-limiting substrate inside the biofilm Soluble substrates are primarily transported into biofilms by a combination of advection and molecular diffusion. Generally, the most important mechanism is molecular diffusion (Zhang and Bishop, 1994). The largest component of biofilm is water, but the diffusivity of a solute inside the biofilm is generally less than that in water because of the tortuosity of the pores and minimal biofilm permeability. Consequently, an effective diffusivity must be applied. Many biofilm reactor models treat this value as 80% of the diffusivity in water (i.e., Dw ¼ Df/0.80) (Stewart, 2003). However, it has been demonstrated that the effective diffusion coefficient (Df,i) for any soluble substrate i can vary with depth inside the biofilm (Beyenal and Lewandowski, 2000). The effective diffusivity decreases with depth because of increasing density and decreasing porosity and permeability of the biofilm with depth. Flow velocity past the biofilm is a major influencing factor determining biofilm density. Varying liquid velocity in the vicinity of the biofilm surface can influence a soluble substrate effective diffusivity inside a biofilm. Consequently, the varying flow rate can affect the rate of internal mass transfer and transformation rates (Bishop, 2003). Turbulent, high-sheer stress environments result in planar and denser biofilms while quiescent, low-sheer stress environments will result in rough and less dense biofilms (van Loosdrecht et al., 1995). Picioreanu (1999) defined a growth number ðG ¼ L2f mmax Xf =ðDf CB ÞÞ that can be related to biofilm roughness. According to Picioreanu (1999), the biofilm may have a dense solid matrix and a flat surface when Go5. However, if G 4 10 the biofilm may develop complex structures such as mushroom clusters and streamers.
4.15.3.4.4 Parameters: estimation and variable coefficients A parameter is an arbitrary constant whose value characterizes a system member. Biokinetic parameter estimation is a serious concern for those who seek to use biofilm models for biofilm reactor process design and research because most parameter values cannot be measured directly in full-scale municipal
Biofilms in Water and Wastewater Treatment
wastewater-treatment plants (Brockmann et al., 2008). Parameters exist for every aspect of biofilm models, including stoichiometry, kinetics, mass transfer, and the biofilm itself. A majority of parameter values in modern process models (e.g., those described by Henze et al. (2000)) have a substantial database that serves to define a relatively narrow range of values that are applicable to a majority of municipal wastewater-treatment systems. Existing biofilm models are relatively insensitive to changes in a majority of the biokinetic parameter values, most of which are described by Henze et al. (2000), within a range of values reported in the literature except for, as an example, the autotrophic nitrifier maximum growth rate (m). However, the mathematical description of some processes includes variable, or lumped, parameters. These parameter values are often system specific and subject to significant uncertainty. The lumped parameters account for an incomplete mechanistic description of the simulated process. Lumped parameters in a majority of biofilm models, including those described in this chapter, are the oxygen affinity constant for autotrophic nitrifiers (KO2,A), endogenous respiration rate constants (bres), attachment rate coefficient (kat), detachment rate coefficient (kdet), mass-transfer boundary layer thickness (LL), ratio of diffusion in biofilm to diffusion in water (Df/Dw). Indentifying parameter subsets that require definition for biofilm model calibration has been the subject of several investigations by Smets et al. (1999), Van Hulle et al. (2004), and Brockmann et al. (2008).
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model for another period. Similarly, Bilyk et al. (2008) reported the calibration of a denitrification filter model by adjusting assumed biofilm thickness and incorporating the assimilative denitrification reaction. Both of these biofilm reactor model calibration efforts were based on bulk-phase measurements, but only Sin et al. (2008) utilized measured characteristics of the biofilm. Such adjustments to systemspecific biofilm and biokinetic parameters in order to match observed data may not produce a properly calibrated model that is capable of describing a variety of design conditions for a wastewater-treatment plant. As previously discussed, the attachment coefficient, for example, has been experimentally demonstrated (and described mathematically) to change as a function of particle (total suspended solids) load (Boltz and La Motta, 2007). Then, it may be argued that adjusting the attachment coefficient (during calibration) to match an observed dataset would naturally render the calibrated model incapable of describing another scenario with a different particle load. Suffice it to say that a reliable and transparent description of recommended approaches for the application and calibration of biofilm models are required for the models to gain general acceptance, understanding, and become effectively used in engineering practice. Protocol defining methodology for sampling, testing, evaluating and applying data to mathematical biofilm reactor models is required. It is likely that existing biofilm reactor models will require improvement for reliable dynamic simulation in practice.
4.15.3.5 Biofilm Reactors in Wastewater Treatment 4.15.3.4.5 Calibration protocol Application of a dynamic biofilm model to describe full-scale municipal wastewater-treatment processes requires a calibration of the selected model. Ad hoc expert-based trial and error and standardized systematic approaches have been used to calibrate process models. Sin et al. (2005) presented a critical comparison of systematic calibration protocols for activated sludge models. These protocols have many similarities that are applicable to biofilm reactor models including goal definition, data collection/testing/reconciliation, and validation. The major differences between protocols reported by Sin et al. (2005) are related to the measurement campaign, experimental methods for influent wastewater characterization, and parameter subset selection and calibration. The major differences speak to areas of systematic calibration protocols for activated sludge models that will almost certainly be exasperated when creating systematic protocol for calibrating a biofilm reactor model. Certainly, additional tests will be required to characterize the physical attributes of both suspended growth and biofilm compartments, and mathematical biofilm models have more parameters than activated sludge models. Furthermore, the biofilm compartment parameters must be estimated from bulk-phase measurements in order to have a timely and costeffective approach to calibrating biofilm reactor models. Sin et al. (2008) reported the calibration of a dynamic biologically active (continuously backwashing) filter model using traditional expert-based manual trial and error. The researchers manipulated system-specific parameters related to attachment, detachment, and biofilm thickness. After calibration, Sin et al. (2008) successfully tested the calibrated
Biofilm reactors play an important role in environmental biotechnology, but many aspects of their design and operation remain poorly understood. Biofilm reactors can be traced to origination of modern water sanitation. Corbett (1903) reported the use of continuously distributed sewage flow over a fixed bed, and Stoddart (1911) reported the use of a coarse biofilm-covered medium dosed with a continuous trickling flow. These accounts are acknowledged as the creation of the TF process. Approximately 100 years following these reports significant advances in the design, academic understanding, and mathematical modeling of biofilms have led to the development of new and emerging biofilm reactors conducive to fundamentally based design approaches and the application of fundamentally based design and operation procedures for traditional biofilm reactors. Two processes – mass transfer and biochemical conversion – are characteristics of all biofilm reactors and influence biofilm structure and function. Compartments that are common to every biofilm reactor exist to optimize mass-transfer and biochemical conversion.
4.15.3.5.1 Biofilm reactor compartments Biofilm reactors have five primary compartments: (1) influent wastewater (distribution) system; (2) containment structure; (3) biofilm carrier; (4) effluent water collection system; and (5) an aeration system (for aerobic processes and scour) or mixing system (for anoxic processes that require bulk-liquid agitation and biofilm carrier distribution). Five components influence local conditions inside the biofilm: (1) biofilm carrier surface (i.e., substratum); (2) biofilm (including both particulate and liquid fractions); (3) mass-transfer boundary
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Biofilms in Water and Wastewater Treatment
layer; (4) bulk liquid; and (5) gas phase (when significant). The components typical of biofilm reactors are described in context of some commercially available biofilm reactors. The five biofilm reactors described include the MBBR, BAF, FBBR, rotating biological contactor (RBC), and TF.
4.15.3.5.2 Moving bed biofilm reactors The MBBR is a two- (anoxic) or three- (aerobic) phase system with a buoyant free-moving plastic biofilm carrier that requires mechanical mixing or aeration to distribute carriers throughout the tank. The process includes a submerged, completely mixed biofilm reactor and liquid–solids separation unit (Ødegaard, 2006). A range of pollutant loading and bulkphase external carbon sources in denitrification MBBRs and dissolved oxygen concentrations in carbon-oxidation and/or nitrification MBBRs have been applied, and system response evaluated (Lazarova and Manem, 1994). It has been demonstrated that MBBRs are capable of processing wastewater to meet a variety of effluent water-quality standards ranging, for example, from the US Environmental Protection Agency definition of secondary treatment (30 mg TSS l1 and 30 mg BOD5/l monthly average) to more stringent enhanced nitrogen removal limits (e.g., total nitrogen less than 3–5 mg l1) under a variety of loading conditions. The MBBR process is capable of meeting similar treatment objectives as the activated sludge process for carbon oxidation, nitrification, and denitrification, but the MBBR makes use of a smaller tank volume than a clarifier-coupled activated sludge system. Biomass retention is clarifier independent; therefore, solids loading in liquid–solids separation unit are significantly reduced when compared with the activated sludge process. Because it is a continuously flowing process, the MBBR does not require a special operational cycle for biofilm thickness control (e.g., backwashing in a BAF or flushing in a TF). Hydraulic head loss and operational complexity is minimal. The MBBR offers much of the same flexibility to manipulate the process flow sheet (to meet specific treatment objectives) as the activated sludge process. Multiple reactors can be configured in series without the need for intermediate pumping or return activated sludge pumping (to accumulate mixed liquor). Liquid–solids separation may be achieved with a variety of processes including sedimentation basins, dissolved air flotation, cloth-disk and membrane filters. The MBBR is well suited for retrofit installation into existing municipal wastewater-treatment plants. An MBBR may be a single reactor or several reactors in a series. Typically, each MBBR has a length-to-width ratio (L:W) in the range of 0.5:1–1.5:1. Plans with an L:W greater than 1.5:1 can result in nonuniform distribution of the biofilm carriers. MBBRs contain a plastic biofilm carrier volume up to 67% of the liquid volume. Screens are typically installed with one MBBR wall and allow treated effluent to flow to the next treatment step while retaining the free-moving plastic biofilm carriers. Aerobic MBBRs use a diffused aeration system to evenly distribute the plastic biofilm carriers and meet process oxygen requirements. On the other hand, anoxic MBBRs use mechanical mixers to evenly distribute the plastic biofilm carriers because there is no process oxygen requirements. Each process mechanical component is submerged. Figure 22
depicts the Williams-Monaco WWTP, Commerce City, Colorado, a two-train bioreactor that consists of four MBBRs in series. The biofilm carriers are extruded or molded from either virgin or recycled high-density polyethylene (HDPE). Table 2 summarizes characteristics of several commercially available plastic biofilm carriers. The carriers are slightly buoyant and have a specific gravity between 0.94 and 0.96 g cm3. Both native and biofilm-covered plastic biofilm carriers have a propensity to float in quiescent water. Biofilms primarily develop on the protected surface inside the plastic biofilm carrier. For this reason, the specific surface areas of plastic biofilm carriers listed in the table exclude areas not inside the plastic carrier. The listed bulk-specific surface area, which is based on 100% carrier fill, is characteristic of a plastic biofilm carrier. The net specific surface area is characteristic of plastic biofilm carrier and fill percentage. For example, if a plastic biofilm carrier has a 500 m2 m3 bulk-specific surface area, then the net specific surface area at 50% carrier fill is 250 m2 m3. Similarly, the net liquid volume displacement at 50% carrier fill is 0.0725 for a plastic biofilm carrier having a characteristic 0.15-bulk-liquid volume displacement (at 100% carrier fill). Plastic biofilm carriers are retained in an MBBR by horizontally configured cylindrical screens or vertically configured flat screens as shown in Figure 23. Aerobic zones typically contain cylindrical screens; anoxic zones contain the flat wall screens. Cylindrical screens are desired. They extend horizontally into the upward-flowing air bubbles imparted by the diffuser grid which aids in scouring any accumulated debris. Energy imparted by the mechanical mixers is insufficient to dislodge debris accumulated on the flat wall screen. Therefore, scouring of flat screens is accomplished with a sparging air header in a denitrification MBBR. Removing the debris retained on a screen aids in maintaining hydraulic throughput. Hydraulically, an MBBR is commonly designed to process a maximum approach velocity (based on the tank cross-sectional area perpendicular to forward flow) in the range 30– 35 m h1. Screen area is defined by the maximum allowable head loss through the screens, which is typically in the range of 5–10 cm. The screen superficial hydraulic load is typically in the range of 50–55 m h1 for average design conditions. The screens and their supporting structural assemblies, if required, are typically constructed from stainless steel and may be from wedge-wire mesh or perforated plates. Low-pressure diffused air is applied to aerobic MBBRs. The airflow enters the reactor through a network of air piping and diffusers that are attached to the tank bottom. Airflow has the dual purpose of meeting process oxygen requirements and uniformly distributing plastic biofilm carriers. To promote uniform distribution of the plastic biofilm carriers, the diffuser grid layout and drop pipe arrangement provide a rolling water circulation pattern. Coarse-bubble diffusers are typically used in moving bed reactors (Figure 25). Coarse-bubble diffusers typically used in MBBRs are stainless steel pipes with circular orifices along the underside. These coarse-bubble diffusers are less affected by scaling and fouling because of the large dimension and turbulent airflow through the discharge orifice (Stenstrom and Rosso, 2008). As a result, coarse-bubble diffusers require less maintenance than fine-bubble diffusers. The coarse-bubble diffusers are designed with a structural end
Biofilms in Water and Wastewater Treatment
551
Aerated reactor #2
Aerated reactor #1
RECIR
Mixer Mixed bed reactor #2
Screen
Effluent overflow Effluent
Airflow distribution area
RECIR pump Effluent basin
RECIR
RECIR
Mixed bed reactor #1
Effluent
Influent Influent splitter box
Aerated reactor #3
RECIR
Aerated reactor #4
(a)
Mixed bed reactor #4
Mixed bed reactor #3
RECIR
Effluent
RECIR
(b)
Figure 22 (a) Moving bed biofilm reactor at the Williams-Monaco Wastewater Treatment Plant, Colorado, USA. (b) Schematic representation of the photographed system which illustrates the system consisting of two parallel trains each with four reactors in series.
support that enables them to withstand the weight of plastic biofilm carriers when the MBBR is out of service and drained. Denitrification MBBRs use mechanical mixers to agitate the bulk of the liquid and to distribute plastic biofilm carriers uniformly throughout the tank. The mechanical mixers are typically rail-mounted submersible (wet motor) units. Stateof-the-art submersible mechanical mixers typically have a maximum 120-rpm impeller speed and a minimum of three blades per impeller. The mixer uses a stainless steel backwardcurve propeller with a round bar welded along its leading edge to avoid damage to the plastic biofilm carriers and impeller wear. The mixer has a large diameter impeller with a fairly low rotational speed (90 rpm at 50 Hz and 105 rpm at 60 Hz). The plastic biofilm carriers float in quiescent water. As a result, the mixers need to be located near the water surface but not so close as to create an air-entraining vortex. A slight negative
inclination of mixer orientation helps maintain the rollingwater circulation pattern and uniformly distribute plastic biofilm carriers (see Figure 24). Rail-mounted units facilitate access to the mixer when maintenance is required. The mixers are typically sized to input 25 W m3 of reactor volume. Carbon-oxidizing MBBRs are classified as low-rate, normalrate, or high-rate bioreactors. Low-rate carbon-oxidizing MBBRs promote conditions for nitrification in downstream reactors. High- and normal-rate MBBRs are strictly carbon-oxidizing bioreactors. In the absence of site-specific pilot-scale observations or a calibrated mathematical model, high-rate MBBRs are typically designed to receive a filtered BOD5 load in the range of 15–20 g m2 d1 at 15 1C. This corresponds to total BOD5 loads as high as 45–60 g m2 d1 at 15 1C (Ødegaard, 2006). To reach secondary treatment effluent standards, a hydraulic residence time less than 30 min is not
552
Biofilms in Water and Wastewater Treatment
Table 2
Moving bed biofilm reactor plastic biofilm carrier characteristicsa
Manufacturer
Name
Bulk specific surface area, weight, gravity
Nominal carrier dimensions (depth; diameter)
Veolia Inc.
AnoxKaldnesTM K1
500 m2 m3 145 kg m3 0.96–0.98
7.2 mm; 9.1 mm
AnoxKaldnesTM K3
500 m2 m3 95 kg m3 0.96–0.98
10 mm; 25 mm
AnoxKaldnesTM Biofilm Chip (M)
1,200 m2 m3 234 kg m3 0.96–1.02
2.2 mm; 45 mm
AnoxKaldnesTM Biofilm Chip (P)
900 m2 m3 173 kg m3 0.96–1.02
3 mm; 45 mm
ActiveCellTM 450
450 m2 m3 134 kg m3 0.96
15 mm; 22 mm
ActiveCellTM 515
515 m2 m3 144 kg m3 0.96
15 mm; 22 mm
ABC4TM
600 m2 m3 150 kg m3 0.94–0.96
14 mm; 14 mm
ABC5TM
660 m2 m3 150 kg m3 0.94–0.96
12 mm; 12 mm
BioPortzTM
589 m2 m3
14 mm, 18 mm
Infilco Degremont Inc.
Aquise
Entex Technologies Inc.
Carrier photo
a
As reported by manufacturer. Modified from Boltz JP, Morgenroth E, deBarbadillo C, et al. (2010b) Biofilm reactor technology and design. In: Design of Municipal Wastewater Treatment Plants, WEF Manual of Practice No. 8, ASCE Manuals and Reports on Engineering Practice No. 76, 5th edn, vol. 2, ch. 13, (ISBN P/N 978-0-07-166360-1 of set 978-0-07-166358-8; MHID P/N 0-07166360-6 of set 0-07-166358-4). New York: McGraw-Hill.
recommended. Medium-rate MBBRs designed for meeting basic secondary treatment standards are typically designed for a loading of 5–10 g BOD5 m2 d1 at 10 1C, depending on the choice of liquid–solids separation process. Values in the higher range are used when coagulation occurs before the separation unit; values in the lower range are used without coagulation. Studying a pilot-scale combined carbon oxidation and nitrification MBBR receiving primary effluent and a (tertiary) nitrification MBBR receiving secondary effluent while maintaining a 4–6 g m3 bulk-liquid dissolved-oxygen concentration in both units, Hem et al. (1994) observed that a total BOD5 load of 1–2 g m2 d1 resulted in nitrification rates
from 0.7 to 1.2 g m2 d1, a total BOD5 load of 2–3 g m2 d1 resulted in nitrification rates from 0.3 to 0.8 g m2 d1, and a total BOD5 load greater than 5 g m2 d1 resulted in virtually no nitrification.
4.15.3.5.3 Biologically active filters BAFs have natural mineral, structured or random plastic media that supports biofilm growth and serves as a filtration medium. Solids accumulated from filtration and biochemical transformation are removed by backwashing. Media density influences BAF configuration and backwash regimes. BAF
Biofilms in Water and Wastewater Treatment
553
(a)
(b)
Figure 23 (a) Horizontal cylindrical screens constructed of wedge wire. Stainless steel coarse-bubble diffusers typically used in aerobic MBBRs are also pictured on the tank floor. (b) Flat wall screen constructed of wedge wire. A single air-header is pictured. Air is periodically introduced to scour debris accumulated on the screen.
A
B
30°
D (a)
(b)
Figure 24 (a) Schematic and (b) picture of mechanical mixers that are specially designed for anoxic moving bed biofilm reactors.
influent requires preliminary and primary treatment. Historically, the acronym BAF has meant biological aerated filters which have been used to refer to aerated biofilters used for secondary treatment. However, Boltz et al. (2010b) revised the acronym BAF to cover all BAFs, including those that operate under anoxic conditions for denitrification. BAFs are characterized by their media configurations and flow regime.
Downflow BAFs with media heavier than water include the Biocarbones process, which was marketed during the 1980s for secondary and tertiary treatment, and packed-bed tertiary denitrification filters such as the Tetra Denites process. These BAFs are backwashed using an intermittent counter-current flow. Upflow BAFs with media heavier than water such as the Infilco Degremont Biofors process have been used for
554
Biofilms in Water and Wastewater Treatment
secondary and tertiary treatment. The systems make use of expanded clay or another mineral media. These BAFs are backwashed using an intermittent concurrent flow. BAFs with floating media such as the Veolia Biostyrs process have also been used for secondary and tertiary treatment, and uses polystyrene, polypropylene, or polyethylene media. These BAFs operate with an intermittent backwash counter-current flow. Continuous backwashing filters operate in an upflow mode and contains media that is heavier than water. The media continuously moves counter-current to the wastewater flow (i.e., downward), and is continuously channeled to a center air lift where it is scoured, rinsed, and returned to the top of the media bed. Nonbackwashing submerged filters consist of a submerged static media bed, and have been called submerged aerated filters (SAFs). Solids are not retained in these filters. Therefore, nonbackwashing submerged filters require a dedicated liquid–solids separation process. A downflow BAF with media heavier than water, such as the Tetra Denites filter, is illustrated in Figure 25. The
Denites process has been used since the late 1970s for meeting stringent total nitrogen limits while providing a filtered effluent. Methanol or another external carbon source is added to the influent wastewater stream to promote biological denitrification. A typical installation includes 1.8 m of 2–3 mm diameter sand media over 457 mm of graded support gravel. In a downflow denitrification BAF, the backwash cycle typically consists of a brief air scour followed by an air–water backwash and water rinse cycle. Backwash water and air scour flow rates are typically 15 and 90 m3 m2 h1, respectively. Backwash water usage is typically 2–3% of the average flow being treated. Nitrogen gas accumulates in the media. A releasing mechanism is pumping backwash water up through the media bed for a short duration. The denitrification capacity between nitrogen release cycles typically ranges from 0.25 to 0.5 kg NOX-N m2. An upflow BAF with media heavier than water, such as the Infilco Degremont Biofors, is illustrated in Figure 26. The Degremont Biofors operates such that solids are trapped Proces air
Raw water
Air
Backwash water extraction
Water Biofilter media Support layer
Air scour Backwash water Treated water Figure 25 Downflow BAF with media heavier than water (e.g., Biocarbones and Tetra Denites). From ATV (1997) Biologische und weitergehende Abwasserreinigung (German), 4th edn. Berlin: Ernst and Sohn as presented by Tschui (1994).
Water
Process air
Biofilter media
Backwash water extraction
Air
Support layer
Air scour Treated water Backwash water
Raw water Figure 26 Upflow BAF with media heavier than water (e.g., Infilco Degremont Biofors). From ATV (1997) Biologische und weitergehende Abwasserreinigung (German), 4th edn. Berlin: Ernst and Sohn as presented by Tschui (1994).
Biofilms in Water and Wastewater Treatment
mostly in the lower part of the filter medium during normal operation and are removed through backwashing and applying scour air. As the backwash consists of concurrent scour air and backwash water, accumulated solids travel up through the media bed before being released at the top. Three types of media can be used in the Biofors depending on the application; the media types include expanded clay, expanded shale (both in the form of spherical grains with an effective size of 3.5 or 4.5 mm), and angular grains (with an effective size of 2.7 mm). The media form a submerged, fixed bed in the bottom of the reactor. The media bed typically has a height of 3–4 m with approximately 1-m freeboard. The grains-specific surface area is approximately 1640 m2 m3. Influent water to the bed flows through a plenum and nozzle air/water distribution system. The nozzles are installed in a false floor located approximately 1 m above the filter floor. Nozzles in the false floor are subject to clogging. Therefore, backwash water and scour air flow through the same plenum/nozzle system. Process air is introduced through separate air diffusers located in the media bed above the inlet nozzles. A key issue with the backwash of sunken media systems is the potential for boils during backwashing. The flow will short-circuit through the line of least resistance. This will result in a boil, or violent eruption of the flow through the point of least resistance. Similar short circuits and boils can also occur if the nozzles are blocked. These boils can result in excessive media loss during backwashing. Therefore, to achieve even backwashing the water must be well distributed across the BAF plan area. Therefore, the headloss across the distribution system must be greater than the headloss through the bed. An upflow BAF with floating media, such as the Veolia Biostyrs, is illustrated in Figure 27. These processes use a floating bed of media to provide area for biofilm development and filtration. Coarse-bubble aeration diffusers exist at the bottom of the media to enhance the contact of air, water, and biomass (Rogalla and Bourbigot, 1990). The Biostyrs process uses light weight expanded polystyrene (specific gravity of 0.05). Alternatively, the Biobeads process uses
555
recycled polypropylene with a specific gravity slightly lower than 1. The Biostyrs reactor is partially filled with (2–6 mm) polystyrene beads. Process objectives determine selection of the bead size; larger beads can be more heavily loaded. The beads, which are lighter than water, form a floating bed in the upper portion of the reactor, typically a height of 3–4 m with approximately 1.5 m of freeboard. The top of the bed is restrained by a slab fitted with filtration nozzles to evenly collect treated wastewater. The clean specific surface area of spherical beads is 1000–1400 m2 m3. In the bottom of the reactor, influent is distributed by troughs formed in the base of the cells. Process air is distributed through diffusers located along the bottom of the reactor or within an aeration grid in the media bed. The latter is used if an anoxic zone is required for denitrification. Backwashing consists of counter-current air scour and backwash water flow. The Biobeads BAF process is similar to Biostyrs, except that the media is larger and heavier, using polypropylene or polyethylene with a density of approximately 0.95. To prevent media attrition, a metal grid is fixed near the top of the reactor. Upflow floating media BAFs may also require a certain number of mini-backwashes (typically 4–8 and, in extreme cases, more than 10) to bump the filter, remove some solids, and lower headloss to achieve a complete filtration cycle of 24 or 48 h (which is the time between normal backwashes). The requirement for minibackwashes plus normal backwashes can generate a significant backwash water volume. During demonstration testing in San Diego, California, USA, a single-stage carbon-oxidation BAF with floating media generated a backwash water volume in the range of 10.3–13.9% of influent flow, compared to a sunken media BAF which produced a backwash water volume in the range of 7.4–7.9% (Newman et al., 2005). An upflow continuous backwash BAF, such as the Parkson Dynsands, is illustrated in Figure 28. Moving bed, continuous backwash filters operate in an upflow mode and consist of media heavier than water. The media continuously moves downward, counter-current to the wastewater flow. These filters are used widely for tertiary suspended solids and turbidity Backwash water
Air scour Process air
Air Aerobic filter zone
Treated water
Anoxic filter zone Water
Recirculation pump Raw water
Backwash water extraction Figure 27 Upflow BAF with floating media (e.g., Veolia Biostyrs). Adapted from ATV (1997) Biologische und weitergehende Abwasserreinigung (German), 4th edn. Berlin: Ernst and Sohn as presented by Tschui (1994).
556
Biofilms in Water and Wastewater Treatment Central reject compartment (H)
Feed (influent) (A)
Rejects (L) Top of airlift pump (G) Filtrate weir (J)
Reject weir (K) Sand washer (L) Effluent (E)
Downward moving sand bed (D)
Downward feed (B) Feed radials (C)
Bottom of airlift pump (F) Figure 28 Parkson Dynasands process schematic, continuous backwash BAF.
removal but have also been applied to separate stage nitrification and denitrification. Two commercially available systems using this technology are the Parkson DynaSands and Paques Astrasands filters. The filter cells are supplied as 4.65-m2 modules with center airlift assembly. The effective media depth is typically 2 m, and sand media size typically ranges from approximately 1 to 1.6 mm. Influent wastewater enters the filter bed through radials located at the bottom of the filter. The flow moves up through the downward-moving sand bed and effluent flows over a weir at the top of the filter. The media, with the accumulated solids, is drawn downward to the bottom cone of the filter. Compressed air is introduced through an airlift tube extending to the conical bottom of the filter and rises upward with a velocity exceeding 3 m s1 creating an air pump that lifts the sand at the bottom of the filter through the center column. The turbulent upward flow in the airlift provides scrubbing action that effectively separates solids from the media before discharge to a wash box. There is a constant upward flow of liquid into the wash box (backwash water) controlled by the wash box discharge weir. Moving bed filter manufacturers typically set the reject weir to provide a wash water flow rate equivalent to approximately 10% of the forward flow at an average filter loading rate of 4.9 m h1. The backwash frequency is quantified by the bed turnover rate. To maintain sufficient biomass for denitrification, the bed turnover rate must be reduced to approximately 100–250 mm h1.
Several media types are available for use in BAFs. Media selection is integral to meeting treatment objectives, flow and backwashing regimes. Typically, media can be categorized as mineral media and plastic media. In most cases, mineral media is denser than water and plastic media is buoyant. The media needs to resist breakdown from abrasion during backwashing and chemical degradation by constituents in municipal wastewater. Commercially available BAF systems and their media are listed in Table 3. Backwashing BAFs maximizes solids capture and filter run time. Proper backwashing requires filter bed expansion and rigorous scouring followed by efficient rinsing. Accumulation of solids and media (mud balling) results in wastewater short-circuiting and can result in excessive media loss. Feed characteristics and type of treatment provided by the BAF affect solids production and frequency requirements for backwashing. Biomass yield in tertiary BAF systems is typically low, so backwashing is relatively infrequent (i.e., one backwash per 36–48 h). Reactor characteristics and media type influence backwash frequency. More openly structured media capture fewer solids which reduces backwash frequency. During backwashing the media bed is typically expanded or fluidized (depending on the system) to allow for grain separation and free movement in order to remove as much accumulated solids as possible. Table 4 compares typical BAF backwashing requirements. BAFs designed for carbon oxidation and suspended solids removal in secondary treatment typically have volumetric BOD loading rates in the range of 1.5–6 kg m3 d1. Average and peak HLRs for secondary and tertiary treatment systems are typically in the range of 4–8 and 10–20 m h1, respectively. As BAFs for secondary treatment are typically placed immediately downstream of primary clarifiers, the applied volumetric mass loading rate is almost always the limiting design parameter. Combined carbon oxidation and nitrification will proceed when the organic loading at lower temperatures is limited to 2.5 kg BOD m3 d1 (Rogalla et al., 1990). Under these conditions a total Kjeldahl nitrogen removal rate of 0.4 kg N m3 d1 may be achieved. Inversely, Rogalla et al. (1990) found that nitrification decreases when soluble COD loadings approach 4 kg m3 d1. Ammonium removal of 80– 90% can be achieved for ammonium loads in the range of 2.5–2.9 kg m3 d1 (Peladan et al., 1996).
4.15.3.5.4 Expanded and fluidized bed biofilm reactors Expanded bed biofilm reactors (EBBRs) and FBBRs use small media particles that are suspended in vertically flowing wastewater, so that the media becomes fluidized and the bed expands. Individual particles become suspended once the drag force of the relatively fast flowing wastewater (30–50 m h1) overcomes gravity and they are separated. In municipal applications, fluidized beds are typically used for tertiary denitrification. Design criteria for denitrifying FBBRs are listed in Table 5. When treating groundwater or industrial wastewater, FBBRs are used for the removal of oxidized contaminants such as nitrate and perchlorate. Suspension of the media maximizes the contact surface between microorganisms and wastewater. It also increases treatment efficiency by improving mass transfer because there
Biofilms in Water and Wastewater Treatment Table 3
557
Biologically active filter systems and commercially available media
Process
Supplier
Flow regime
Media
Specific gravity
Size (mm)
Astrasands Biobeads Biocarbones Biofors Biolest Biopur
Paques/Siemens Brightwater F.L.I. OTV/Veolia Degremont Stereau Sulzer/Aker Kvaerner Kruger/Veolia Severn Trent Severn Trent Parkson FB Leopold Severn Trent
Upflowa Upflow Downflow Upflow Upflow Downflow
Sand Polyethylene Expanded shale Expanded clay Pumice/pouzzolane Polyethylene
42.5 0.95 1.6 1.5–1.6 1.2
1–1.6
Upflow Upflow Downflow Upflowa Downflow Up/down
Polystyrene Sand Sand Sand Sand Slag
0.04–0.05 2.6 2.6 2.6 2.6 2–2.5
3.3–5 2–3 2–3 1–1.6 2 28–40
Washed gravel
2.6
19–38
Biostyrs ColoxTM Denites Dynasands Eliminites Submerged activated filter
2–6 2.7, 3.5, and 4.5
Specific surface area (m2 m3)
1400–1600
Structured 1000 656 656
240
a
Moving bed. From Boltz JP, Morgenroth E, deBarbadillo C, et al. (2010b) Biofilm reactor technology and design. In: Design of Municipal Wastewater Treatment Plants, WEF Manual of Practice No. 8, ASCE Manuals and Reports on Engineering Practice No. 76, 5th edn, vol. 2, ch. 13 (ISBN P/N 978-0-07-166360-1 of set 978-0-07-166358-8; MHID P/N 0-07-166360-6 of set 0-07-166358-4). New York: McGraw-Hill.
Table 4
Summary of biologically active filter (BAF) backwashing (BW) requirements
Upflow, sunken media Normal BW Energetic BWa Upflow, floating media Normal BW Mini-BWb Downflow, sunken media Upflow, moving bedf
Backwash water rate, m h1
Air scour rate, m h1
Total duration minc
Total backwash water volume per cellc
Total backwash water volume per celld
20 (8.2) 30 (12.3)
97 (5.3) 97 (5.3)
50 25
9.2 m3 m2 9.2 m3 m2
12 m3 m2 10 m3 m2
55 (22.5) 55 (22.5) 15 (6) 0.5–0.6
12 (0.65) 12 (0.65) 90 (5) Continuous through air lift
16 5 20–25 Continuous
2.5 m3 m3 mediae 1.5 m3 m3 mediae 3.75–5 m3 m2 55–67 m3 d1
2.5 m3 m3 mediae 1.5 m3 m3 mediae 3.75–5 m3 m2 55–67 m3 d1
(0.2–0.24) a
Energetic backwash once every 1–2 months depending on trend in ‘‘clean bed’’ headloss following normal backwash. Mini-backwash applied as interim measure when pollutant load exceeds design load. c Backwash duration reflects total duration of the typical backwash cycle, which includes valve cycle time and pumping and nonpumping steps. The duration of each step is adjustable via programmable logic controller and supervisory control and data acquisition control systems. d The total backwash water volume includes drain and filter to waste steps, where applicable. e Backwash volume requirements for upflow floating media BAF typically are based on media volume rather than cell area because depths vary. f Continuous backwash filter BW is based on a standard 4.65 m2 cell and a typical weir setting for reject flow of approximately 2.3–2.8 m3 h1 cell1. From Boltz JP, Morgenroth E, deBarbadillo C, et al. (2010b) Biofilm reactor technology and design. In: Design of Municipal Wastewater Treatment Plants, WEF Manual of Practice No. 8, ASCE Manuals and Reports on Engineering Practice No. 76, 5th edn, vol. 2, ch. 13 (ISBN P/N 978-0-07-166360-1 of set 978-0-07-166358-8; MHID P/N 0-07-166360-6 of set 0-07-166358-4). New York: McGraw-Hill. b
is significant relative motion between the biofilm and flowing wastewater. Because of the balance of forces involved in particle fluidization and bed expansion, the smallest particles are found at the top and the largest at the bottom of the fluid bed. Therefore, the media particles should be graded to a relatively tight size range. The degree of bed expansion determines whether a bed is deemed expanded or fluidized. The transition lies between 50% and 100% expansion over the static bed height. This discussion assumes the upper limit: beds less than double static bed height
(o100% expanded) are considered expanded; those more than double the static bed height (4100% expanded) are fluidized. A lower degree of bed expansion is advantageous, because it requires a lower flow velocity, less energy, and increases effective biomass concentration, which reduces the reactor footprint. In aerobic processes, however, it increases volumetric oxygen demand because of increased biomass concentration. The FBBR/EBBR is illustrated in Figure 29. The system consists of a column in which the particles are fluidized and a
558
Biofilms in Water and Wastewater Treatment
Table 5
Design criteria for denitrifying fluidized bed biofilm reactors
Parameter
Value
Packing Type Effective size Sphericity Uniformity coefficient Specific gravity Initial depth Bed expansion Empty-bed upflow velocity Hydraulic loading rate Recirculation ratio NO 3 N loading: 13 1C 20 1C Empty-bed contact time C:N (methanol) Specific surface areaa Biomass concentrationa
Unit
Range
Typical
mm Unitless Unitless Unitless m % m h1 m3effluent m2 Bioreactor Unitless
Sand 0.3–0.5 0.8–0.9 1.25–1.50 2.4–2.6 1.5–2.0 75–150 36–42 400–600 2:1–5:1
Sand 0.4 0.8–0.85 r1.4 2.6 2.0 100 36 500 3:1
2.0–4.0 3.0–6.0 10–20 3.0–3.5 1000–3000 15 000–40 000
3.0 5.0 15 3.2 2000 30 000
area
kg m3 d1 kg m3 d1 min Unitless m2 m3 mg l1
d1
a
Specific surface area range based on sand particles; alternate media used in fluidized bed reactors such as carbon or glassy coke may have a different specific surface area range. From Boltz JP, Morgenroth E, deBarbadillo C, et al. (2010b) Biofilm reactor technology and design. In: Design of Municipal Wastewater Treatment Plants, WEF Manual of Practice No. 8, ASCE Manuals and Reports on Engineering Practice No. 76, 5th edn, vol. 2, ch. 13 (ISBN P/N 978-0-07-166360-1 of set 978-0-07-166358-8; MHID P/N 0-07-166360-6 of set 0-07-166358-4). New York: McGraw-Hill.
Figure 29 Fluidized bed biofilm reactor process flow diagram (Shieh and Keenan, 1986).
Process flow enters at the bottom of the reactor and flows through a distribution system to ensure even dispersion and uniform fluidization. Silica sand (0.3–0.7 mm diameter) and granular activated carbon (GAC; 0.6–1.4 mm) are typically used. Other materials, however, have been used at pilot scale, such as 0.7–1.0 mm glassy coke (McQuarrie et al., 2007). Small carrier particles (1 mm) provide a large specific surface area for biofilm growth (up to 2400 m2 m3 when expanded 50%), which is one of the key advantages of this process technology. In a study of tertiary nitrification of activated sludge-settled effluent using a pilot-scale EBBR, Dempsey et al. (2006) found that the process also removed up to 56% CBOD and 62% TSS from the influent stream. Removal of these materials was attributed to the activities of protozoa (free-living and stalked) and metazoa (rotifers, nematodes, and oligochaetes) as shown in Figure 30.
recycle line that is used to maintain a fixed, vertical hydraulic flow. In this way, bed expansion is kept constant and biofilm covered particles are retained independent of influent flow. Aeration typically is achieved during recycle, during which influent wastewater mixes with effluent recycled from the top of the bed. If aeration is conducted within the fluidized bed, then a significant volume of gas disturbs the fluidized state by causing turbulence and increased force of interparticle collisions. This can cleave biofilm from the substratum. Nevertheless, this approach has been used. The advantage of adding air to the recycle stream is that biomass is not stripped from the media by turbulence of rising gas bubbles; therefore, the treated effluent typically has a lower suspended concentration (Jeris et al., 1981).
The RBC process has been applied where average effluent water-quality standards are less than or equal to 30 mg l1 BOD5 and TSS. The RBC employs a cylindrical, synthetic media bundle that is mounted on a horizontal shaft. Figure 31 illustrates the shaft-mounted media. The bundled media is partially submerged (typically 40%) and slowly (1–1.6 rpm) rotates to expose the biofilm to substrate in the bulk of the liquid (when submerged), and to air (when not submerged). Detached biofilm fragments suspended in the RBC effluent stream are removed by liquid– solids separation units. The RBC process is typically configured with several stages operating in series. Each reactorin-series may have one or more shafts. Parallel trains are
Effluent
Excess biomass
Recycle Separator
Influent
Media Reactor Bioparticle
O2 Chemicals (optional) 1 2
1 Medium 2 Biofilm
4.15.3.5.5 Rotating biological contactors
Biofilms in Water and Wastewater Treatment
(a)
(b)
(c)
(d)
(e)
(f)
559
Figure 30 Particulate biofilms with associated protozoa and metazoan from expanded bed: (a) bioparticles in expanded bed; (b) bioparticles with surface attached; (c) closeup of rotifer attached to bioparticle; (d) stalked protozoa on surface of particulate biofilms; (e) testate amoeba grazing on biofilm; and (f) oligochaete worm grazing on bioparticles (Dempsey et al., 2006).
implemented to provide additional surface area for biofilm development. Media-supporting shafts typically are rotated by mechanical drives. Diffused air-drive systems and an array of airentraining cups that are fixed to the periphery of the media (to capture diffused air) have been used to rotate the shafts. RBCs have failed as a result of shaft, media, or media support system structural failure; poor treatment performance; accumulation of nuisance macrofauna; poor biofilm thickness control; and inadequate performance of air-drive systems for shaft rotation. Typically, the RBC tank is sized at 4.9 103 m3 m2 of media for low-density units. Disks typically have a 3.5-m diameter and are situated on a 7.5-m-long rotating shaft. The RBCs may contain low- or high-density media. Low-density media has a 118-m2 m3 biofilm active specific surface; high-density units have 180 m2 m3. Low-density media typically are used in the first stages of RBC systems which are designed for BOD5 removal to reduce potential media clogging and weight problems resulting from substantial biofilm accumulation. High-density media typically is used for nitrification. Mechanical shaft drives consist of an electric motor, speed reducer, and belt or chain drive. Typically, 3.7-kW mechanical drives have been provided for full-scale RBCs. Air-driven shafts require a remote blower for air delivery. Air headers are equipped with coarse-bubble diffusers. The air flow rate is typically in the range of 4.2–11.3 m3 min1 per shaft. Air quantity required by systems using air-driven shaft rotation, however, must be evaluated on a site-specific basis. Mechanical drive units have been designed for operation from 1.2 to 1.6 rpm. Air-drive units have been designed for 1.0–1.4 rpm. Ideally, shaft rotational speed is consistent. The development of an evenly distributed biofilm is desirable to avoid an uneven weight distribution, which may cause cyclical loadings in mechanical-drive systems and loping (uneven rotation) in air-driven shaft rotating systems. A loping condition often
accelerates rotational speed and, if not corrected, may lead to inadequate treatment and the inability to maintain shaft rotation. Air-drive systems should provide ample reserve air supply to maintain rotational speeds, restart stalled shafts, and provide short-term increased speeds (2–4 times normal operation) to control excessive or unbalanced biofilm thicknesses. Available data indicate that in excess of an 11.3-m3 min1 airflow rate per shaft may be required to maintain a 1.2-rpm shaft rotational speed during peak organic loading conditions (Brenner et al., 1984). Large-capacity air cups (150 mm diameter) typically are provided in the first stage of the process to exert a greater torque on the shaft and reduce loping. The RBC process is typically covered to avoid ultraviolet (UV) light-induced media deterioration and algae growth, to prevent excessive cooling, and to provide odor control. RBCs have been installed in buildings or under prefabricated fiberglass-reinforced plastic (FRP) covers (as pictured in Figure 31).
4.15.3.5.6 Trickling filters The TF is a three-phase biofilm reactor with fixed carriers. Wastewater enters the bioreactor through a distribution system, trickles downward over the biofilm surface, and air moves upward or downward in the third phase where it diffuses through the flowing liquid and into the biofilm. TF components generally include an influent water distribution system, containment structure, rock or plastic media, and underdrain and ventilation system. Wastewater treatment using the TF results in a net production of total suspended solids. Therefore, liquid–solids separation is required, and is typically achieved with circular or rectangular secondary clarifiers. The TF process generally includes an influent/ recirculation pump station, the TF(s), and liquid–solids separation unit(s).
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Figure 31 Photograph of the Envirexs rotating biological contactor cylindrical synthetic media bundle mounted on a horizontal shaft (a) and rotating biological contactor covers (b). Photographs courtesy Siemens Water Technologies.
Figure 32 (a) Hydraulically driven rotary distributors use variable frequency drive controlled gates that either open or close distributor orifices which adjust with varying pumped flow rates to maintain a constant preset rotational speed. (b) Electrically driven rotary distributor. Photographs courtesy WesTech, Inc.
Primary effluent or screened and degritted wastewater is either pumped or flows by gravity to the TF distribution system. Essentially, there are two types of TF distribution systems: fixed-nozzle and rotary distributors. Because their efficiency is poor, distribution with fixed nozzles should not be used (Harrison and Timpany, 1988). Rotary distributors may be hydraulically or electrically driven. A properly designed rotary distribution system allows for effective media wetting and the intermittent application of wastewater to biofilm carriers. The intermittent application of influent wastewater allows the biofilm to have periods of resting which primarily serves as a process aeration mechanism. Poor media wetting may lead to dry pockets, ineffective treatment zones, and odor. An electrically or modern hydraulically driven rotary distributor
(Figure 32) controls rotational speed independent of the influent wastewater flow rate, and may be used to maintain the desired hydraulic dosing rate. Ideal TF media provides a high specific surface area, low cost, high durability, and high enough porosity to avoid clogging and promote ventilation (Metcalf and Eddy, 2003). TF media types include rock (RO), random (RA) (synthetic), vertical flow (synthetic) (VF), and cross-flow (synthetic) (XF). Both VF and XF media are constructed with smooth and/or corrugated plastic sheets. Another commercially available synthetic media, although not commonly used, is vertically hanging plastic strips. Horizontal redwood or treated wooden slats have also been used, but are generally no longer considered viable because of high cost or limited supply. Modules
Biofilms in Water and Wastewater Treatment
of plastic sheets (i.e., self-supporting VF or XF modules) are used almost exclusively for new and improved TFs, but several TFs with rock media exist, and have proven capable of meeting treatment objectives when properly designed and operated. Table 6 compares the characteristics of some TF media. The higher specific surface area and void space in modular synthetic media allow for higher hydraulic loading, enhanced oxygen transfer, and biofilm thickness control in comparison to rock media. Rock media has, ideally, a 50-mm diameter, but may range in size. Due to structural requirements associated with the large unit weight of rock, rock-media TFs are shallow in comparison to synthetic-media TFs. Their large surface area makes them more susceptible to excessive cooling. Generally, rock media is considered to have a low specific surface area, void space, and high unit weight. Although recirculation is common, the low void ratio in rock-media TFs limits hydraulic application rates. Excessive hydraulic application can result in ponding, limited oxygen transfer, and poor bioreactor performance. Performance of existing rock-media TFs may sometimes be improved by providing mechanical ventilation, solids contact channels, and/or deepened secondary clarifiers that include energy dissipating inlets and flocculator-type feed wells. Grady
Table 6
et al. (1999) suggested that under low organic loading (i.e., o1 kg BOD5 d1 m3) rock- and synthetic-media TFs are capable of equivalent performance. However, as organic loading increases, synthetic-media TFs are less susceptible to operational problems and have reduced potential for plugging. Synthetic TF media has a higher specific surface area and void space, and lower unit weight than rock media. Modular synthetic media is generally manufactured with the following specific surface areas: 223 m2 m3 as high density, 138 m2 m3 as medium density, and 100 m2 m3 as low density. Both VF and XF media are reported to remove BOD5 and NH3–N (Harrison and Daigger, 1987), but sufficient scientific evidence exists to surmise that there is a difference in the treatment efficiency of TFs constructed with XF and VF media even when manufactured with virtually identical specific surface areas. Plastic modules with a specific surface area in the range of 89–102 m2 m3 are well suited for carbon oxidation and combined carbon oxidation and nitrification. Parker et al. (1989) recommended medium-density XF media against the use of high-density XF media in nitrifying TFs. This is supported by observations from a pilot-scale nitrifying TF application data and conclusions of Gujer and Boller (1983, 1984)
Properties of some trickling filter media Nominal size (m)
Bulk density (kg m3)
Specific surface area (m2 m3)
Void space (%)
0.024–0.076
1442
62
50
0.076–0.128
1600
46
60
0.61 0.61 1.22
24–45
100, 138, and 223
95
Vertical flow
0.61 0.61 1.22
24–45
102 and 131
95
Randomb
0.185 ø 0.051 H
27
98
95
Media type Rock River
Slag
Plastica Cross flow
a
561
Material
Manufacturers of modular plastic media: (formerly) BF Goodrich, American Surf-Pac, NSW, Munters, (currently) Brentwood Industries, Jaeger Environmental, and SPX Cooling. Manufacturers of random plastic media: (formerly) NSW Corp. and (currently) Jaeger Environmental.
b
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Biofilms in Water and Wastewater Treatment
which show lower nitrification (flux) rates for lower-density modular synthetic media. The researchers claim that lower rates occur with high-density media due to the development of dry spots below the flow interruption points (i.e., higherdensity media has more flow interruptions and, therefore, less effective wetting). Using medium-density media also reduces plugging potential. Vertically oriented modular synthetic (VF) media is generally accepted as being ideally suited for highstrength wastewater (perhaps industrial) and high organic loadings such as with a roughing TF. In some cases, XF media has been placed in the top layer to enhance wastewater distribution and VF media comprises the remainder of the TF media. Typically, the top layer of a TF’s modular plastic media is covered with FRP or HDPE grating. The grating protects modular plastic media from deterioration by UV light and potential structural damage that may result from waterinduced load exerted during periods of high-intensity dosing. Figure 33 illustrates a typical TF column and a picture of the grating. Rock and random synthetic media are not self-supporting and require structural support to contain the media within the bioreactor. These containment structures are typically precast or panel-type concrete tanks. When self-supporting media such as plastic modules are used, other materials such as wood, fiberglass, and coated steel have been used as containment structures. The containment structure serves to avoid wastewater splashing, and to provide media support, wind protection, and flood containment. In some cases TF containment structures have been designed to allow flooding of the media, which increases operator flexibility in controlling macrofauna accumulation. The TF underdrain system is designed to meet two objectives: collect treated wastewater for conveyance to downstream unit processes and create a plenum that allows for the transfer of air throughout the TF media (Grady et al., 1999). Clay or concrete underdrain blocks are commonly used for rock-media TFs because of the required structural support. A variety of support systems including concrete piers and reinforced fiberglass plastic grating are used for other media types. The volume created between concrete and media bottom creates the underdrain.
TFs require oxygen to sustain aerobic biochemical transformation processes. The VF of air through the media can be induced mechanically or by natural draft. Natural air ventilation results from a difference in ambient air temperature outside and inside the TF. The temperature causes air to expand when warmed or contract when cooled. The net result is an air-density gradient throughout the TF, and an air front either rises or sinks depending on the differential condition. This rising or sinking action results in a continuous air flow through the bioreactor. Natural ventilation may become unreliable or inadequate in meeting process air requirements when neutral temperature gradients do not produce air movement. Currently, the provision of adequate underdrain and effluent channel sizing to permit free air flow is standard. Passive devices for ventilation include vent stacks on the TF periphery, extensions of underdrains through TFs side walls, ventilating manholes, louvers on the sidewall of the tower near the underdrain, and discharge of TF effluent to the subsequent settling basin in an open channel or partially filled pipes. Drains, channels, and pipes should be sufficiently sized to prevent submergence greater than 50% of their crosssectional area under design hydraulic loading. Ventilating access ports with open grating covers should be installed at both ends of the central collection channel. Large diameter TFs typically have branch channels (to collect the treated wastewater). These branches should also include ventilating manholes or vent stacks installed at the TF periphery. The open area of the slots in the top of the underdrain blocks should not be less than 15% of the TF area. One square meter gross area of open grating in ventilating manholes and vent stacks should be provided for each 23 m2 of TF area. Typically, 0.1 m2 of ventilating area is provided for every 3–4.6 m of TF periphery, and 1–2 m2 of ventilation area in the underdrain area per 1000 m3 of TF media. Another criterion for rockmedia TFs is the provision of a vent area at least equal to 15% of the TF cross-sectional area. Mechanical ventilation enhances and controls air flow with low-pressure fans that continuously circulate air throughout the TF. Therefore, a majority of new and improved TFs use low-pressure fans to mechanically promote air flow. The air flow resulting from natural draft will distribute itself. This will
Figure 33 Skid-resistant (polyethylene or fiberglass-reinforced plastic) grating placed on top of a typical modular plastic media trickling filter column.
Biofilms in Water and Wastewater Treatment
not occur with mechanical ventilation. Pressure loss through synthetic TF media is typically low, often less than 1-mm H2O/ m of TF depth (Grady et al., 1999). The low pressure drop typically results in low fan power requirements (B3–5 kW). The head on the fan is typically less than 1500-mm H2O. Unfortunately, the low pressure drop allows air to rise upward through the TF media without distributing itself across the bioreactor section. Therefore, fans are typically connected to distribution pipes. The air flow distribution piping has openings sized such that air flow through each is equal and air flow distribution is uniform. The pipes typically have a velocity in the range of 1100–2200 m h1 in order to further promote uniform air flow distribution. Air flow requirements are calculated based on process oxygen requirements and characteristic oxygen-transfer efficiency which is typically in the range of 2–10%. The mechanical air stream may flow upward to downward. Down-flow systems can be designed without covers. However, covers are required for systems that do not have air distribution through a network of pipes under the media. Covering TFs offers a wintertime benefit of limiting cold airflow and minimizing wastewater cooling. Mechanical ventilation and covered TFs may be used to destroy odorous compounds. A critical unit in the TF process is the pump station that lifts primary effluent (or screened raw sewage), and recirculates unsettled trickling effluent (here, referred to as underflow) to the influent stream. In general, TF underflow is recirculated to the distribution system to achieve the hydraulic load (influent þ recirculation) required for proper media wetting and biofilm thickness control, and decouple hydraulic and organic loading. TF influent is generally pumped to allow TF underflow to flow by gravity to the suspended growth reactor (or solids contact basin), secondary clarifier, or other downstream of the TF. When fit with weirs, a single pump station can be used to convey both influent and recirculation streams.
Table 7
563
TFs can be classified as roughing, carbon oxidation, carbon oxidation and nitrification, and nitrification. Table 7 summarizes characteristics of each TF. The performance ranges are associated with average design conditions. Single day or average week observations may significantly be greater.
4.15.4 Part III. Undesirable Biofilms: Examples of Biofilm-Related Problems in the Water and Wastewater Industries Biofilms are unavoidably associated with water environments, so biofilm control, a component of many industrial processes, is especially important in water and wastewater treatment. Depending on the particular setting, biofilms may cause process performance problems, material performance problems, health problems, and esthetic problems. The specific problems that biofilms cause in industrial settings are as diverse as the technological processes affected by the biofilms. In this section, we discuss four biofilm-related problems that have been reported in the water and wastewater industries: 1. biofilms on metal surfaces and MIC; 2. biofilms on concrete surfaces and crown corrosion of sewers; 3. biofilms on filtration membranes in drinking water treatment; and 4. biofilms on filtration membranes in wastewater treatment.
4.15.4.1 Biofilms on Metal Surfaces and MIC In the manufacturing of metals and metal alloys, raw materials – the ores – are chemically reduced and their internal chemical energy increases. These materials are used by microorganisms as sources of energy in a sequence of processes in which the chemical energy of the affected material decreases, bringing the energy levels of the products closer to
Trickling filter classification
Design parameter
Roughinga
Carbon oxidizing (cBOD5 removal)a
Carbon oxidation and nitrificationa
Nitrificationa
Media typically used
VF
RO, XF, or VF
RO, XF, or VF
XF
Wastewater source
Primary effluent
Primary effluent
Primary effluent
Secondary effluent
Hydraulic loading m3 d1 m2 BOD5 and NH3 N Load kg m3 d1 g m2 d1
52.8–178.2
14.7–88.0
14.7–88.0
35.2–88.0
1.6–3.52 NA
0.32–0.96 NA
0.08–0.24 0.2–1.0
NA 0.5–2.4
Conversion (%) or effluent concentration (mg l1) Macro fauna
50–75% filtered cBOD5 conversion No appreciable growth
20–30 mg l1 cBOD5 and TSSb Beneficial
0.5–3 mg l1 as NH3 Nb
Depth, m (ft)
0.91–6.10
r12.2
o10 mg l1 as cBOD5; o3 mg l1 as NH3 Nb Detrimental (nitrifying biofilm) r12.2
a
Detrimental r12.2
Applicable to shallow trickling filters. gpm ft2, gallons per minute per square foot of trickling filter plan area. Concentration remaining in the clarifier effluent stream. From Boltz JP, Morgenroth E, deBarbadillo C, et al. (2010b) Biofilm reactor technology and design. In: Design of Municipal Wastewater Treatment Plants, WEF Manual of Practice No. 8, ASCE Manuals and Reports on Engineering Practice No. 76, 5th edn, vol. 2, ch. 13, p. 238 (ISBN P/N 978-0-07-166360-1 of set 978-0-07-166358-8; MHID P/N 0-07166360-6 of set 0-07-166358-4). New York: McGraw-Hill. b
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Biofilms in Water and Wastewater Treatment
the energy levels of the materials from which they were made. MIC can affect a variety of materials, both metallic and nonmetallic. If nonmetallic materials are affected, the term biodeterioration of materials is more often used than MIC, although this terminology is not very consistent and, for example, the term crown corrosion of sewers, which in fact refers to the biodeterioration of concrete, is quite popular among water professionals. Accelerated corrosion of metals in the presence of microorganisms stems from microbial modifications to the chemical environment near metal surfaces (Beech et al., 2005; Geiser et al., 2002; Lee and Newman, 2003; Lewandowski et al., 1997). Such modifications depend, of course, on the properties of the corroding metal and on the microbial community structure of the biofilm deposited on the metal surface (Beech and Sunner, 2004; Dickinson et al., 1996; Flemming, 1995; Olesen et al., 2000, 2001). Beech et al. (2005) described MIC as a consequence of coupled biological and abiotic electron transfer reactions, that is, redox reactions of metals enabled by microbial ecology (Beech et al., 2005). Hamilton (2003) attempted to generate a unified concept of MIC but found common features in only some of the possible mechanisms (Hamilton, 2003). It is unlikely that a unified concept of MIC can be generated at all. Rather, there are many mechanisms by which microorganisms may affect metal surfaces, and we demonstrate some of them here. These do not exhaust the possibilities, of course, but are rather used to exemplify the possible mechanisms. As we have restricted the discussion of MIC to metal surfaces only, it is convenient to define corrosion as anodic dissolution of a metal. In this way we can easily separate the corrosion reaction, the anodic dissolution of the metal, from many other anodic reactions that can occur at a metal surface covered with a biofilm. These other anodic reactions deliver electrons originating from substances metabolized near the metal surface, but only the reaction in which the metal itself is oxidized is defined as corrosion. The presence of the other anodic reactions causes confusion in MIC studies, as the current between the anode and the cathode is made up of electrons originating from many anodic reactions occurring at the surface, not only from the corrosion reaction. Microorganisms generate chemical environments that are conducive to corrosion reactions even if they do not take part in the process themselves. As in most industrial processes microorganisms are always present on metal surfaces, it is not immediately obvious whether the microorganisms attached to the surface accelerate the corrosion process or are just innocent bystanders. The only way to resolve this is by demonstrating that a specific mechanism of MIC is present because a product of microbial metabolism consistent with this mechanism can be detected. Many mechanisms of MIC have been proposed. Accelerated corrosion may result from the action of acid-producing bacteria, such as Thiobacillus thiooxidans and Clostridium aceticum; iron-oxidizing bacteria, such as Gallionella, Sphaerotilus, and Leptothrix; MOB, such as L. discophora; or hydrogenproducing bacteria. These mechanisms have been studied and the results described in numerous publications. We describe here representative examples of such mechanisms: the effects of differential aeration cells, sulfate-reducing bacteria (SRB corrosion), and MOB corrosion.
4.15.4.1.1 Differential aeration cells on iron surfaces MIC caused by differential aeration cells is an example of a nonspecific mechanism of MIC, because it depends on the presence of biofilm, and not on the type of microorganisms that reside in the biofilm. If the oxygen concentrations at two adjacent locations on an iron surface are different, then the half-cell potentials at these locations are different as well. The location where the oxygen concentration is higher will have a higher potential (more cathodic) than the location where the oxygen concentration is lower (more anodic). The difference in potential will give rise to a current flow from the anodic locations to the cathodic locations and to the establishment of a corrosion cell. This is the mechanism of differential aeration cells, and the prerequisite to this mechanism is that the concentration of oxygen varies among locations (Acuna et al., 2006; Dickinson and Lewandowski, 1996; Hossain and Das, 2005). Indeed, many measurements using oxygen microsensors have demonstrated that oxygen concentrations in biofilms can vary from one location to another (Lewandowski and Beyenal, 2007). If the anodic reaction is the oxidation of iron,
Fe-Fe 2þ þ 2e
ð17Þ
and the cathodic reaction is the reduction of oxygen,
O2 þ 2H2 O þ 4e -4OH
ð18Þ
then the overall reaction describing the process is
2Fe þ O2 þ 2H2 O-2Fe2þ þ 4OH
ð19Þ
The Nernst equation quantifying the potential for this reaction is
E ¼ Eo
0:059 ½Fe 2þ 2 ½OH 4 log 4 pðO2 Þ
ð20Þ
Figure 34 visualizes this mechanism.
4.15.4.1.2 SRB corrosion SRB causes corrosion of cast iron, carbon, and low alloy steels and stainless steels. SRB corrosion of potable water mains is a common (US EPA, 1984) and well-recognized problem (Seth and Edyvean, 2006; Tuovinen et al., 1980). MIC caused by SRB is an example of a mechanism that depends on the activity of a specific group of microorganisms in a biofilm. The corrosion of mild steel caused by SRB is the most notorious case of MIC, and it provides a direct and easy-to-understand link between microbial reactions and electrochemistry (Javaherdashti, 1999). According to the mechanism that was originally proposed by Von Wohlzogen Kuhr in 1934, SRB oxidizes cathodically generated hydrogen to reduce sulfate ions to H2S, thereby removing the product of the cathodic reaction and stimulating the progress of the reaction (Al Darbi et al., 2005). This mechanism was later found to be inadequate to explain the field observations. More involved mechanisms were implicated in this type of microbial corrosion, including the puzzling effect of oxygen, which can stimulate what is apparently an anaerobic process. It is now certain that the
Biofilms in Water and Wastewater Treatment
565
Aerated water Cathodic site; corrosion products
Biofilm
OH−
OH−
Cathode − e
Biof ilm
O2
e−
Anodic site
Biofilm
O2
O O2 Aerobic 2 O2 O2 O2 Anaerobic O2 O O2 2 O2 O2 Anaerobic O 2 O2 O2 O2 M+ M+ M+ Anode
O2 O2 OH−OH− e−
Cathode e−
1 mm
Metal (b)
(a)
Figure 34 Biofilm heterogeneity results in differential aeration cells. (a) This schematic shows pit initiation due to oxygen depletion under a biofilm (Borenstein, 1994). (b) An anodic site and pit under the biofilm and corrosion products deposited on mild steel.
possible pathways for cathodic reactions include sulfides and bisulfides as cathodic reactants (Videla, 2001; Videla and Herrera, 2005). The currently accepted mechanism of SRB corrosion is composed of a network of reactions that reflects the complexity of the environment near corroding metal surfaces covered with biofilms; the following paragraphs illustrate some of this complexity. The process starts with the microbial metabolism of SRB producing hydrogen sulfide by reducing sulfate ions. Hydrogen sulfide can serve as a cathodic reactant, thus affecting the rate of corrosion (Antony et al., 2007; Costello, 1974):
2H2 S þ 2e -H2 þ 2HS
ð21Þ
Ferrous iron generated from anodic corrosion sites precipitates with the metabolic product of microbial metabolism, hydrogen sulfide, forming iron sulfides, FeSx:
Fe 2þ þ HS ¼ FeS þ H þ
ð22Þ
This reaction may provide protons for the cathodic reaction (Crolet, 1992). The precipitated iron sulfides form a galvanic couple with the base metal. For corrosion to occur, the iron sulfides must have electrical contact with the bare steel surface. Once contact is established, the mild steel behaves as an anode and electrons are conducted from the metal through the iron sulfide to the interface between the sulfide deposits and water, where they are used in a cathodic reaction. Surprisingly, the most notorious cases of SRB corrosion often occur in the presence of oxygen. As SRB is anaerobic microorganisms, this fact has been difficult to explain. This effect of oxygen can be explained based on a mechanism in which iron sulfides (resulting from the reaction between iron ions and sulfide and bisulfide ions) are oxidized by oxygen to elemental sulfur, which is known to be a strong corrosion agent (Lee et al., 1995). Biofilm heterogeneity plays an important role in this process, because the central parts of microcolonies are anaerobic while the outside edges remain aerobic
(Lewandowski and Beyenal, 2007). This arrangement makes this mechanism of microbial corrosion possible, because the oxidation of iron sulfides produces highly corrosive elemental sulfur, as illustrated by the following reaction:
2H2 O þ 4FeS þ 3O2 -4So þ 4FeOðOHÞ
ð23Þ
Hydrogen sulfide can also react with the oxidized iron to form ferrous sulfide and elemental sulfur (Schmitt, 1991), thereby aggravating the situation by producing even more elemental sulfur, and closing the loop through production of the reactant used in the first reaction, FeS:
3H2 S þ 2FeOðOHÞ-2FeS þ So þ 4H2 O
ð24Þ
The product of these reactions – elemental sulfur – increases the corrosion rate. Schmitt (1991) has shown that the corrosion rate caused by elemental sulfur can reach several hundred mpy (Schmitt, 1991). We have demonstrated experimentally that elemental sulfur is deposited in the biofilm during SRB corrosion (Nielsen et al., 1993), thereby detecting the component vital for this mechanism to occur. It is also well known that the sulfur disproportionation reaction that produces sulfuric acid and hydrogen sulfide is carried out by sulfur-disproportionating microorganisms (Finster et al., 1998). Also, several microbial species, such as T. thiooxidans, can oxidize elemental sulfur and sulfur compounds and produce sulfuric acid:
4S o þ 4H2 O-3H2 S þ H2 SO4
ð25Þ
In summary, the SRB corrosion of mild steel in the presence of oxygen is an acid corrosion: Anodic reaction:
Fe-Fe 2þ þ 2e
ð26Þ
2H þ þ 2e-H2
ð27Þ
Cathodic reaction:
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Biofilms in Water and Wastewater Treatment 4.15.4.3 Biofilms on Filtration Membranes in Drinking Water Treatment
O2
2−
FeO(OH)
SO4
SO42−
SO42− H2S S
0
FeS2 FeS
H+
O2
Fe2+ Metal
S0
HS− H2
O2
e
Figure 35 The SRB corrosion of mild steel in the presence of oxygen is an acid corrosion (Lewandowski et al., 1997).
The mechanism of SRB corrosion involves several loops, cycles in which reactants are consumed in one reaction and recycled in other reaction; the process is spontaneous at the expense of the energy released by the oxidation of the metal. This mechanism also demonstrates how the reactants and products of corrosion processes are included in the metabolic reactions of the microorganisms. For example, hydrogen, the product of the cathodic reaction above, is oxidized by some species of SRB to reduce sulfate and generate hydrogen sulfide, H2S (Cord-Ruwisch and Widdel, 1986), which is the reactant in the first reaction we referred to in this section. Hydrogen sulfide then dissociates to bisulfides:
H2 S ¼ Hþ þ HS
ð28Þ
which are then used in the reactions described above. Figure 35 shows the network of reactions described above.
4.15.4.2 Biofilms on Concrete Surfaces: Crown Corrosion of Sewers The mechanism of crown corrosion of sewers is very similar to the mechanism of MIC corrosion of metals caused by SRB. In sewers, SRB reduces sulfate ions to sulfides, which are oxidized by oxygen to elemental sulfur. Then the elemental sulfur is further oxidized, mainly by T. thiooxidans, but also by other Thiobacillus species, such as T. novellus/intermedius and T. neapolitanus, in a complex ecosystem on the sewer pipe (Vincke et al., 2001). As a result, sulfuric acid is produced, which dissolves the concrete and damages the sewers (Padival et al., 1995; Islander et al., 1991; Sand and Bock, 1984). The following reactions illustrate this action:
H2 SO4 þ CaCO3 -CaSO4 þ H2 CO3
ð29Þ
H2 SO4 þ CaðOHÞ2 -CaSO4 þ 2H2 O
ð30Þ
Crown corrosion of sewers depends on the presence of biofilm on the concrete surface and on the generation of sulfuric acid in immediate proximity to the concrete surface.
The common use of membranes in various technologies of water and wastewater treatment is probably the most visible mark of the changes that occurred in these applications in the last decade, and it is expected that filtration membranes will be even more popular in the future than they are now (Shannon et al., 2008). The traditional use of membranes in water treatment has been in the desalination of sea and brackish waters using the reverse osmosis (RO) process, and there is a large body of knowledge accumulated on this application. RO membrane filtration is becoming even more popular as the cost of desalination decreases because of various improvements in the technology that reduce the energy consumption and because of the use of new materials that produce less expensive and more robust membranes (Veerapaneni et al., 2007). Membrane processes have been introduced into other types of water treatment, besides desalination, such as water softening (Conlon et al., 1990). The main advantages of using membrane filtration in water treatment are that the process does not require using chemicals and that the membrane modules have a smaller footprint than the conventional treatment facilities. Membrane filtration can be used instead of other traditional processes in water treatment, such as coagulation, sand and activated carbon filtration, or ion exchange, without the necessity of adding chemicals to the water, which helps prevent the formation of disinfection byproducts, for example. Membrane filtration can be used alone in water treatment or in combination with other processes, in hybrid arrangements. For example, it can be used in combination with powdered activated carbon (PAC) to remove disinfection byproducts that exist in the raw water (Khan et al., 2009). Excessive biofouling of membranes is a problem in all membrane applications, but RO and nanofiltration (NF) processes are the most sensitive to biofouling (Vrouwenveldera et al., 2009). Much research has been done toward understanding the process of biofilm formation on these membranes and developing methods for cleaning the membranes. The removal of biofilm from RO membranes can be accomplished by mechanical or by chemical methods, or by a combination of mechanical and chemical methods. Mechanical methods include flushing with water or with water and air. Mechanical cleaning can be used alone or it can be followed by chemical cleaning. The simplest method of mechanical cleaning is the forward flush, in which the water flow rate above the membrane is increased to increase the shearing force and remove the deposits from the membrane. To increase the shearing force even further, air can be introduced into the conduit delivering the cleaning water. The air bubbles introduce additional instability into the flow field and increase the shearing force exerted on the surface. The backward flush is based on reversing the direction of filtration: cleaning water is filtered in the opposite direction and the particles trapped in membrane pores are removed. Depending on the contaminants deposited on the membranes, the surface can be cleaned chemically using various type of chemicals. If the deposits are predominantly inorganic scale, then the chemical cleaning can include agents that act mostly on scale, such as hydrochloric acid (HCl) or nitric acid (HNO3). If the
Biofilms in Water and Wastewater Treatment
biofilm is the main problem, then the cleaning substance may include antimicrobial agents to remove the biofilms. Two types of antimicrobial agents are in common use for this purpose: oxidizing and nonoxidizing biocides. The oxidizing biocides popular in membrane cleaning processes include chlorine, bromine, chloramine, chlorine dioxide, hydrogen peroxide, peroxyacetic acid, and ozone. Nonoxidizing biocides include formaldehyde, glutaraldehyde, and quaternary ammonium compounds. One recent study targeted cell–cell communications in biofilms to develop a novel approach in controlling membrane fouling (Yeon et al., 2009). Much effort has been directed toward the development of membranes with new or modified materials that can resist biofouling and toward modifying the surfaces of ultrafiltration (UF) and NF membranes by the graft polymerization of hydrophilic monomers that resist biofouling or allow more aggressive chemical treatment of the membranes (Hester et al., 2002; Wang et al., 2005; Asatekin et al., 2006, 2007). According to recent studies, in spiral-wound membrane modules, biofilm accumulation has a major impact on the spacer channel but the actual fouling of the membrane contributes to the overall pressure drop to a much smaller extent than previously assumed (Vrouwenveldera et al., 2009).
4.15.4.4 Biofilms on Filtration Membranes in Wastewater Treatment Membrane filtration is used in two types of wastewater technologies: (1) membrane bioreactors (MBRs) and (2) membrane biofilm reactors (MBfRs). This terminology is somewhat confusing: the names sound similar, and the fact that the obvious acronyms for the two technologies are the same does not help. It is therefore customary to call the MBRs and the MBfRs. From the biofouling point of view, microbial growth on membranes is undesirable (Le-Clech et al., 2006) while in MBfRs biofilm growth on the membrane is necessary for process performance. MBfRs are used to deliver dissolved gases, such as oxygen, hydrogen, and methane, to the microorganisms attached to the membrane (Brindle and Stephenson, 1996; Brindle et al., 1998; Suzuki et al., 2000; Lee and Rittmann, 2000; Pankhania et al., 1999; Modin et al., 2008). MBRs are used to replace gravity settling in the secondary sedimentation tanks used in traditional biological wastewater treatment; for example, the activated sludge process where membrane processes can be used to separate the biomass of suspended microorganisms from the effluent. The membranes used in MBRs are typically UF membranes. MBR technology is well established in wastewater treatment: it has been implemented on large scales (Melin et al., 2006), and textbooks have been published describing its application (Stephenson et al., 2000; Judd, 2006). Using membrane filtration to replace gravity settling has many advantages, and one of them is avoidance of the notorious problems with sludge bulking that plague many activated sludge treatment plants. Membranes in MBRs suffer from biofouling, which decreases the permeate flow (Howell et al., 2003; Young et al., 2006; Kimura et al., 2005) Large-scale operations suffer from this problem, particularly the irreversible fouling that cleaning does not remove (Wang et al., 2005). The most common solution to the excessive accumulation of biomass is bubbling air near the
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membrane’s surface, which creates high shear and removes the biomass (Hong et al., 2002). Basic studies on biofilm formation (Davies et al., 1998) indicate that bacteria regulate their group behaviors, such as biofilm formation, in response to population density using small signal molecules called autoinducers, or quorumsensing molecules. It is expected that interference with microbial communication systems in biofilms may lead to novel approaches to preventing biofouling in many areas. Three strategies for interfering with autoinducer molecules have been proposed: blockage of autoinducer production, interference with signal receptors, and inactivation of autoinducer molecules (Rassmusen and Givskov, 2006). In a recent study, Yeon et al. (2009) demonstrated that inactivating the autoinducer molecules in a batch-type MBR reactor decreased the amount of EPS deposited on the membrane and that interfering with cell–cell communication in biofilms can alleviate the fouling of filtration membranes.
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4.16 Membrane Biological Reactors FI Hai, University of Wollongong, Wollongong, NSW, Australia K Yamamoto, University of Tokyo, Tokyo, Japan & 2011 Elsevier B.V. All rights reserved.
4.16.1 4.16.2 4.16.2.1 4.16.2.2 4.16.3 4.16.3.1 4.16.3.2 4.16.3.2.1 4.16.3.2.2 4.16.3.2.3 4.16.3.2.4 4.16.3.2.5 4.16.3.2.6 4.16.3.3 4.16.3.4 4.16.3.4.1 4.16.3.4.2 4.16.3.4.3 4.16.3.4.4 4.16.3.4.5 4.16.4 4.16.4.1 4.16.4.2 4.16.4.3 4.16.4.4 4.16.4.4.1 4.16.4.4.2 4.16.4.4.3 4.16.4.4.4 4.16.4.4.5 4.16.4.4.6 4.16.4.4.7 4.16.4.5 4.16.4.6 4.16.4.7 4.16.4.7.1 4.16.4.7.2 4.16.4.7.3 4.16.4.7.4 4.16.5 4.16.5.1 4.16.5.1.1 4.16.5.1.2 4.16.5.1.3 4.16.5.1.4 4.16.5.2 4.16.5.3 4.16.5.4 4.16.5.4.1 4.16.5.4.2 4.16.5.4.3 4.16.5.5 4.16.5.5.1
Introduction Aeration and Extractive Membrane Biological Reactors Aeration Membrane Biological Reactor Extractive Membrane Biological Reactor History and Fundamentals of Biosolid Separation MBR Historical Development Process Comparison with Conventional Activated Sludge Process Treatment efficiency/removal capacity Sludge properties and composition Sludge production and treatment Space requirements Wastewater treatment cost Comparative energy usage Relative Advantages of MBR Factors Influencing Performance/Design Considerations Pretreatment Membrane selection and applied flux Sludge retention time Mixed liquor suspended solids concentration Oxygen transfer Worldwide Research and Development Challenges Importance of Water Reuse and the Role of MBR Worldwide Research Trend Modeling Studies on MBR Innovative Modifications to MBR Design Inclined plate MBR Integrated anoxic–aerobic MBR Jet-loop-type MBR Biofilm MBR Nanofiltration MBR Forward osmosis MBR Membrane distillation bioreactor Technology Benefits: Operators’ Perspective Technology Bottlenecks Membrane Fouling – the Achilles’ Heel of MBR Technology Fouling development Types of membrane fouling Parameters influencing MBR fouling Fouling mitigation Worldwide Commercial Application Installations Worldwide Location-specific drivers for MBR applications Plant size Development trend and the current status in different regions Decentralized MBR plants: Where and why? Commercialized MBR Formats Case-Specific Suitability of Different Formats MBR Providers Market share of the providers Design considerations Performance comparison of different providers Standardization of Design and Performance-Evaluation Method Standardization of MBR filtration systems
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Standardization of MBR characterization methods Future Vision Conclusion
4.16.1 Introduction
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Membrane biological reactors combine the use of biological processes and membrane technology to treat wastewater. The use of biological treatment can be traced back to the late nineteenth century. It became a standard method of wastewater treatment by the 1930s (Rittmann, 1987). Both aerobic and anaerobic biological treatment methods have been extensively used to treat domestic and industrial wastewater (Visvanathan et al., 2000). After removal of the soluble biodegradable matter in the biological process, any biomass formed needs to be separated from the liquid stream to produce the required effluent quality. In the conventional process, a secondary settling tank is used for such solid/liquid Apprx. molecular weight 200 µm 0.001
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separation and this clarification is often the limiting factor in effluent quality (Benefield and Randall, 1980). Membrane filtration, on the other hand, denotes the separation process in which a membrane acts as a barrier between two phases. In water treatment, the membrane consists of a finely porous medium facilitating the transport of water and solutes through it (Ho and Sirkar, 1992). The separation spectrum for membranes, illustrated in Figure 1, ranges from reverse osmosis (RO) and nanofiltration (NF) for the removal of solutes, to ultrafiltration (UF) and microfiltration (MF) for the removal of fine particulates. MF and UF membranes are predominantly used in conjunction with biological reactors (Pearce, 2007). UF can remove the finest particles found in water supply, with the removal rating dependent upon the
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Figure 2 Market drivers for membranes in wastewater. Information from Howell JA (2004) Future of membranes and membrane reactors in green technologies and for water reuse. Desalination 162: 1–11; and Pearce G (2007) Introduction to membranes: Filtration for water and wastewater treatment. Filtration and Separation 44(2): 24–27.
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et al., 2000); for bubble-less aeration of the bioreactor (Brindle and Stephenson, 1996); and for extraction of priority organic pollutants from hostile industrial wastewaters (Stephenson et al., 2000). There are other forms of membrane biological reactors such as enzymatic membrane bioreactor (Charcosset, 2006) for production of drugs, vitamins, etc., or membrane biological reactors for waste-gas treatment (Reij et al., 1998), a discussion about which is beyond the scope of this chapter. Solid–liquid membrane-separation bioreactors employ UF or MF modules for the retention of biomass to be recycled into the bioreactor. Gas-permeable membranes are used to provide bubble-less oxygen mass transfer to degradative bacteria
Oxygen transfer
pore size of the active layer of the membrane. The complete pore-size range for UF is approximately 0.001–0.02 mm, with a typical removal capability of UF for water and wastewater treatment of 0.01–0.02 mm. MF typically operates at a particle size that is up to an order of magnitude coarser than this. In water treatment, the modern trend is to use a relatively tight MF with a pore size of approximately 0.04–0.1 mm, whereas wastewater normally uses a slightly more open MF with a pore size of 0.1–0.4 mm (though wastewater can be treated using UF membranes, or MF membranes used for water applications). The market drivers for membranes in wastewater are illustrated in Figure 2. However, as in any separation process, in membrane technology too, the management and disposal of concentrate is a significant issue. Environment-friendly management and disposal of the resulting concentrates at an affordable cost is a significant challenge to water and wastewater utilities and industry. To eradicate the respective disadvantages of the individual technologies, the biological process can be integrated with membrane technology. Although some recent studies have demonstrated case-specific feasibility of direct UF of raw sewage (Janssen et al., 2008), membranes by themselves are seldom used to filter untreated wastewater, since fouling prevents the establishment of steady-state conditions and because water recovery is very low (Schrader et al., 2005; Fuchs et al., 2005; Judd and Jefferson, 2003). However, membrane filtration can be efficiently used in combination with a biological process. The biological process converts dissolved organic matter into suspended biomass, reducing membrane fouling and allowing increase in recovery. On the other hand, in the membrane filtration process, the membranes introduced into the bioreactors not only replace the settling unit for solid–liquid separation but also form an absolute barrier to solids and bacteria and retain them in the process tank. As our understanding of membrane technology grows, we learn that membrane technology is now being applied to a wider range of industrial applications and is used in many new forms for wastewater treatment. Combining membrane technology with biological reactors for the treatment of municipal and industrial wastewaters has led to the development of three generic membrane processes within bioreactors (Figure 3): for separation and recycle of solids (Visvanathan
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Biofilm Nutrient biomedium (c) Figure 3 Simplified representation of membrane biological reactors: (a) biosolid separation, (b) aeration, and (c) extractive membrane biological reactors.
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present in the bioreactor. Additionally, the membrane can act as support for biofilm development, with direct oxygen transfer through the membrane wall in one direction and nutrient diffusion from the bulk liquid phase into the biofilm in the other direction. An extractive membrane process has been devised for the transfer of degradable organic pollutants from hostile industrial wastewaters, via a nonporous silicone membrane, to a nutrient medium for subsequent biodegradation. Biosolid separation is, however, the most widely studied process and has found full-scale applications in many countries. In a comprehensive review published in 2006, Yang et al. (2006) pointed out that the vast majority of research on membrane biological reactors since 1990 focused on biosolidseparation-type applications. There was no significant increase in the number of studies on gas diffusion and extractive membrane biological reactors over time. Publications on extractive and diffusive membrane biological reactors became available during 1994–95, after which a steady output of less than five publications a year was observed. This indicates that current research is predominantly in the water and wastewater-filtration area, in parallel with the commercial success in this field. In line with the current trend of research and commercial application, this chapter focuses on the biosolidseparation membrane biological reactors, which is more commonly known as membrane bioreactor (MBR). However, a brief outline of the other two types of membrane biological reactors is furnished in Section 4.16.2. The remainder of this chapter elaborates on the history, fundamentals, research and development challenges, as well as the commercial application of the biosolid-separation membrane biological reactors, which are henceforth referred to as MBRs.
4.16.2 Aeration and Extractive Membrane Biological Reactors 4.16.2.1 Aeration Membrane Biological Reactor Wastewater-treatment processes using high-purity oxygen have a greater volumetric degradation capacity compared to the conventional air-aeration process. However, conventional oxygenation devices have high power requirements associated with the need for high mixing rate, and cannot be used in conjunction with biofilm processes. In the membraneaeration biological reactors (MABRs), the capability of biofilm to retain high concentrations of active bacteria is coupled with the high oxygen transfer rate to the biofilm. The key characteristic advantages of MABRs are summarized as follows:
• •
High oxygen transfer rate, especially suitable for highoxygen-demanding wastewaters. In conventional aerobic biological wastewater treatment, volatile organic compounds (VOCs) can escape to the atmosphere without being biodegraded as a result of air bubbles stripping out the compounds from the bulk liquid. Since no oxygen bubbles are formed in MABRs, gas stripping of VOCs and foaming due to the presence of surfactants can be prevented (Rothemund et al., 1994; Semmens 1991; Wilderer et al., 1985) to a greater extent.
•
Membrane-attached biofilms are in intimate contact with the oxygen source, with direct interfacial transfer and utilization of oxygen within the biofilm. In contrast to conventional biofilm processes, in MABR biofilms, oxygen from the membrane and pollutant substrate(s) from the bulk liquid are transferred across the biofilm in countercurrent directions (Figure 4). Biofilm stratification in MABRs results from this distribution of the maximum oxygen and pollutant-substrate concentrations at different locations within the biofilm and the relative thickness of MABR biofilms; this enables the removal of more than one pollutant type. The high oxygen concentrations coupled with the low organic carbon concentrations near the membrane/biofilm interface encourage nitrification, an aerobic heterotrophic layer above this facilitates organic carbon oxidation, and an anoxic layer near the biofilm/ liquid interface supports denitrification (Stephenson et al., 2000).
MABRs have been used to treat a variety of wastewater types at the laboratory scale (Brindle and Stephenson, 1996). However, in line with the characteristics of MABRs discussed above, most investigations show that the process is particularly suitable for the treatment of high-oxygen-demanding wastewaters, biodegradation of VOCs, combined nitrification, denitrification, and/or organic carbon oxidation in a single biofilm. Bubble-less oxygen mass transfer can be accomplished using gas-permeable dense membranes or hydrophobic microporous membranes (Cote et al., 1988). Both plate and frame and hollow-fiber membrane configurations have been used to supply the oxygen. Oxygen diffusion through dense membrane material can be achieved at high gas pressures without bubble formation. In hydrophobic microporous membranes, the pores remain gas filled; and oxygen is transported to the shell side of the membrane through the pores by gaseous diffusion or Knudsen flow-transport mechanisms. The partial pressure of the gas is kept below the bubble point to ensure the bubble-less supply of oxygen (Ahmed and Semmens, 1992; Rothemund et al., 1994; Semmens, 1991; Semmens and Gantzer, 1993). Pressurized hollow fibers have been investigated in the dead-end and flow-through modes of operation. The evacuation of carbon dioxide from the bioreactor is a benefit of flow-through operation, though no quantitative work to determine removal rates has been undertaken (Cote et al., 1997; Kniebusch et al., 1990). Deadend operation has usually been avoided, due to significantly decreased performance and condensate formation in the lumen (Cote et al., 1997). The nonbiological fouling and loss of performance of dead-end porous hollow fibers due to iron oxidation, absorption of free oils and greases into pores, surfactants, and suspended solids, and fiber tangling have been reported (Semmens and Gantzer, 1993). Chemical treatment of the dead ends of these hollow fibers may provide a means for the condensate to escape. The liquid boundary layer normally has a greater impact upon the overall oxygen mass transfer than the membrane, with mixing of the liquid a key operational parameter (Cote et al., 1997; Kniebusch et al., 1990; Wilderer et al., 1985). However, wall thickness significantly affects the transport of
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Biofilm Figure 4 Simplified representation of the steady-state concentration profiles of oxygen, carbon substrate, and microbial activity in case of MABR biofilm and conventional biofilm.
oxygen through dense polymer membranes (Wilderer et al., 1985). Oxygen transport is also controlled by the presence of membrane-attached biofilm and its thickness; the partial pressure of oxygen and flow-velocity characteristics on the lumen side; and the wastewater flow-velocity characteristics on the shell side of the membrane (Kniebusch et al., 1990; Pankania et al., 1994). Oxygen partial pressure provides the means for controlling the depth of oxygen penetration into the wastewater, with an increase in partial pressure resulting in an increase in the metabolic activity of the membraneattached biofilm (Rothemund et al., 1994). In bioreactors, most membranes used for oxygen mass transfer operate with the biofilm attached to the membrane surface. These biofilms are in intimate contact with the oxygen source and are protected against abrasion and grazing (Kniebusch et al., 1990; Rothemund et al., 1994). Scanning electron micrographs show that some attached bacteria inhabit the membrane pores, with the location of the oxygen and wastewater interphase very close to the bacteria (Rothemund et al., 1994). Thus, oxygen-transfer resistance due to the thickness of the porous membrane and the liquid boundary layer are not necessarily decisive limiting factors (Kniebusch et al., 1990; Rothemund et al., 1994; Wilderer et al., 1985).
Excessive biofilm accumulation can result in the transport limitation of oxygen and nutrients, plugging of membrane fibers, a decline in biomass activity, metabolite accumulation deep within the biofilm, and the channeling of flow in the bioreactor such that steady-state conditions may not be maintained (Debus and Wanner, 1992; Pankania et al., 1994; Yeh and Jenkins, 1978). To operate at maximum efficiency, occasional membrane washing, air scouring, backwashes, and high recirculation rate of wastewater to achieve high shear velocities have all been employed to control biomass accumulation. In the MABR process, oxygen is transferred without forming bubbles and therefore cannot be utilized to mix the bulk liquid. In laboratory scale MABRs, liquid-phase mixing has been achieved using recirculation pumps, impellers, magnetic stirrers, nitrogen, or air sparging.
4.16.2.2 Extractive Membrane Biological Reactor The extractive membrane biological reactor (EMBR) process enables the transfer of degradable organic pollutants from hostile industrial wastewaters, via a dense silicone membrane,
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to a nutrient medium for subsequent degradation (Brindle and Stephenson, 1996). Membranes used for the extraction of pollutants into a bioreactor have been developed using pervaporation by exchanging the vacuum phase with a nutrient biomedium phase where biodegradation mechanisms maintain the concentration gradient needed to transfer organic pollutants present in hostile industrial wastewaters (Lipski and Cote, 1990; Nguyen and Nobe, 1987; Yun et al., 1992). The inorganic composition of the nutrient medium is unaffected by the industrial wastewater within the hydrophobic hollow-fiber membrane. Hence, the conditions within the bioreactor can be optimized to ensure high biodegradation rate (Brookes and Livingston, 1993; Livingston, 1993, 1994). The extraction and biodegradation of toxic volatile organic pollutants, such as chloroethanes, chlorobenzenes, chloroanilines, and toluene from hostile industrial wastewaters, with high salinity and extremes of pH, using EMBRs have been demonstrated at the laboratory scale (Stephenson et al., 2000). Further information on these two generic types of MBRs can be derived from the review papers by Brindle and Stephenson (1996) and McAdam and Judd (2006), and the book by Stephenson et al. (2000). Yang et al. (2006) argued that extractive or aeration MBRs present a significant opportunity for researchers as niche areas of application as these novel processes remain unexplored. Hazardous waste treatment and toxic waste cleanup present two potential areas for the EMBR (Brookes and Livingston, 1994; Dossantos and Livingston, 1995; Livingston et al., 1998), whereas hydrogenotrophic denitrification of groundwater (Clapp et al., 1999; Mo et al., 2005; Modin et al., 2008; Nuhoglu et al., 2002; Rezania et al., 2005) and gas-extractionassisted fermentation (Daubert et al., 2003; Lu et al., 1999) are potential research areas for the AMBR. It is also important to recognize the fact that these three membrane processes are not mutually exclusive and, if necessary, could be coupled into one bioreactor (Brindle and Stephenson, 1996). Once the research field has gained momentum, commercial interest might correspondingly follow.
4.16.3 History and Fundamentals of Biosolid Separation MBR 4.16.3.1 Historical Development Membranes have been finding wide application in water and wastewater treatment ever since the early 1960s when Loeb and Sourirajan invented an asymmetric cellulose acetate membrane for RO (Visvanathan et al., 2000). Many combinations of membrane solid/liquid separators in biological treatment processes have been studied since. The first descriptions of the MBR technology date from the late 1960s. The trends that led to the development of today’s MBR are depicted in Figure 5. When the need for wastewater reuse first arose, the conventional approach was to use advanced treatment processes. The progress of membrane manufacturing technology and its applications could lead to the eventual replacement of tertiary treatment steps by MF or UF (Figure 5(a)). Parallel to this development, MF or UF was used for solid/liquid separation in the biological treatment
process and thereby sedimentation step could be eliminated (Figure 5(b)). The original process was introduced by DorrOlivier Inc. and combined the use of an activated sludge bioreactor with a cross-flow membrane-filtration loop (Smith et al., 1969). By pumping the mixed liquor at a high pressure into the membrane unit, the permeate passes through the membrane and the concentrate is returned to the bioreactor (Hardt et al., 1970; Arika et al., 1966; Krauth and Staab, 1988; Muller et al., 1995). The flat-sheet membranes used in this process were polymeric and featured pore size ranging from 0.003 to 0.01 mm (Enegess et al., 2003). Although the idea of replacing the settling tank of the conventional activated sludge (CAS) process was attractive, it was difficult to justify the use of such a process because of the high cost of membranes, low economic value of the product (tertiary effluent), and the potential rapid loss of performance due to fouling. Due to the poor economics of the first-generation MBRs, apart from a few examples such as installations at the basement level of skyscrapers in Tokyo, Japan, for wastewater reuse in flushing toilets, they usually found applications only in niche areas with special needs such as isolated trailer parks or ski resorts. The breakthrough for the MBRs occurred in 1989, the process involved submerging the membranes in the reactor itself and withdrawing the treated water through the membranes (Yamamoto et al., 1989; Kayawake et al., 1991; Chiemchaisri et al., 1993; Visvanathan et al., 1997). In this development, membranes were suspended in the reactor above the air diffusers (Figure 5(c)). The diffusers provided the oxygen necessary for treatment to take place and scour the surface of the membrane to remove deposited solids. There have been other parallel attempts to save energy in membrane-coupled bioreactors. In this regard, the use of jet aeration in the bioreactor was investigated (Yamagiwa et al., 1991). The main feature of this process was that the membrane module was incorporated into the liquid recirculation line for the formation of the liquid jet such that aeration and filtration could be accomplished using only a single pump. Jet aeration works on the principle that a liquid jet, after passing through a gas layer, plunges into a liquid bath entraining a considerable amount of air. Using only one pump makes it mechanically simpler and therefore useful to small communities. The limited amount of oxygen transfer possible with this technique, however, restricts this process only to such small-scale applications. The invention of air-backwashing techniques for membrane declogging led to the development of using the membrane itself as both clarifier and air diffuser (Parameshwaran and Visvanathan, 1998). In this approach, two sets of membrane modules are submerged in the aeration tank. While the permeate was extracted through one of the sets, the other set was supplied with compressed air for backwashing. The cycle was repeated alternatively, and there was a continuous airflow into the aeration tank, which was sufficient to aerate the mixed liquor. Eventually, two broad trends have emerged in recent times, namely submerged MBRs and sidestream MBRs. Submerged technologies tend to be more cost effective for largerscale lower-strength applications, and sidestream technologies are favored for smaller-scale higher-strength applications. The sidestream MBR envelope has been extended in recent years by the development of the air-lift concept, which
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Figure 5 Evolution of membrane use in conjunction with bioreactor.
bridges the gap between submerged and cross-flow sidestream MBR, and may have the potential to challenge submerged systems in larger-scale applications (Pearce, 2008b). The economic viability of the current generation of MBRs depends on the achievable permeate flux, mainly controlled by effective fouling control with modest energy input (typically r1 kW h1 m3 product). More efficient fouling-mitigation methods can be implemented only when the phenomena occurring at the membrane surface are fully understood. Detailed discussion on the technology bottlenecks and the design aspects are provided in Sections 4.16.4 and 4.16.5, respectively. It is worth noting that as the oxygen supply limits maximum mixed-liquor suspended solids (MLSSs) in aerobic MBR, anaerobic MBRs (AnMBRs) were also developed. The first test of the concept of using membrane filtration with anaerobic treatment of wastewater appears to have been reported by Grethlein (1978). The first commercially available AnMBR was developed by Dorr-Oliver in the early 1980s for high-strength whey-processing wastewater treatment. The process, however, was not applied at full scale, possibly due to high membrane costs (Sutton et al., 1983). The Ministry of International Trade and Industry (MITI), Japan, launched a 6-year research and development (R&D) project named Aqua-Renaissance ’90 in 1985 with the particular objective of developing energy-saving and smaller footprint water-
treatment processes utilizing sidestream AnMBR to produce reusable water from industrial wastewater and sewage. However, a high cross-flow velocity and frequent physicochemical cleaning was required to maintain the performance of such a high-rate MBR (Yamamoto, 2009). It was difficult to reduce the energy consumption significantly by adopting the sidestream operation using a big recirculation pump. On the other hand, commencing in 1987, a system known as anaerobic digestion ultrafiltration (ADUF) was developed in South Africa for industrial wastewater treatment (Ross et al., 1992). This process is currently in operation. Further details on AnMBRs can be derived from the comprehensive review by Liao et al. (2006). This chapter, however, focuses on aerobic MBRs.
4.16.3.2 Process Comparison with Conventional Activated Sludge Process Some important basic characteristics of CAS and MBR are compared in this section.
4.16.3.2.1 Treatment efficiency/removal capacity The MBR process involves a suspended growth-activated sludge system that utilizes microporous membranes for solid/ liquid separation in lieu of secondary clarifiers. The biological treatment in MBR is performed according to the principles
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known from activated sludge treatment. However, higher suspended solids, biological oxygen demand (BOD), and chemical oxygen demand (COD) removals in MBR have been reported throughout the literature. With CAS, the colloidal fraction (that represents about 20% of the organic content of wastewater) has a residence time (hydraulic residence time (HRT)) in the range of few hours while with MBR, due to total SS retention, the residence time of this fraction (sludge retention time (SRT)) is in the range of several days. Thus, the biodegradation for this fraction is higher in MBR than in CAS. Some soluble compounds too, after being adsorbed on SS, can be retained in MBR and can be biodegraded to a better extent. Thus, some studies have ascribed the better removal of soluble COD in MBR to the fact that the effluent is particle free (Cote et al., 1997; Engelhardt et al., 1998; De Wilde et al., 2003). MBR produces quality effluent suitable for reuse applications or as a high-quality feedwater source for RO treatment. Indicative output quality includes suspended solids o1 mg l1, turbidity o0.2 nephelometric turbidity unit (NTU), and up to 4 log removal of virus (depending on the membrane nominal pore size). In addition, it provides a barrier to certain chlorine-resistant pathogens such as Cryptosporidium and Giardia. In comparison to the CAS process, which typically achieves 95%, COD removal can be increased to 96–99% in MBRs (Stephenson et al., 2000). Nutrient removal is one of the main concerns in modern wastewater treatment especially in areas that are sensitive to eutrophication. As in the CAS, currently, the most widely applied technology for N removal from municipal wastewater is nitrification combined with denitrification. Total nitrogen removal through the inclusion of an anoxic zone is possible in MBR systems. Besides phosphorus precipitation, enhanced biological phosphorus removal (EBPR) can be implemented, which requires an additional anaerobic process step. Some characteristics of MBR technology render EBPR in combination with post-denitrification as an attractive alternative that achieves very low nutrient effluent concentrations (Drews et al., 2005b).
4.16.3.2.2 Sludge properties and composition The presence of a membrane for sludge separation has many consequences. This influences the rheological properties and composition of the sludge. Defrance et al. (2000) observed in a sidestream MBR with high cross-flow velocity that MBR sludge was less viscous than conventional sludge. The same was observed by Rosenberger et al. (2002). Furthermore, with increasing shear rate, viscosity of the sludge decreases (Rosenberger et al., 2002), although in some cases, the activated sludge behaves as a Newtonian fluid (Xing et al., 2001). Defrance and Jaffrin (1999) found out that filtering-activated sludge from an MBR resulted in fouling that could be totally, physically removed, whereas filtration of CAS led to physically irremovable fouling. It is quite difficult to generalize information about sludge composition from different installations, since each installation promotes different types of activated sludge. This has its effect on the microbial community that can be found in an activated sludge system. Nevertheless, it is obvious that the presence of the membrane in an MBR system influences the biomass composition. Since no suspended solids are washed
out with the effluent, the only sink is surplus sludge. From a secondary clarifier, lighter species are washed out, whereas in an MBR they are retained in the system by the membrane. Furthermore, changes in SRT and higher MLSS concentrations might lead to changes in the microbial community. Microbialcommunity analyses have revealed significant differences between CAS system and an MBR and a higher fraction of bacteria was found in the nongrowing state in the MBR (Witzig et al., 2002; Wagner and Rosenwinkel, 2000).
4.16.3.2.3 Sludge production and treatment Small-scale laboratory studies revealed a great advantage of MBRs, that is, lower or even zero excess sludge production, caused by low loading rates and high SRTs (Benitez et al., 1995). When longer SRTs are applied, sludge production, of course, decreases in the MBR (Wagner and Rosenwinkel, 2000). However, the amount of excess secondary sludge produced in larger MBR installations operated under the practical range of SRTs is somewhat lower than or even equal to that in conventional systems (Gu¨nder and Krauth, 2000). Table 1 provides a general comparison of the sludge-production rates from different treatment processes. It should be noted that the primary sludge production in the case of the MBR is lower. The suited pretreatment for the MBR is grids and/or sieves, and in an average, screened water was observed to contain 30% more solids than settled water (Jimenez et al., 2010). MBR sludge treatment is almost the same compared to CAS systems. The dewaterability of waste-activated sludge from the MBR seems to pose no additional problem, compared to aerobic stabilized waste sludge from CAS systems (Kraume and Bracklow, 2003).
4.16.3.2.4 Space requirements One of the advantages of the MBR is its compactness, because large sedimentation tanks are not needed. An interesting parameter in this respect is the surface-overflow rates for the two systems. The overflow rate of a secondary clarifier is defined as the ratio of its flow and footprint, that is, the volume of water that can be treated per square meter of tank. In practice, values around 22 m d1 are used. For an MBR filtration tank, an overflow rate can also be estimated from the permeate flux and the membrane-packing density within the
Table 1
Sludge production in case of different treatment processes
Treatment process
Sludge production kg (kg BOD)1
Submerged MBR Structured media biological aerated filter Trickling filter Conventional activated sludge Granular media BAF
0.0–0.3 0.15–0.25 0.3–0.5 0.6 0.63–1.06
Data from Stephenson T, Judd S, Jeferson B, and Brindle K (2000) Membrane Bioreactors for Wastewater Treatment. London: IWA. Gander MA, Jefferson B, and Judd SJ (2000) Membrane bioreactors for use in small wastewater treatment plants: Membrane materials and effluent quality. Water Science and Technology 41: 205–211, and Metcalf and Eddy, Inc. (2003) Wastewater Engineering – Treatment and Reuse, 4th edn. New York: McGraw-Hill.
Membrane Biological Reactors
tank. Following this method, Evenblij et al. (2005a) showed that with an average permeate flux of 15 l m2 h1, the overflow rates of the membrane tanks are in the range 25–62 m d1 which is up to 3 times higher than the overflow rate of a conventional secondary clarifier. Compared to an average overflow rate of 22 m d1 with a secondary clarifier, the space consumption for sludge-water separation in an MBR is 10–60% lower when flux is 15 l m2 h1 and 50–80% lower when flux is 25 l m2 h1. A further reduction in footprint is caused by the higher MLSS concentration that can be applied in an MBR. This estimate however did not take into account backflushing or relaxation periods, which reduce the overflow rate. Nevertheless, full-scale MBR plants also manifest these space-saving characteristics. For instance, Brescia WWTP, in Italy, which is the world’s largest MBR retrofit of an existing conventional plant, gives a full-scale example of a ratio of 2 when comparing area needed by CAS and MBR (Brepols et al., 2008).
4.16.3.2.5 Wastewater treatment cost
Relative cost (1994 cost equals 1)
The high cost connected with MBR is often mentioned in discussions about applicability of MBR. However, it is not easy to make a general economical comparison between MBR and CAS systems. First of all, the reference system should not simply be an activated sludge system, but a system that produces an effluent of the same quality. Moreover, an MBR is a modular system, that is, easily expandable, which is often mentioned as an advantage of the system. However, this makes the system less competitive with conventional systems, since these become relatively less expensive per population equivalent (p.e.) at larger scale. It should be noted that although the equipment and energy costs of an MBR are higher than systems used in conventional treatment, total water costs can be competitive due to the lower footprint and installation costs (Pearce, 2008b; Lesjean et al., 2004; Cote et al., 2004; De Wilde et al., 2003). MBR costs have declined sharply since the early 1990s, falling typically by a factor of 10 in 15 years. As MBR technology has become accepted, and the scale of installations has increased, there has been a steady downward trend in membrane prices (Figure 6), which is still continuing. This is particularly notable with the acceptance of the MBRs in the municipal sector. The uptake of membrane technology for municipal applications has had the affect of
1.0 Membrane cost (per unit flow rate) 0.8
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downward pressure on price. A detailed holistic cost comparison may reveal reasonably comparable results between the cost of the MBR option versus other advanced treatment options, especially if land value is considered. Studies show that depending on the design and site-specific factors the total water cost associated with MBR may be less or higher than the CAS-UF/MF option. For example, a cost comparison by the US consultant HDR in 2007 showed that MBR was 15% more expensive on a 15 million liters a day (MLD) case study, whereas a study by Zenon in 2003 gave MBR 5% lower costs (Pearce, 2008a). The differences were due to the design fluxes assumed and the capital charge rate for the project. Neither study allocated a cost advantage from the reduced footprint, which could typically translate to a treated water cost saving of up to 5%. It is interesting to evaluate the development in cost estimates over the past several years. Davies et al. (1998) made a cost comparison for two wastewater treatment plants (WWTPs), with capacities of 2350 and 37 500 p.e. With the assumptions they made (e.g., a membrane lifetime of 7 years) they conclude that depending on the design capacity (i.e., 2 times DWF to be treated) MBR is competitive with conventional treatment up to a treatment capacity of 12 000 m3 d1 (Table 2). Engelhardt et al. (1998) after carrying out pilot experiments also made a cost calculation for an MBR with a capacity of 3000 p.e., designed for nitrification/denitrification and treatment of 2*DWF. Investment costs were estimated at h3104 000 (including pretreatment) and operational cost at h194 000 yr1. Adham et al. (2001) made a cost comparison between MBR oxidation ditch followed by membrane filtration and CAS followed by membrane filtration. They concluded that MBR is competitive with the other treatment systems (Table 3). Chang et al. (2001) report experiments with low-cost membranes. The effect of membrane cost on the investment cost is considerable, but operational problems hinder further application of low-cost membranes. A drawback of the applied membranes is its limited disinfecting capacity. Van Der Roest et al. (2002a) described a cost comparison between an MBR installation and a CAS system with tertiary sand filtration. The calculations were carried out for two new WWTPs with the aim of producing effluent with low
Table 2 Capital and operating cost ratios of MBR and conventional activated sludge (CAS) process assuming a capacity of 2*(dry weather flow) Parameter
Cost ratio (MBR:ASP)
Capital cost 2350 p.e 37 500 p.e Operating costs per year
0.63 2.00
2350 p.e 37 500 p.e
1.34 2.27
0.6 0.4 0.2 0.0 (1994)
(1995)
(1997)
(1999)
(2000)
Figure 6 Sharp cost decline of membranes for MBR (cost of Zenon membranes as an example).
Data from Davies WJ, Le MS, and Heath CR (1998) Intensified activated sludge process with submerged membrane microfiltration. Water Science and Technology 38(5): 421–428.
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concentrations of nitrogen and phosphorus. Almost the same investment costs and 10–20% higher operating costs, depending on the capacity of the plant, for MBR were estimated (Table 4). Cost differences between an MBR and a traditional WWTP concerning manpower, chemical consumption, and sludge treatment were noted to be minimal. WERF (2001) summarized operating and water-quality data obtained over 1 year from two MBR pilot plants located at the Aqua 2000 Research Center at the City of San Diego (California) North City Plant. Preliminary cost estimates of the MBR technology were also developed. MBRs demonstrated that their effluent was suitable to be fed directly into an RO process from a particulate standpoint with silt density index (SDI) values averaging well below 3. The MBR effluent water quality was superior to the quality of a full-scale tertiary conventional WWTP. The preliminary cost estimate in this report was performed for a 1 million gallons a day (mgd) scalping facility (WWTP drawing a designated amount of flow from the sewer system; excess sewage flow is treated at another plant located at the end of the sewer line). This facility produced an effluent suitable as feedwater for an RO process. Based upon this estimate, the present value was estimated as $0.81 m3, $0.96 m3, and $1.16 m3 for the MBR process, oxidation ditch with MF, and oxidation ditch with conventional tertiary lime pre-treatment, respectively. Therefore, the MBR process was reported as the most cost-effective alternative for water reclamation where demineralization or indirect drinking water-production (RO) is required. McInnis (2005) reported a detailed comparative cost analysis of two membrane-based tertiary treatment options: (1)
Table 3 Capital and total cost ratios of MBR and tertiary MF following alternative biological processes Alternatives
Oxidation ditch-MF CAS-MF
Cost ratio (MBR:alternative) Capital
Total per year
0.91 0.85
0.89 0.9
Data from Adham S, Mirlo R, and Gagliardo P (2000) Membrane bioreactors for water reclamation – phase II. Desalination Research and Development Program Report No. 60, Project No. 98-FC-81-0031. Denver, CO: US Department of the Interior, Bureau of Reclamation, Denver Office.
MBR and, (2) CAS process followed by MF (CAS/MF). According to that study, irrespective of design flow rate, the MBR entails slightly higher unit capital costs as compared to CAS/ MF process, while, depending on the design flow rate, the operation and maintenance costs (O&M) of the former are higher than or comparable to that of the latter. Comparative O&M cost breakdown revealed that MBR entails less labor cost, considerably higher power and chemical consumptions and slightly higher membrane cost, other costs remaining virtually the same. In the CAS/MF process, labor cost induces the highest cost, while in case of the MBR process, labor and electrical power-consumption costs are almost similar. Overall, the MBR imposes slightly higher capital and operating/ maintenance cost over that of CAS/MF. Cote et al. (2004) explored two membrane-based options available to treat sewage for water reuse, tertiary filtration (TF) of the effluent from a CAS process, and an integrated MBR. These options were compared from the point of view of technical performance and cost using ZeeWeed immersed membranes. The analysis showed that an integrated MBR is less expensive than the CAS-TF option. The total life cycle costs for the treatment of sewage to a quality suitable for irrigation reuse or for feeding RO decreased from 0.40$ m3 to 0.20$ m3 as plant size increased to 75 000 m3 d1. It was also shown that the incremental life-cycle cost to treat sewage to indirect potable water-reuse standards (i.e., by UF and RO) was only 39% of the cost of seawater desalination. A recent market research report (BCC Research, 2008) estimated the capital cost of a 50 000 gallons per day (gpd) (190 m3 d1) plant at US$350 000, a 100 000 gpd plant at US$500 000, and a 500 000 gpd plant at US$2 million. For systems of 1 mgd (million gallons per day) and larger, capital costs start at US$3.5 million (Table 5). The largest percentage of new system installations, 93%, continue to fall into the 5000–500 000 gpd range (most of those, about 57% of them, have capacities of less than 25 000 gpd), 2% of installations range from 0.5–1 mgd, and 5% of them are larger than 1 mgd. Tables 2–5 list cost values reported during the period 1998–2008. Obviously, the data from the initial stage of the MBR development holds little relevance today. However, these are listed here to provide a general trend of cost-data evolution.
4.16.3.2.6 Comparative energy usage Table 4 Capital and total cost ratios of MBR and tertiary sand filtration following CAS Parameter
Cost ratio
Capital cost 10 000 p.e 50 000 p.e Operating costs per year
0.92 1.01
10 000 p.e 50 000 p.e
1.09 1.21
Data from Van Der Roest HF, Lawrence DP, and Van Bentem AG (2002a) Membrane Bioreactors for Municipal Wastewater Treatment (Water and Wastewater Practitioner Series: Stowa Report). London: IWA.
MBR provides an equivalent treatment level to CAS-UF/MF, but at the expense of higher energy cost since the efficiency of air usage in MBR is relatively low. The MBR process uses more Table 5
Capital cost of MBR depending on plant sizea
Plant size, gpd 103
Capital cost, US$ 103
50 100 500 1000
350 500 2000 3500
a
1 m3 d1 ¼ 264.17 gpd. Data from BCC Research (2008) Membrane bioreactors: Global markets. Report Code MST047B, Report Category – Membranes & Separation Technology.
Membrane Biological Reactors
air, and hence higher energy than conventional treatment. This is because aeration is required for both the biological process and the membrane cleaning, and the type, volume, and location of air required for the two processes are not matched. Biotreatment utilizes fine air bubbles, since oxygen needs to be absorbed for the biological reaction step. In contrast, fouling control is best achieved by larger bubbles, since the air is required to scour the membrane surface or shake the membrane to remove the foulant. Accordingly, although the concept of MBR was first developed to exploit the fact that the biological wastewater-treatment process and the process of membrane-fouling control can both use aeration (Pearce, 2008b), the potential for dual-purpose aeration is strictly limited. Based on a survey of conventional wastewater-treatment facilities in the US, Metcalfe and Eddy, Inc. (2003) reported that the energy usage range was 0.32–0.66 kW h1 m3. Energy usage in wastewater treatment is somewhat lower in Europe, partly due to a greater consciousness for energy efficiency, and partly due to the fact that average BOD loading/ capita in the US is 20–25% greater than that in Europe (due to the use of kitchen disposal units). Long-term monitoring of wastewater-treatment systems has shown usages as low as 0.15 kW h1 m3 for activated sludge, increasing to 0.25 kW h1 m3 if a biological aerated filter (BAF) stage is included (Pearce, 2008a). Membrane filtration after conventional treatment is estimated to add 0.1–0.2 kW h1 m3 to the energy, equivalent to a total energy use for CAS-UF/MF of 0.35–0.5 kW h1 m3 in a new facility (Lesjean et al., 2004). Experience in large-scale commercial MBRs shows an energy usage of around 1.0 kW h1 m3, although smaller-scale facilities typically operate at 1.2–1.5 kW h1 m3 or higher (Judd, 2006). However, in comparison to these values, energy consumption of around 1.9 kW h1 m3 was reported in 2003 (Zhang et al., 2003) and up to 2.5 kW h1 m3 in 1999 (Ueda and Hata, 1999). This proves that there is a gradual improvement in MBR design (Figure 7). Further improvements in air efficiency and membrane-packing density are expected
Energy consumption, kW hr−1 m−3
3.0
2.0
1.0
to improve the current values in the future. Even so, it seems likely that MBR energy costs will continue to exceed those of CAS-UF/MF by 0.4 kW h1 m3 or more (Pearce, 2008a). However, the fact that membrane filtration after conventional treatment is estimated to add only 0.1–0.2 kW h1 m3 to the energy points out that the higher energy consumption of MBR over CAS-UF/MF is due to the difference in consumption in the respective biological processes. MBRs are generally operated at quite low F/M ratios (less than 0.2), or high MLSS concentrations, and this is one of the reasons for the excellent biodegradation efficiency, and high aeration cost as well. CAS plants, on the other hand, are operated at higher F/M ratios, implying lower oxygen need for biodegradation. Table 6 lists typical energy-use rates of different biologicalbased treatment combinations. Section 4.16.5 provides further information on energy comparison of the MBR formats.
4.16.3.3 Relative Advantages of MBR There are several advantages associated with the MBR technology, which make it a valuable alternative over other treatment techniques. The combination of activated sludge with membrane separation in the MBR results in efficiencies of footprint, effluent quality, and residual production that cannot be attained when these same processes are operated in sequence. The MBR system is particularly attractive when applied in situations where long biological solid-retention times are necessary and physical retention and subsequent hydrolysis are critical to achieving biological degradation of pollutants (Chen et al., 2003). The prime advantages of MBR are the treated water quality, the small footprint of the plant, less sludge production, and flexibility of operation (Visvanathan et al., 2000). First of all, the retention of all suspended matter and most of the soluble compounds within the bioreactor leads to excellent effluent quality capable of meeting stringent discharge requirements and paving the way for direct water reuse. The possibility of retaining all bacteria and viruses results in a sterile effluent, eliminating extensive disinfection and the corresponding hazards related to disinfection by-products. As the entire process equipment can be made airtight, odor dispersion can be prevented quite successfully. Since suspended solids are not lost in the clarification step, total separation and control of the SRT and hydraulic retention time (HRT) are possible enabling optimum control of the microbial population and flexibility in operation. The absence of a clarifier, which also acts as a natural selector for settling organisms, enables sensitive, slow-growing
0.0 1999
2003 Year
2006
Figure 7 Gradual reduction in reported values of energy consumption by MBR. Data from Ueda T and Hata K (1999) Domestic wastewater treatment by a submerged membrane bioreactor with gravitational filtration. Water Research 33: 2888–2892; Zhang SY, Van Houten R, Eikelboom DH, et al. (2003) Sewage treatment by a low energy membrane bioreactor. Bioresource Technology 90: 185–192; and Judd S (ed.) (2006) The MBR Book: Principles & Applications of MBRs in Water & Wastewater Treatment. Oxford: Elsevier.
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Table 6 Comparative typical energy consumption by different treatment options Treatment option
Energy use (kW h1 m3)
CAS CAS-BAF CAS-MF/UF MBR
0.15 0.25 0.35–0.5 0.75–1.5a
a
Power consumption range for large- to smaller-scale plants.
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species (nitrifying bacteria, bacteria capable of degrading complex compounds) to develop and persist in the system (Cicek et al., 2001; Rosenberger et al., 2002). The membrane not only retains the entire biomass but also prevents the escape of exocellular enzymes and soluble oxidants creating a more active biological mixture capable of degrading a wider range of carbon sources (Cicek et al., 1999b). MBRs eliminate process difficulties and problems associated with settling, which is usually the most troublesome part of wastewater treatment. The potential for operating the MBR at very high SRTs without the obstacle of settling allows high biomass concentrations in the bioreactor. Consequently, higher-strength wastewater can be treated and lower biomass yields are realized (Muller et al., 1995). This also results in more compact systems than conventional processes, significantly reducing plant footprint and making it useful in waterrecycling applications (Konopka et al., 1996). The low sludge load in terms of BOD forces the bacteria to mineralize poorly degradable organic compounds. The higher biomass loading also increases shock tolerance, which is particularly important where feed is highly variable (Xing et al., 2000). The increased endogenous (autolytic) metabolism of the biomass (Liu and Tay, 2001) under long SRT allows development of predatory and grazing communities, with the accompanying trophiclevel energy losses (Ghyoot and Verstraete, 1999). These factors, in addition to resulting in lower overall sludge production, lead to higher mineralization efficiency than those of a CAS process. High molecular weight soluble compounds, which are not readily biodegradable in conventional systems, are retained in the MBR (Cicek et al., 2002). Thus, their residence time is prolonged and the possibility of oxidation is improved. The system is also able to handle fluctuations in nutrient concentrations due to extensive biological acclimation and retention of decaying biomass (Cicek et al., 1999a).
4.16.3.4 Factors Influencing Performance/Design Considerations This section sheds light on some important design considerations of MBR. More detailed information on some of these parameters is provided in Section 4.16.4.7, in relation to membrane fouling.
4.16.3.4.1 Pretreatment All MBRs require pretreatment, for example, screening and grit removal, to protect the membranes. Screening has historically been limited to 3 mm; however, hair and fiber can still pass through this size of the screen and become embedded or wrapped around the hollow fibers. The MBR providers have standardized their screen selections to a 2-mm traveling band, punched screen. Conversely, the flat-sheet membranes experience less problems with hair and fiber, and are standardized to a 3-mm screen. Further discussion regarding mechanical pretreatment is provided in Section 4.16.4.6.
4.16.3.4.2 Membrane selection and applied flux An MBR membrane needs to be mechanically robust, chemically resistant to high Cl2 concentrations used in cleaning, and nonbiodegradable (Pearce, 2008a). Clean-water permeability
is not as important in an MBR as in membrane-filtration applications, since the membrane transport properties are strongly influenced by the accumulation of foulant particles at the membrane surface. However, process flux in treating a wastewater feed is important since it directly affects capital cost, due to its effect on membrane area and footprint, and operating costs due to the effect of membrane area on chemical and air use. Most MBRs operate at an average flux rate between 12.5 and 25 l m2 h1, with Mitsubishi’s unit operating in the lower range. The key flux rates that determine the number of membranes required are associated with the peak flow rates. For plants with peaking factors of less than two, an MBR can handle the plant flow variation without having a significantly impact on the average design flux rate. Otherwise, equalization needs to be provided with either a separate tank at the head of the facility or within the aeration basin, allowing sidewater depth variations during peak flow.
4.16.3.4.3 Sludge retention time In the past, most MBR systems were designed with extremely long SRTs, of the order of 30–70 days, and very few were operated at less than about 20 days. Two reasons prompted such practice: (1) the drive to minimize sludge production or eliminate it all together and (2) the concern over the reduced flux resulting from short SRT operation, presumably due to the fouling effect of extracellular excretions from younger sludge. Currently, the selection of SRT is based more on the treatment requirements, and SRTs as low as 8–10 days can now be contemplated.
4.16.3.4.4 Mixed liquor suspended solids concentration From the point of view of bioreactor volume reduction and minimization of excess sludge, submerged MBR systems have been typically operated with MLSS concentrations of more than 12 000 mg l1, and often in the range of 20 000 mg l1. Hence, they offer greater flexibility in the selection of the design SRT. However, excessively high MLSS may render the aeration system ineffective and reduce membrane flux. A trade-off, therefore, comes into play. Current design practice is to assume the MLSS to be closer to 10 000 mg l1 to ensure adequate oxygen transfer and to allow for higher membrane flux. With larger systems, it is more cost effective to reduce the design MLSS because of the high relative cost of membranes when compared to the cost of additional tank volume.
4.16.3.4.5 Oxygen transfer At high MLSS concentrations, the demand for oxygen can be significant. In some cases, the demand can exceed the volumetric capacity of typical oxygenation systems. The oxygentransfer capacity of the aeration system must also be carefully analyzed. Submerged membranes are typically provided with shallow coarse bubble air to agitate the membranes as a means to control fouling. Such aeration provides some oxygenation, but at low efficiency. In compact systems, fine bubble aeration may be placed at greater depth below the membrane aeration; however, the combined efficiency and the bubblecoalescing effects require further consideration during design (Visvanathan et al., 2000).
Membrane Biological Reactors
The lower operating cost obtained with the submerged configuration along with the steady decrease in the membrane cost encouraged an exponential increase in MBR plant installations from the mid-1990s onward. Since then, further improvements in the MBR design and operation have been introduced and incorporated into larger plants. The key steps in the recent MBR development are summarized below:
• •
•
The acceptance of modest fluxes (25% or less of those in the first generation), and the idea of using two-phase bubbly flow to control fouling. While early MBRs were operated at SRTs as high as 100 days with MLSS up to 30 g l1, the recent trend is to apply a lower SRT (around 10–20 days), resulting in more manageable MLSS levels (10–15 g l1). Thanks to these new operating conditions, the fouling propensity in the MBR has tended to decrease and overall maintenance has been simplified, as less-frequent membrane cleaning is necessary.
Further discussion on these aspects is provided in the following sections.
4.16.4 Worldwide Research and Development Challenges 4.16.4.1 Importance of Water Reuse and the Role of MBR The need for pure water is a problem of global proportions. In the Earth’s hydrologic cycle, freshwater supplies are fixed and constant, while global water demand is growing (Howell, 2004; Bixio et al., 2006). With each passing year, the quality of the planet’s water measurably deteriorates, presenting challenges for the major users: the municipal, industrial, and environmental sectors. Increasing demand for water, and drought and water scarcity are now common issues facing many urban and rural communities around the world (Howell, 2004; Tadkaew et al., 2007; Jimenez and Asano, 2008). Water treatment has, therfore, become an area of global concern as individuals, communities, industries, countries, and their national institutions strive for ways to keep this essential resource available and suitable for use. Water recycling is a pragmatic and sustainable approach for many countries to mitigate or solve the problems of water supply. There is a growing interest in using nontraditional water resources by means of water reclamation and water recycling for long-term sustainability. It can be divided into two categories, internal domestic or industrial recycling and external recycling, where high-quality reclaimed water from a sewage treatment plant is used for aquifer recharge or irrigation. With the current focus on water-reuse projects and the role they play in the water cycle, the search for cost-competitive advanced wastewater-treatment technologies has never before been so important. Treatment technology for water recycling encompasses a vast number of options. A general paucity of legislative and socioeconomic information has led to the development of a diverse range of technical solutions (Jefferson et al., 2000). Membrane processes are regarded as key elements of advanced wastewater reclamation and reuse schemes and are included in a number of prominent schemes worldwide, for example, for artificial groundwater recharge, indirect
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potable reuse, as well as for industrial-process water production (Melin et al., 2006; Bixio et al., 2008). Among the many treatment alternatives, MBRs, which combine membrane filtration and biological process for wastewater treatment, are seen to have an effective technology capable of transforming various types of wastewater into high-quality effluent exceeding most discharge requirements and suitable for a variety of nonpotable water-reuse applications such as flushing toilets and for irrigation (Tadkaew et al., 2007; Jimenez and Asano, 2008). In some cases, treated water can be applied to recharge groundwater to halt saltwater intrusion into coastal aquifers, abate subsidence in areas sinking due to overpumping groundwater, and support aquifer storage and recovery. Issues of water quality, water quantity, and aging/nonexistent infrastructure propel the market for MBRs. Escalating water costs due to dwindling supplies for communities and businesses also drive the growing acceptance of MBRs. Anticipated stricter environmental regulations are driving sales of MBRs to industry, municipalities, and are prompting maritime users to consider MBR technology (Jefferson et al., 2000; Jimenez and Asano, 2008). This is probably due to the effectively disinfected high-quality effluent and high performance in trace organic removal for safe and environmentally benign discharge that MBRs can offer. In practical terms, the process has many benefits, which make it suitable for the size of the systems applicable to recycling. The ability to run independently of load variation and produce no sludge are critical and highlight MBRs as possibly the most viable small-footprint, high-treatment option for water recycling (Jefferson et al., 2000; Melin et al., 2006; Tadkaew et al., 2007). Comparison with other technologies used for water recycling reveals that MBRs not only produce lower residual concentrations but do so more robustly than the alternatives (Jefferson et al., 2000; Melin et al., 2006). The favorable microbiological quality of the effluent of MBRs is a major factor in their frequent selection for water reuse, even if full disinfection cannot be expected, particularly considering the distribution and storage components of a full-scale system, which can be prone to regrowth of microorganism and contamination from various sources. However, the MBR effluent is adequate for many water-reuse applications with little residual chlorine disinfection for subsequent distribution. The MBR then does provide a dual layer of protection against pathogen breakthrough, greatly lowering the risk during operation. MBRS have the greatest efficacy toward water recycling, albeit contingent upon a loading rate constrained by the operable flux. Not only do they comply with all likely waterquality criteria for domestic recycling but they also produce a product that is visibly clear and pathogen free, both of which are likely to be key concerns in terms of public acceptability. There are some issues that still need to be addressed and these are highlighted throughout Sections 96.4.6 and 96.4.7 of this chapter.
4.16.4.2 Worldwide Research Trend Early development efforts in MBR technology were concentrated in UK, France, Japan, and South Korea, whereas extensive research in China and Germany began after 2000. Much
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of the research in the newcomer countries is building on pioneering work from the UK, France, Japan, and South Korea. Three stages may be identified in the worldwide MBR research: 1. An entry-level stage spanning from 1966 to 1980, during which lab-scale research was mainly conducted. Membranes of that period had low flux and short life span due to undeveloped membrane-manufacturing technology. 2. A slow-to-moderate growth period from 1980 to 1995, when MBR technology was well investigated especially in Japan, Canada, and the USA. During this stage, new membrane-material development, MBR configuration design, and MBR operation were critically studied. The submerged MBR concept was put forward by Japanese researchers in 1989. 3. The rapid development stage started in 1995 and continues even now, when MBR technology underwent a rapid development prompted by deep understanding of the technology in research communities and by the installation of full-scale MBRs. Much of the published information on MBRs to date has mainly focused on bench or pilot-scale studies, performance results of treating a specific type of wastewater, and short-term operations. Regardless of the source of wastewater, whether it is municipal or industrial, very few publications involved fullscale studies for long-term operational periods. In a comprehensive review, Yang et al. (2006) grouped the available worldwide publications regarding MBR into six main research areas: (1) literature and critical reviews; (2) fundamental aspect; (3) municipal and domestic wastewater treatment; (4) industrial wastewater and landfill leachate treatment; (5) drinking-water treatment; and (6) others, which include gas removal, sludge treatment, hydrogen production, and gas diffusion. The fundamental research category was based on studies that exclusively looked at membrane fouling, operation and design parameters, sludge properties, microbiological characteristics, cost, and modeling. Studies, which focused on applied research and general reactor performance, were categorized by influent (feed) type (groups 3–6). Membrane fouling, which has been widely considered as one of the major limitations to faster commercialization of MBRs, has been investigated from various perspectives including the causes, characteristics, mechanisms of fouling, and methods to prevent or reduce membrane fouling. More than one-third of studies in the fundamental aspects group were found to deal with issues related to membrane fouling.
4.16.4.3 Modeling Studies on MBR Models that can accurately describe the MBR process are important for the design, prediction, and control of MBR systems. Due to the intrinsic complexity and uncertainty of MBR processes, basic models that can provide a holistic understanding of the technology at a fundamental level are of great necessity. Complex models that are also practical for real applications can greatly assist in capitalizing on the benefits of MBR technology. However, compared to experimental R&D, followed by commercialization of the technology, modeling studies for system-design analysis and performance prediction
are at a relatively preliminary stage. In an attempt to identify the required research initiatives in this regard, this section looks briefly into the state-of-the-art MBR modeling efforts. Effluent quality and the investment and operating costs are the primary concerns for any given wastewater treatment system. Therefore, model development should center on components for which water-quality standards have been set and parameters which are strongly correlated to cost. Ng and Kim (2007) put forward a few key model components and parameters for MBR modeling:
•
•
•
•
•
The ability to quantify individual resistance (i.e., resistance from cake formation, biofilm formation, and adsorptive fouling) as a function of the various influencing parameters is important in determining which parameters have the greatest influence on fouling and for designing and optimizing the system to achieve an economical balance between production and applied pressure. Determining the relationship between biomass concentration and other parameters can aid in identifying an optimal biomass concentration for operation, which can lead to significant economical savings. Aeration accounts for a significant portion of energy costs in the operation of MBR systems. The factors that influence oxygen requirement (wastewater and biomass concentration/growth rates) and the oxygen-transfer rate (MLSS concentration, MBR configuration, type of bubbles used, and specific airflow rate) should receive due consideration in the model to optimize aeration. Carbon and nutrient (nitrogen and phosphorous components) concentrations and their influencing factors (e.g., respective concentrations and growth rates of the various types of organisms and concentration of oxygen) should be incorporated into the models. Soluble microbial products (SMPs), which comprise a major portion of the organic matter in effluents from biological treatment processes and are potentially associated with issues such as disinfection by-product formation, biological growth in distribution systems, and membrane fouling, should be given proper consideration in models.
MBR models available in the literature can be broadly classified into three categories: biomass kinetic models, membranefouling models, and integrated models to describe the complete MBR process (Ng and Kim, 2007; Zarragoitia-Gonza´lez et al., 2008). Models describing biomass kinetics in an MBR include the activated sludge model (ASM) family (Henze et al., 2000), the SMP model (Furumai and Rittmann, 1992; Urbain et al., 1998; de Silva et al., 1998), and the ASM–SMP hybrid model (Lu et al., 2001; Jiang et al., 2008). The ASMs were developed to model the activated sludge process. The MBR process is the activated sludge process with the secondary clarification step replaced by membrane filtration; therefore, it is reasonable to use ASMs to characterize the biomass dynamics in an MBR system. However, their ability to describe the MBR process accurately has not been verified by in-depth experiments. Research suggests that SMPs are important components in describing biomass kinetics due to high SRTs in MBR systems. Accordingly, the SMP model demonstrated the capability of
Membrane Biological Reactors
characterizing the biomass with a reasonable-to-high degree of accuracy. Lu et al. proposed that the modified versions of ASM1 (Lu et al., 2001) and ASM3 (Lu et al., 2002), which incorporate SMPs, demonstrated fairly reasonable accuracy in quantifying COD and soluble nitrogen concentrations. Jiang et al. (2008) extended the existing ASM No. 2d (ASM2d) to ASM2dSMP with introduction of only four additional SMPrelated parameters. In addition to minimizing model complexity and parameter correlations, the model parameter estimation resulted in reasonable confidence intervals. Models describing membrane fouling include the empirical hydrodynamic model (Liu et al., 2003), fractal permeation model (Meng et al., 2005), sectional resistance model (Li and Wang, 2006), subcritical fouling behavior model (Saroj et al., 2008), and the resistance-in-series models that were presented as a part of the integrated models. Some of them are simply based on solid–liquid separation and simulate filtration processes (Chaize and Huyard, 1991; Gori et al., 2004). Other models consider specific physical approaches: cross-flow filtration (Cheryan, 1998; Hong et al., 2002; Beltfort et al., 1994) and mass-transport models (Beltfort et al., 1994; Bacchin et al., 2002). Nevertheless, membrane fouling is generally evaluated by employing the resistance-in-series model (Wintgens et al., 2003; Wisniewski and Grasmick, 1998) or, rarely, using empirical models (Benitez et al., 1995; De Wilde et al., 2003). The integrated models, basically, couple the kinetic models with the fouling ones (such as the resistance-in-series model) and they often consider the formation and degradation of SMPs (Ng and Kim, 2007). The models reported to date are valuable preliminary attempts, but require further improvements. For instance, the empirical hydrodynamic model is too simple to describe the membrane-fouling phenomenon, and the sectional resistance model lacks accuracy. Both the fractal permeation model and resistance-in-series model by Lee et al. (2002) provide good scientific insight, but specific experimental verification is necessary for general use of the models. The resistance-in-series model developed by Wintgens et al. (2003) shows the most promise, as it is fairly accurate, accounts for cleaning cycles, and can predict permeability changes over time. Further tests are needed to determine whether the model requires calibration or if the model parameters are applicable to other MBR systems. Recently, Zarragoitia-Gonza´lez et al. (2008) included the biological kinetics and the dynamic effect of the sludge attachment and detachment from the membrane, in relation to the filtration and a strong intermittent aeration in a hybrid model. The model was established considering SMP formation–degradation kinetic based on previous published models (Cho et al., 2003; Lu et al., 2001). A modification of Li and Wang’s model (Li and Wang, 2006) allows to calculate the increase of the transmembrane pressure (TMP), evaluating, at the same time, the influence of an intermittent aeration of bubbles synchronized with the filtration cycles on fouling control, and to analyze the effects of shear intensity on sludge cake removal. On the other hand, in order to describe the biological system behavior, a modified ASM1 model was used. The final hybrid model was developed to calculate the evolution of sludge properties, its relation to sludge cake growth, and the influence of sludge properties on membrane fouling.
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A simple model for evaluating energy demand arising from aeration of an MBR was presented by Verrecht et al. (2008) based on a combination of empirical data for membrane aeration and biokinetic modeling for biological aeration. The model assumes that aeration of the membrane provides a portion of the dissolved oxygen needed for biotreatment. The model also assumes, based on literature information sources, a linear relationship between membrane permeability and membrane aeration up to a threshold value, beyond which permeability is unchanged with membrane aeration. An analysis reveals that significant reductions in energy demand are attained through operating at lower MLSS levels and membrane fluxes. The complete organic removal in MBR is due to all the inseries phenomena: biological degradation of biomass, biological filtration of cake layer, and final filtration of physical membrane. Di Bella et al. (2008) set up a mathematical model for the simulation of physical–biological wastewater organic removal for SMBR system. The model consists of two submodels: the first one for the simulation of the biological processes and a second one for the physical processes. In particular, regarding the biological aspects, it is based on the ASM concept. On the other hand, organic-matter removal due to filtration (the physical process) was described by simple models proposed in the literature (Kuberkar and Davis, 2000; Jang et al., 2006; Li and Wang, 2006). It is conceivable that several of the existing models, particularly the ASMs, require validation to determine their applicability for modeling the MBR process and to evaluate whether they can serve as a base for future MBR model development. Membrane fouling in MBRs is affected by the biotransformation processes in the system; therefore, a more effective integration of biomass kinetics and membrane fouling into the models is required. Moreover, examination of alternative empirical modeling approaches, such as the application of artificial neural networks, is worthwhile to establish a thorough link between inputs and outputs of MBR systems and to find phenomenological interrelationships among components and parameters (Ng and Kim, 2007).
4.16.4.4 Innovative Modifications to MBR Design Researchers have put forth different modifications to the conventional design of MBRs in order to enhance removal performance and/or mitigate membrane fouling. This section highlights some of such examples (Table 7). The commercialized MBR formats are discussed separately in Section 4.16.5.2.
4.16.4.4.1 Inclined plate MBR Theoretically, an infinite SRT provides a possibility of naturally achieving zero-excess sludge discharge from MBR under normal environment. It should, however, be noted that zeroexcess sludge production is just a theoretical concept which can only be obtained with a feed containing only solutes. In real life, sewage or industrial effluents contain nonbiodegradable suspended solids and colloids that accumulate in the reactor, continuously increasing the sludge concentration. Therefore, an immediate challenge encountered at infinite SRT is the extremely high sludge concentration
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Table 7
Examples of innovative modifications to MBR design
Modified design
Main purpose
Selected reference
Inclined plate MBR
Omit excess sludge production and thereby realize long-term stable membrane filtration Derive the simultaneous advantages of efficient nutrient removal and mitigate membrane fouling (Chae et al., 2006 a,b;). Treatment of high-strength wastewater without encountering severe fouling Enhanced removal of recalcitrant compound and/or membrane fouling mitigation Obtain in one step indirect potable reuse standard effluent Indirect potable reuse along with energy demand reduction Obtain in one step indirect potable reuse standard effluent
Xing et al. (2006)
Integrated anoxic–aerobic MBR
Jet-loop-type MBR Biofilm MBR Nanofiltration MBR Forward osmosis MBR Membrane distillation bioreactor (MDBR)
produced in the bioreactor (Wen et al., 1999). Consequently, the method to achieve zero-excess sludge discharge translates into how to realize long-term stable membrane filtration of high-concentration sludge beyond the guideline value of 10– 20 g l1 recommended for submerged MBRs when applied to domestic wastewater treatment. In order to omit excess sludge production, Xing et al. (2006) proposed an innovative MBR design comprising an anoxic tank equipped with settlingenhancer inclined plates and a subsequent aerobic tank containing the membrane. The inclined plates together with intermittent air blowing (to blow off gaseous content generated by denitrification, etc.) proved to be quite effective in confining high MLSS sludge within the anoxic tank leading to an MLSS difference of 0.1– 13.1 g l1 between the aerobic and anoxic sludge. Consequently, the capability of MBRs in handling the extremely high MLSS challenge encountered especially at zero-excess sludge could be extended. Results indicated that at an HRT of 6 h, average removals of COD, ammonia nitrogen, and turbidity were 92.1, 93, and 99.9%, resulting in daily averages of 12.6 mg COD l1, 1.3 mg NH3–N l1, and 0.03 NTU, respectively.
4.16.4.4.2 Integrated anoxic–aerobic MBR In contrast to separate anoxic tanks for denitrification or creation of alternating anoxic/oxic conditions within the same tank by intermittent aeration, an integrated anoxic/oxic MBR, containing anoxic/oxic compartments in one reactor, was developed to derive simultaneous advantages of efficient nutrient removal (Chae et al., 2006a, 2006b) and mitigated membrane fouling (Chae et al., 2006a, 2006b; Hai, 2007; Hai et al., 2007; Hai et al., 2006b; Hai et al., 2008a). Under the optimal volume ratio of anoxic and oxic zones of 0.6 and the desirable internal recycle rate and HRT of 400% and 8 h, respectively, the average removal efficiencies of total nitrogen (T-N) and total phosphorus (T-P) were 75% and 71%, respectively (Chae et al., 2006b). Furthermore, comparison with sequential anoxic/oxic MBR under the same conditions revealed the membrane-fouling reduction potential of this specific design (Chae et al., 2006a).
Chae et al. (2006a,b), Hai et al. (2006b, 2008a)
Park et al. (2005), Yeon et al. (2005) Lee et al. (2006), Leiknes and Odegaard (2007), Ngo et al. (2008), Hai et al. (2008) Choi et al. (2002) Achilli et al. (2009), Cornelissen et al. (2008) Phattaranawik et al. (2008, 2009)
Working with a high-strength industrial wastewater, Hai et al. (2006a, 2006b, 2008a) demonstrated minimization of excess sludge growth and maintenance of less MLSS concentration in contact with the membrane at the aerobic zone by exploring a similar reactor design along with a strategy of splitting the feed through the two zones.
4.16.4.4.3 Jet-loop-type MBR The so-called high-performance compact reactor (HCR) which is a jet-loop-type reactor with a draft tube and a two-phase nozzle was coupled with a submerged membrane by Park et al. (2005). The HCR is able to deal with very high organic loading rates due to the high efficiency of oxygen transfer, mixing, and turbulence achieved. The significant amount of bubbles and turbulence present in the HCR can be beneficial in retarding fouling of the submerged membrane. The developed MBR showed much greater membrane permeability than the conventional MBR, promising very high potential for the treatment of high-strength wastewater without encountering severe fouling (Park et al., 2005; Yeon et al., 2005).
4.16.4.4.4 Biofilm MBR Membrane-coupled moving-bed biofilm reactor system, wherein the membrane is submerged within the same tank (Lee et al., 2006) or in an additional tank (Leiknes and Odegaard, 2007), has been extensively studied in association with different kinds of biocarriers. Powdered activated carbon (PAC) which also acts as an adsorbent is commonly added into the bioreactor as the biocarrier (Ng et al., 2006; Hai, 2007; Hai et al., 2008b). However, carriers made of inert materials, such as plastic (Leiknes and Odegaard, 2007) and sponge (Lee et al., 2006; Ngo et al., 2008), have also been used. Biomass granulation with shell-support media coupled with membrane separation is also worth mentioning in this context (Thanh et al., 2008). The mechanisms of enhanced removal and/or membranefouling mitigation depend on the specific design and the utilized biocarrier type. For example, in an integrated membrane-coupled moving-bed biofilm reactor using sponge as the biocarrier, frictional force exerted by the circulating
Membrane Biological Reactors
carrier on the submerged membrane reduced the formation of cake layer on the membrane surface and thus enhanced the membrane permeability (Lee et al., 2006). On the other hand, Leiknes and Odegaard (2007) demonstrated that operation under high volumetric-loading rates of 2–8 kg COD m3 d1and HRTs up to 4 h and maintenance of membrane fluxes around 50 l m2 h1 were possible by placing the moving-bed biofilm reactor prior to the submerged MBR. The specific purpose of the biofilm reactor in this case was to reduce the organic loading on MBR. Ng et al. (2006) contend that the improved membrane performance of the MBR with added PAC could be due to a number of factors including, PAC providing sink for some of the fouling components and the scouring action of PAC. Hai et al. (2008b) reported that simultaneous PAC adsorption within a fungiMBR treating dye wastewater resulted in multiple advantages including co-adsorption of dye and fungal enzyme onto activated carbon and subsequent enzymatic dye degradation.
4.16.4.4.5 Nanofiltration MBR The potential for using NF technology in wastewater treatment and water reuse is noteworthy. A new concept with the addition of RO membrane after conventional MBR has been recently developed to reclaim municipal wastewater. The new MBR-RO process demonstrated the capability of producing the same or more consistent product quality (in terms of total organic carbon (TOC), NH4, and NO3) and sustained higher flux compared to the CAS-MF-RO process in reclamation of domestic sewage (Qin et al., 2006). Choi et al. (2002, 2007), on the other hand, demonstrated the technical feasibility of a submerged NF-MBR. For the initial 130 days, the NF-MBR achieved high permeate quality (DOC concentration ¼ 0.5–2.0 mg l1) and maintained reasonable water productivity. With low electrolyte rejection, operation under a low suction pressure was possible, and electrolyte accumulation in the bioreactor, which may hinder biological activity, did not occur. The permeate quality, however, deteriorated to some extent (DOC concentration ¼ 3.0 mg l1) due to the deterioration of the cellulose membrane.
4.16.4.4.6 Forward osmosis MBR The forward osmosis (FO)–MBR is an innovative technique for the reclamation of wastewater, which combines activated sludge treatment and FO membrane separation with an RO posttreatment. FO membranes, either submerged or external, are driven by an osmotic pressure difference over the membrane. Through osmosis, water is transported from the mixed liquor across the semipermeable membrane into a draw solution (DS) with a higher osmotic pressure. To produce potable water, the diluted DS is then treated in an RO unit, and the concentrated DS is reused in the FO process. The FO-MBR is expected to have the same advantages as conventional MBRs; however, it has to deal with the most important drawback, that is, a high energy demand. In this system, FO membranes with structures comparable with NF or RO membranes are used instead of MF/UF membranes for the separation of suspended solids, multivalent ions, natural organic matter, and biodegradable materials. Since fluxes are generally lower and no internal fouling occurs, fouling of NF
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or RO membranes, compared to that of the MF or UF membranes in conventional MBR, may be dealt with easily. The RO system after FO-MBR can be operated with higher fluxes because all the bivalent ions are removed in the FO-MBR. Recent studies have demonstrated high sustainable flux and relatively low reverse transport of solutes from the DS into the mixed liquor, along with very high removal performance (Achilli et al., 2009; Cornelissen et al., 2008).
4.16.4.4.7 Membrane distillation bioreactor A novel wastewater-treatment process known as the membrane distillation bioreactor (MDBR) incorporating membrane distillation in an SMBR operated at an elevated temperature was developed and experimentally demonstrated by Phattaranawik et al. (2008, 2009). The ability of membrane distillation (MD) to transfer only volatiles means that very high quality treated water is obtainable, with TOC levels below 1 ppm and negligible quantity of salts. A unique feature is that the MDBR allows for organic retention times to be much greater than the HRT. The TOC in the permeate was consistently lower than 0.7 mg l1 for all experiments. Stable fluxes in the range 2–5 l m2 h1 have been sustained over extended periods. The MDBR was described to have the potential to achieve in a single step, the reclamation obtained by the combined MBR þ RO process. It was also suggested that for viable operation, it would be necessary to use low-grade (waste) heat and water cooling. Several other emerging approaches are also noticeable in contemporary literature. These include hybrid MBR-CAS concept (De Wilde et al., 2009), anaerobic baffled reactor-MBR combination (Pillay et al., 2008), etc.
4.16.4.5 Technology Benefits: Operators’ Perspective The relative advantages of MBR over the CAS process were outlined in Section 4.16.3.3. This section highlights the technical benefits of MBRs cited by the operators: 1. high-quality effluent, ideal for post membrane treatments (e.g., NF and UF); 2. space savings, enabling upgrading of plants without land expansion; 3. shorter start-up time compared to conventional treatment systems; 4. low operating and maintenance manpower requirement (average of 1.7 working hours per MLD); and 5. (5) automated control.
4.16.4.6 Technology Bottlenecks MBR technology is facing some research and development challenges. The technology bottlenecks as reported in the literature include (Howell, 2002, 2004; Lesjean et al., 2004; Le-Clech et al., 2005a; Yang et al., 2006; Melin et al., 2006) 1. Membrane fouling. Further understanding the mechanisms of membrane fouling and developing more effective and easier methods to control and minimize membrane fouling. 2. Pretreatment. Effective methods to limiting membrane clogging and operational failures.
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3. Membrane life span. Increasing membrane mechanical and chemical stability. 4. Cost. Further reduction of costs for maintenance and replacement of membranes, energy requirement, and labor requirements. 5. Plant capacity. Scaling up for large plants. 6. Exchangeability of modules. Module exchangeability between different brands (reduction of costs for replacement of membranes). Some other problems often encountered by the operators include (Leslie and Chapman, 2003; Adham et al., 2004; LeClech et al., 2005a; Yang et al., 2006)
• • • • • • • • • •
membrane fouling during permeate backpulsing, entrained air impacting suction-pump operation, bioreactor foaming, inefficient aeration due to partial clogging of aerator holes, no significant decrease of biosolid production, scale buildup on membrane and piping, corrosion of concrete, hand rails, and metallic components due to corrosive vapor produced during high temperature NaOCl cleaning, membrane delamination and breakage during cleanings, odor from screening, compaction, drying beds, and storage areas (although normally less than in CAS), and failure of control system.
Although the commercialization of MBRs has expanded substantially in the past 20 years, target markets have not been tapped to a large extent and new potential areas of applications are continually developing. The R&D challenges mentioned above, when tackled, will lead to a more competitive and mature market for MBR applications. Lesjean et al. (2004) contend that academic research is addressing only some of these issues. For instance, while many publications on fouling are being produced and some cost studies are conducted, no significant research efforts have addressed membrane life span, pretreatment, and scale-up issues. Academic researchers can expect interest from MBR companies and plant operators on these subjects, and should direct some of their research programs to address these needs. Among the challenges underscored by the experts, membrane fouling is one of the most serious problems that has retarded faster commercialization of MBR technology. The causes, characteristics, mechanisms of fouling, and methods to prevent or reduce membrane fouling are discussed elaborately in Section 4.16.4.7; Section 4.16.5.5 sheds light on the issue of exchangeability of modules. The remainder of the current section will be devoted to the issues closely related to membrane fouling and performance, that is, mechanical pretreatment and membrane integrity:
•
Pretreatment. Pretreatment is one of the most critical factors for ensuring a stable and continuous MBR operation. Due to membrane sensitivity to the presence of foreign bodies, fine prescreening of the feed (and sometimes of the mixed liquors) must occur. The type of sieve installed is very important with regard to the total screening of hair and fibers. Recent studies (Frechen et al., 2006; Schier et al., 2009) have shown sieves with smaller gap sizes and with
•
two-dimensional gap geometries to perform better. On the other hand, even intensive long-term pilot plant trials can fail to suggest the effective scale-up design of the sieve (Melin et al., 2006). If too many clogging problems occur, the original pre-screen systems are usually upgraded to finer screens. However, when both the influent and the mixed liquor are filtered with a fine prescreen, a large amount of trash is produced (up to 3.8 m3 per week for a 1.4 MLD plant) (Le-Clech et al., 2005a; Melin et al., 2006; Schier et al., 2009). It should be noted that the investment in pretreatment is of little use if the bioreactor is uncovered, in which case, different sorts of debris can easily enter the bioreactor. It is recommended to remove these items using a high-pressure water hose. However, many MBR users report that this type of manual cleaning causes membrane-fiber breakage. In order to keep the membrane effectively separated from the fibrous materials, Schier et al. (2009)proposed the following mechanical-treatment concept: conventional pretreatment including screen and grit chamber/grease trap to be placed before the biological tank, causing braid of hair and fibers formed therein to be removed by the sieve placed before the separate filtration chamber housing the membrane modules. Membrane integrity. A major problem facing MBR systems is the loss of membrane integrity, which leads to the permeate-quality deterioration and ineffective backwashing. When breakage occurs in a submerged hollow-fiber MBR system, continuous filtration may allow solids and particles to quickly clog the broken fiber. However, application of backwash would force the solids out of the fiber. Accordingly, once damaged, disinfection of the product water would be compromised and it would also cause the loss of the backwash efficiency; and the faulty membrane/module would need to be changed quickly.
Faulty installation is one obvious reason for membrane failure. Once under pressure, an incorrectly installed membrane module can be compressed. Other reasons associated with regular operation include frequent and/or extended contact between membrane and cleaning solution causing delamination of the membrane, scoring and cleaving of the membrane resulting from the presence of abrasive or sharpedged materials in the influent, and operating stress and strain occurring in the system due to fiber movement and membrane backwashing. A better understanding of the effect of membrane material, age, and fouling on membrane integrity may be gained from hollow-fiber-tensile test reported in the literature (Childress et al., 2005; Gijsbertsen-Abrahamse et al., 2006). Even flat-sheet membranes used in MBRs are not immune to occasional failure (Cornel and Krause, 2003). The construction of current flat-sheet MBR membrane panels is a labor-intensive, multistep operation. These are typically sandwich constructions with three separate layers. Two of them are pre-fabricated membrane layers, while the third one is a permeate drainage layer which is sandwiched between them. The three layers of the sandwich are held together by gluing or laminating techniques over their entire surface or just at their edges. Flat-sheet membranes have been found to be sensitive to breaking near the top
Membrane Biological Reactors
due to poor adhesion of the membrane to the support layer (Doyen et al., 2010).
4.16.4.7 Membrane Fouling – the Achilles’ Heel of MBR Technology Although MBR has become a reliable alternative to CAS processes and an option of choice for many domestic and industrial applications, membrane fouling and its consequences in terms of plant maintenance and operating costs limit the widespread application of MBRs (Le-Clech et al., 2006). Membrane fouling can be defined as the undesirable deposition and accumulation of microorganisms, colloids, solutes, and cell debris within pores or on membrane surface (Meng et al., 2009). It results from the interaction between the membrane material and the components of the activated sludge liquor, which include biological flocs formed by a large range of living microorganisms along with soluble and colloidal compounds. Thus, it is not surprising that the fouling behavior in MBRs is more complicated than that in most membrane applications. The suspended biomass has no fixed composition and varies with both feedwater composition and MBR operating conditions employed. Accordingly, although many investigations of membrane fouling have been published, the diverse range of operating conditions and feedwater matrices employed, and the limited information reported in most studies on the biomass composition in suspension or on the membrane, have made it difficult to establish any generic behavior pertaining to membrane fouling in MBRs. Three fouling phenomena need to be recognized and duly addressed:
• • •
Cake formation. This results from the balance of forces (shear stress at the membrane wall and filtration force) and is evidently linked to the biomass characteristics. Blockage of bundle of fibers. The bundle of fibers act as a deep bed filter (depending on biomass characteristics and structure of the bundle). Biofilm formation. This is not strictly dependent upon biomass characteristics as, very often, the microorganisms involved in the biofilm formation are not the dominant species in the biomass.
4.16.4.7.1 Fouling development Zhang et al. (2006a) proposed a three-stage history for membrane fouling in MBRs:
• • •
Stage 1. An initial short-term rise in TMP due to conditioning. Stage 2. Long-term rise in TMP, either linear or weakly exponential. Stage 3. A sudden rise in TMP, with a sharp increase in dTMP/dt, also known as the TMP jump.
When operating at fluxes well below the apparent critical flux of the MLSS, a slow steady rise in TMP (stage 2) is observed which eventually changes to a rapid rise in TMP (stage 3). For sustainable operation, the aim would be to limit the extent of stage 1, prolong stage 2, and avoid stage 3, since it could be difficult to restore.
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4.16.4.7.2 Types of membrane fouling Definitions based on ease of removal and a variety of confusing terminologies have been proposed in the literature to describe fouling. For example, based on the ease of removal, some authors prefer to use the term ‘irreversible fouling’ to the fouling that can be removed by chemical cleaning but not by physical cleaning. Recently, Meng et al. (2009) proposed a somewhat changed definition and used the terms ‘removable’ and ‘irremovable’ for the fouling which is easily eliminated and which requires chemical cleaning, respectively. This chapter, however, uses the more direct terms – physically removable fouling and chemically removable fouling. The formation of a cake layer which can be described as a porous media with a complex system of interconnected interparticle voids has been reported as the major contributor to membrane fouling in MBRs (Jeison and van Lier, 2007; Ramesh et al., 2007). Such fouling is usually physically removable. Recently, a large number of scientific investigations have been performed in order to gain a better understanding of cake-layer formation and cake-layer morphology employing techniques such as confocal laser-scanning microscopy (CLSM), multiphoton microscopy, etc. (Yang et al., 2007; Hughes et al., 2006, 2007). During initial filtration, colloids, solutes, and microbial cells pass through and deposit inside the membrane pores. However, during the long-term operation of MBRs, the deposited cells multiply and yield extracellular polymeric substance (EPS), which clog the pores and form a strongly attached fouling layer. Chemical cleaning is usually required to remove such fouling. Evaluation of physically removable and chemically removable fouling propensity of MBR mixed liquor has been the focus of many studies to date (Field et al., 1995; Ognier et al., 2004; Pollice et al., 2005; Bacchin et al., 2006; Guglielmi et al., 2007; Lebegue et al., 2008; Wang et al., 2008b). Some of the definitions are based on the fouling components. The fouling in MBRs can be classified into three major categories: biofouling, organic fouling, and inorganic fouling, although, in general, all of them take place simultaneously during membrane filtration of activated sludge. Biofouling refers to the deposition, growth, and metabolism of bacteria cells or flocs on the membranes. Biofouling may start with the deposition of individual cell or cell cluster on the membrane surface, after which the cells multiply and form a biocake (Liao et al., 2004; Pang et al., 2005; Wang et al., 2005; Ramesh et al., 2007). Techniques such as scanning electron microscopy (SEM), CLSM, atomic force microscopy (AFM), and direct observation through the membrane (DOTM) have been extensively used to derive valuable information regarding floc/cell-deposition process and the microstructure or architecture of the cake layer. Certain studies have also analyzed the microbial community structures and microbial colonization on the membranes in MBRs (Chen et al., 2004; Jin et al., 2006; Jinhua et al., 2006; Zhang et al., 2006b; Miura et al., 2007; Lee et al., 2009) employing molecular techniques. Such studies reported that the microbial communities on membrane surfaces were quite different from those in the suspended biomass and initially a specific phylogenetic group of bacteria may play the key role in development of the mature biofilm. However, a temporal change
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of microbial-community structure can take place due to the development of anoxic conditions in the cake layer. Organic fouling in MBRs refers to the deposition of biopolymers on the membranes (Meng et al., 2009). Due to the small size, the soluble biopolymers can be deposited onto the membranes more readily, but they have lower back-transport velocity in comparison to large particles (e.g., colloids and sludge flocs). Powerful analytical tools such as Fourier transform infrared (FTIR) spectroscopy, solid-state 13C-nuclear magnetic resonance (NMR) spectroscopy, and high-performance size-exclusion chromatography (HP-SEC) are usually utilized for identification of the deposited biopolymers (Kimura et al., 2005; Rosenberger et al., 2006; Zhou et al., 2007; Teychene et al., 2008) and studies have confirmed that SMP or EPS is the origin of organic fouling in MBR. Inorganic elements such as Mg, Al, Fe, Ca, Si, etc. and metals can enhance the formation of biofouling and organic fouling and can together form a recalcitrant cake layer (Lyko et al., 2007; Wang et al., 2008b). Inorganic fouling can form in two ways – due to concentration-polarization-led chemical precipitation and entrapment within biopolymer gel layer (Meng et al., 2009). Chemical cleaning agents such as ethylenediaminetetraacetic acid (EDTA) might efficiently remove inorganics on the membrane surface (Al-Amoudi and Lovitt, 2007); however, the fouling caused by inorganic scaling may not be easy to eliminate even by chemical cleaning (You et al., 2006).
Figure 8 lists the membrane-fouling parameters, while Figure 9 illustrates the interrelations and combined effect of those parameters. Some of the membrane characteristics and the parameters that influence the performance of the MBRs are discussed in the following: 1. Physical parameters.
•
Pore size and distribution. Studies revealed that the pore size alone could not predict hydraulic performances. The effects of pore size (and distribution of pore size) on membrane fouling are strongly related to the feedsolution characteristics and in particular the particlesize distribution. The complex and changing nature of
Membrane fouling
4.16.4.7.3 Parameters influencing MBR fouling All the parameters involved in the design and operation of MBR processes have an influence on membrane fouling (Le-Clech et al., 2006; Meng et al., 2009). While some of these parameters have a direct influence on MBR fouling, many others result in subsequent effects on phenomena exacerbating fouling propensity. However, three main categories of factors can be identified – membrane and module characteristics, feed and biomass parameters, and operating conditions.
Membrane characteristics • Physical parameters -Pore size and distribution -Porosity/roughness -Membrane configuration • Chemical parameters -Hydrophobicity -Materials
Mixed liquor characteristics
Feed
Biomass
Figure 9 Interrelations and combined effect of the membrane fouling parameters.
Feed–biomass characteristics
Operating conditions
• Nature of feed and concentration • Biomass fractionation • Biomass (bulk) parameters -MLSS concentration -Viscosity -Temperature -Dissolved oxygen (DO) • Floc characteristics -Floc size -Hydrophobicity/surface charge
• Aeration, cross-flow velocity • Sludge retention time (SRT) • Unsteady state operation
• Extracellular polymeric substance (EPS) • Soluble microbial products (SMP) Figure 8 Membrane fouling parameters at a glance.
Operating conditions
Membrane characteristics
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•
•
the biological suspension present in MBR systems and the large pore-size distribution of the membrane generally used in MBR systems are the main reasons for the undefined general dependency of the flux propensity on pore size (Chang et al., 2002a; Le-Clech et al., 2003b). It is generally expected that smaller-pore membranes would reject a wider range of materials, and the resulting cake layer would feature a higher resistance compared to large-pore membranes. However, this type of fouling is easily removed during the maintenance cleaning than fouling due to internal pore clogging obtained in larger-pore membrane systems. The chemically removable fouling, due to the deposition of organic and inorganic materials onto and into the membrane pores, is the main cause of the poor longterm performances of larger pore-size membranes (Chang et al., 2001; He et al., 2005). However, the opposite trend is sometimes reported (Gander et al., 2000). The duration of the experiment and other operating parameters such as cross-flow velocity and constant pressure or constant flux operation have a direct influence on the determination of the optimization of the membrane pore size and are responsible for contradictory reports in the literature. Porosity/roughness. Membrane roughness and porosity along with membrane microstructure, material, and pore-size distribution were suggested as potential reasons for the different fouling behaviors observed (Kang et al., 2006; Ho and Zydney, 2006). For instance, a track-etched membrane, with its dense structure and small but uniform cylindrical pores, featured the lowest resistance due to pore fouling in contrast to the other membranes having interwoven sponge-like highly porous network (Fang and Shi, 2005). Other studies have pointed out the importance of pore-aspect ratio (mean major-axis length/mean minor-axis length) (Kim et al., 2004) or roughness (He et al., 2005) on fouling in an MBR. Membrane configuration. In submerged MBR processes, the membrane can be configured as vertical flat plates, vertical or horizontal hollow fine fibers (filtration from out to in) or, more rarely as tubes (filtration from in to out). Each of hollow-fiber and flat-sheet membrane types has specific footprint and air scouring and chemical cleaning requirement, which may favor one process over another for a given application (Judd, 2002; Hai et al., 2005). Nevertheless, hollow-fiber modules are generally more economical to manufacture, provide high specific membrane area, and can tolerate vigorous backwashing (Stephenson et al., 2000). For low-flux operation, hollow fibers are attractive due to their high packing density. A higher fiber-packing density would increase productivity; however, increasing the packing density may lead to severe interstitial blockage due to the impeded propagation of air bubbles toward the core, limiting their effect on fouling limitation (Kiat et al., 1992; Yeo and Fane, 2005; Sridang et al., 2005). However, Hai et al. (2008a) developed a spacer-filled module in order to utilize high packing density without encountering
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severe fouling. Studies have also revealed the effects of other membrane characteristics including hollowfiber orientation, size, and flexibility ( Cui et al., 2003; Ognier et al., 2004; Chang and Fane, 2002; Lipnizki and Field, 2001; Zheng et al., 2003; Zhongwei et al., 2003). 2. Chemical parameters.
•
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Hydrophobicity. The influence of the membrane hydrophobicity on the early stage of the fouling formation may be significant; however, this parameter is expected to play only a minor role during extended filtration periods in MBRs (Le-Clech et al., 2006). Once initially fouled, the membrane’s chemical characteristics would become secondary to those of the sludge materials covering the membrane surface. Nevertheless, because of the hydrophobic interactions occurring between solutes, microbial cells and membrane material, membrane fouling is expected to be more severe with hydrophobic rather than hydrophilic membranes (Madaeni et al., 1999; Chang et al., 1999; Yu et al., 2005a), although different results have also been reported (Fang and Shi, 2005). In many reported studies, change in membrane hydrophobicity often occurs with other membrane modifications such as pore size and morphology, which make the correlation between membrane hydrophobicity and fouling more difficult to assess. Materials. The large majority of the membranes used in MBRs are polymeric based. A direct comparison between polyethylene (PE) and polyvinylidene fluoride (PVDF) membranes clearly indicated that the latter leads to a better prevention of physically irremovable fouling and that PE membrane fouled more quickly (Yamato et al., 2006). Zhang et al. (2008b) studied the affinity between EPS and the three polymeric UF membranes, and observed that the affinity capability of the three membranes was of the order polyacrylonitrile (PAN)oPVDFopolyethersulfone (PES). Although featuring superior chemical, thermal, and hydraulic resistances, ceramic (Fan et al., 1996; Scott et al., 1998; Luonsi et al., 2002; Xu et al., 2003; Judd et al., 2004) and stainless steel (Zhang et al., 2005) membrane modules are not the preferred option for MBR applications due to their high cost (around an order of magnitude more expensive than the polymeric materials).
3. Feed–biomass characteristics.
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Nature of feed and concentration. Fouling in the MBR is mostly affected by the interactions between the membrane and the biological suspension rather than wastewater itself (Choi et al., 2005). Nevertheless, the fouling propensity of the wastewater has to be indirectly taken into consideration during the characterization of the biomass, as the wastewater nature can significantly influence the physicochemical changes in the biological suspensions (Le-Clech, 2003b; Jefferson et al., 2004), which in turn may aggravate fouling.
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Biomass fractionation. The many studies (Bae and Tak, 2005; Li et al., 2005a; Itonga et al., 2004; Lee et al., (2003); Lee et al., 2001a; Wisniewski and Grasmick, 1998; Bouhabila et al., 2001) that are available on the contribution of different fractions of the biomass to fouling usually report contradictory results. Although the relatively low fouling role played by the suspended solids (biofloc and the attached EPS) compared to those of the soluble and colloids (generally defined as soluble microbial products or SMP) is usually reported, the reported relative contribution of the SMP to overall membrane fouling ranges from 17% (Bae and Tak, 2005) to 81% (Itonga et al., 2004). These wide discrepancies may be explained by the different operating conditions and biological states of the suspension used in the reported studies (Figure 10). Although an interesting approach for studying MBR fouling, the fractionation experiments neglect any coupling or synergistic effects which may occur among the different components of the biomass.
•
•
4. Biomass (bulk) parameters.
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MLSS concentration. Although the increase in MLSS concentration has often been reported to have a mostly negative impact on the MBR hydraulic performances (Cicek et al., 1999b; Chang and Kim, 2005), controversies exist (Defrance and Jaffrin, 1999; Hong et al., 2002; Le-Clech et al., 2003b; Lesjean et al., 2005; Brookes et al., 2006). The existence of threshold values above (Lubbecke et al., 1995) or below (Rosenberger et al., 2005) which the MLSS concentration has a negative influence was also reported. Figure 11 depicts the influence of shift in MLSS concentration on flux as reported in different studies. Nowadays, information on additional biomass characteristics (e.g., composition and concentration of EPS) is deemed necessary to furnish a comprehensive picture. On the other hand,
100
Variable: Membrane type
•
Hai et al. (2006a)showed that the extent of fouling was independent of MLSS concentration itself, and was rather more influenced by the efficiency of the foulingprevention strategies adopted. Viscosity. The importance of MLSS viscosity is that it modifies bubble size and can dampen the movement of hollow fibers in submerged bundles (Wicaksana et al., 2006). The net result of this phenomenon would be a greater rate of fouling. Increased viscosity also reduces the efficiency of mass transfer of oxygen and can therefore effect dissolved oxygen (DO) (Germain and Stephenson, 2005); fouling, as discussed later, tends to be worse at low DO. Critical MLSS concentrations have been reported in the literature (Itonga et al., 2004) above which, suspension viscosity tends to increase exponentially with the solid concentration. Temperature. Experiments conducted under moderate temperature usually report greater deposition of materials on the membrane surface at lower temperatures. Temperature may impact membrane filtration by increasing fluid viscosity, causing defloculation of biomass and higher EPS secretion, reducing biodegradation rate, etc. (Jiang et al., 2005; Rosenberger et al., 2006). Dissolved oxygen. The effects of DO on MBR fouling are multiple and may include changes in biofilm structure, SMP levels, and floc-size distribution (Lee et al., 2005). The average level of DO in the bioreactor is controlled by the aeration rate, which not only provides oxygen to the biomass but also tends to limit fouling formation on the membrane surface. Optimum aeration would result in lower specific cake resistance of the fouling layer featuring larger particle sizes and greater porosity (Kang et al., 2003; Kim et al., 2006). Therefore, in general, higher DO tends to lead to better filterability, and lower fouling rate.
Variable: Sludge type
Variable: SRT
80 60 40 20 0 (a)
(b)
Soluble
Colloids
Suspended solids
(c)
Colloid + soluble
Figure 10 Influence of different parameters (membrane type, sludge type, and SRT) on the relative contributions (in %) of the different biomass fractions to MBR fouling. Data from (a) Bae TH and Tak TM (2005) Interpretation of fouling characteristics of ultrafiltration membranes during the filtration of membrane bioreactor mixed liquor. Journal of Membrane Science 264: 151–160; (b) Meng F and Yang F (2007) Fouling mechanisms of deflocculated sludge, normal sludge, and bulking sludge in membrane bioreactor. Journal of Membrane Science 305: 48–56; and (c) Lee W, Kang S, and Shin H (2003) Sludge characteristics and their contribution to microfiltration in submerged membrane bioreactors. Journal of Membrane Science 216: 217–227.
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Limiting or critical or stabilized flux, (l m−2 h−1)
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irregular floc shape, and higher hydrophobicity (Meng et al., 2006).
(1)
100 80
(2)
(7)
60
(3) (4)
40 (5) 20
(6)
0 0
5
10
15
20
25
MLSS concentration, gl−1 Figure 11 Influence of shift in MLSS concentration on flux (fouling) as reported in different studies. Data from (1) Cicek N, Franco JP, Suidan MT, and Urbain V (1998) Using a membrane bioreactor to reclaim wastewater. Journal of American Water Works Association 90: 105–113; (2) Beaubien A, Baty M, Jeannot F, Francoeur E, and Manem J (1996) Design and operation of anaerobic membrane bioreactors: Development of a filtration testing strategy. Journal of Membrane Science 109: 173–184; (3) Madaeni SS, Fane AG, and Wiley D (1999) Factors influencing critical flux in membrane filtration of activated sludge. Journal of Chemical Technology and Biotechnology 74: 539–543; (4) Han SS, Bae TH, Jang GG, and Tak TM (2005) Influence of sludge retention time on membrane fouling and bioactivities in membrane bioreactor system. Process Biochemistry 40: 2393–2400; (5) Bouhabila EH, Ben Aim R, and Buisson H (1998) Microfiltration of activated sludge using submerged membrane with air bubbling (application to wastewater treatment). Desalination 118: 315–322; (6) Bin C, Xiaochang W, and Enrang W (2004) Effects of TMP, MLSS concentration and intermittent membrane permeation on a hybrid submerged MBR fouling. In: Proceedings of the IWA – Water Environment – Membrane Technology (WEMT) Conference. Seoul, Korea, 7–10 June; and (7) Defrance L and Jaffrin MY (1999) Reversibility of fouling formed in activated sludge filtration. Journal of Membrane Science 157: 73–84.
5. Floc characteristics.
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•
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Floc size. The floc-size distribution obtained with the MBR sludge is lower than the results generally obtained from CASP (Zhang et al., 1997; Wisniewski and Grasmick, 1998; Lee et al., 2003; Cabassud et al., 2004; Bae and Tak, 2005). Unlike in the CAS systems, the effective separation of suspended biomass from the treated water is not critically dependent on aggregation of the microorganisms, and the formation of large floc. However, independent of their size, biological floc play a major role in the secretion of EPS and formation of the fouling cake on the membrane surface. Hydrophobicity/surface charge. The direct effect of floc hydrophobicity on MBR fouling is difficult to assess. Conceptually, hydrophobic flocs would lead to high flocculation propensity, less secretion of EPS, and low interaction with the hydrophilic membrane (Jang et al., 2006). However, reports of highly hydrophobic flocs fouling MBR membranes can be found in the literature. For instance, the excess growth of filamentous bacteria, known to be responsible for severe MBR fouling, also resulted in higher EPS levels, lower zeta potential, more
6. Extracellular polymeric substances. The term EPS is used as a general and comprehensive concept for different classes of macromolecules such as polysaccharides, proteins, nucleic acids, (phosphor-)lipids, and other polymeric compounds which have been found at, or outside, the cell surface and in the intercellular space of microbial aggregates (Flemming and Wingender, 2001). EPS are the construction materials for microbial aggregates such as biofilms, flocs, and activated sludge liquors. The functions of EPS matrix are multiple and include aggregation of bacterial cells in flocs and biofilms, formation of a protective barrier around the bacteria, retention of water, and adhesion to surfaces (Laspidou and Rittmann, 2002). With its heterogeneous and changing nature, EPS can form a highly hydrated gel matrix in which microbial cells are embedded (Nielson and Jahn, 1999). Therefore, they can be responsible for the creation of a significant barrier to permeate flow in the membrane processes. Contemporary literature is replete with reports identifying EPS as a major fouling parameter (Chang and Lee, 1998; Cho and Fane, 2002; Nagaoka et al., 1996, 1998; Rosenberger and Kraume, 2002). On the other hand, since the EPS matrix plays a major role in the hydrophobic interactions among microbial cells and thus in the floc formation (Liu and Fang, 2003), it was proposed that a decrease in EPS levels may cause floc deterioration and may be detrimental for the MBR performances. This indicates the existence of an optimum EPS level for which floc structure is maintained without featuring high fouling propensity. Many parameters including gas sparging, substrate composition (Fawehinmi et al., 2004), and loading rate (Cha et al., 2004; Ng et al., 2005) affect EPS characteristics in the MBR, but SRT probably remains the most significant of them (Hernandez et al., 2005). A functional relationship between specific resistance, mixed liquor volatile suspended solids (MLVSS), TMP, and permeate viscosity, and EPS is believed to exist (Cho et al., 2005). 7. Soluble microbial products. SMPs are defined as soluble cellular components that are released during cell lysis, diffuse through the cell membrane, and are lost during synthesis or are excreted for some purpose (Laspidou and Rittmann, 2002; Li et al., 2005a). During filtration, SMPs adsorb on the membrane surface, block membrane pores, and/or form a gel structure on the membrane surface where they provide a possible nutrient source for biofilm formation and a hydraulic resistance to permeate flow (Rosenberger et al., 2005). Since direct relationships between the carbohydrate level in SMP (SMPc) solution with fouling rate (Lesjean et al., 2005), filtration index and capillary suction time (CST) (Greiler et al., 2005; Evenblij et al., 2005b; Tarnacki et al., 2005), critical flux tests (Le-Clech et al., 2005b), and specific flux (Rosenberger et al., 2005) have been clearly described, it reveals SMPc to be the major foulant indicator in MBR systems. However, controversy over the relative contribution of carbohydrate and protein portions of SMP to fouling exists (Evenblij and Van der Graaf, 2004; Drews et al., 2005a; Drews et al., 2006).
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The operating conditions of MBrs are discussed as follows:
•
•
•
Aeration, cross-flow velocity. Since the origin of the SMBR, bubbling has been defined as the strategy of choice to induce flow circulation and shear stress on the membrane surface. Aeration used in MBR systems has three major roles: providing oxygen to the biomass, maintaining the activated sludge in suspension, and mitigating fouling by constant scouring of the membrane surface (Dufresne et al., 1997). However, an optimum aeration rate, beyond which a further increase has no significant effect on fouling suppression, has been observed on many occasions (Ueda et al., 1997; Le-Clech et al., 2003a, 2003b; Liu et al., 2003; Psoch and Schiewer, 2005b). It is also important to note that too intense an aeration rate may damage the floc structure reducing their size, and release EPS into the bioreactor (Park et al., 2005; Ji and Zhou, 2006), and thereby aggravate fouling. Solid retention time. SRT (and thereby the F/M ratio), which greatly controls biomass characteristics, is regarded as the most important operating parameter influencing fouling propensity in MBRs. Considering the advantages of this process over the conventional activated sludge process (CASP), the early MBRs were typically run at very long SRTs to minimize excess sludge (Liu et al., 2005; Gao et al., 2004; Nuengjamnong et al., 2005). But unlike in bench-scale studies employing simpler synthetic feed, the progressive accumulation of nonbiodegradable materials (such as hair and lint) in an MBR fed with real sewage definitely leads to clogging of the membrane module (Le-Clech et al., 2005b). Operating an MBR at higher SRT leads inevitably to increase of MLSS concentration (Zhang et al., 2006c). The increase in aeration intensity to retain high MLSS levels in suspension and maintain proper oxygenation may not be a sustainable option for the treatment process. In this scenario, the increased shear provided to control fouling could cause biofloc deterioration as well as cell lysis and enhanced EPS secretion, and lead to fatal fouling. On the other hand, at infinite SRT, most of the substrate is consumed to ensure the maintenance needs and the synthesis of storage products. The very low apparent net biomass generation observed can explain the low fouling propensity observed for high SRT operation in certain studies (Orantes et al., 2004). It is likely that there is an optimal SRT, between the high fouling tendency of very low SRT operation and the high viscosity suspension prevalent for very long SRT. Unsteady state operation. In practical applications, unsteady state conditions such as variations in operating conditions (flow input/HRT and organic load) and shifts in oxygen supply could occur regularly (Drews et al., 2005a). The start-up phase can also be considered as unsteady operation and data collected before biomass stabilization (including the period necessary to reach acclimatization) may become relevant in the design of MBRs (Cho et al., 2005). Such unsteady state conditions have also been defined as additional factors leading to changes in MBR fouling propensity. For instance, the addition of a spike of acetate in the feedwater significantly decreased the filterability of the biomass in an MBR due to the rise in SMP levels resulting from the feed spike (Evenblij et al., 2005a).
4.16.4.7.4 Fouling mitigation The complex interactions between the fouling parameters complicate the perception of MBR fouling and it is therefore crucial to have a complete understanding of the biological, chemical, and physical phenomena occurring in MBRs to assess fouling propensity and mechanisms and thereby formulate mitigation strategies. As membrane fouling increases with increasing flux in all membrane separation processes, the operating flux should be lower than the critical flux. When the operating flux is below the critical flux, particle accumulation in the region of membranes can be effectively prevented. However, due to physicochemical solute–membrane material interactions, the membrane permeability decreases over time, even when MBRs are operated in subcritical (below critical flux) conditions. Other preventative methods need to be considered to maintain stable operation of MBR systems (Figure 12). Fouling can be removed by various methods and they are as discussed herein: 1. Physical cleaning. The following methods are usually used in combination to remove membrane fouling:
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Permeate backwashing. Membrane backwashing or backflushing refers to pumping permeate in the reverse direction through the membrane. Backwashing has been found to successfully remove most of the reversible fouling due to pore blocking, transport it back into the bioreactor, and partially dislodge loosely attached sludge cake from the membrane surface (Bouhabila et al., 2001; Psoch and Schiewer, 2005a; Psoch and Schiewer, 2006). Frequency, duration, the ratio between those two parameters, and its intensity are the key parameters in the design of backwashing and different combinations of these parameters have proved to be more efficient in different studies (Jiang et al., 2005; Schoeberl et al., 2005). Between 5% and 30% of the produced permeate is used for backwashing. This also
Removal of fouling
Limitation of fouling
• Physical cleaning --Backwashing --Air backwashing --Intermittent operation --Sonification and other energy-intensive processes
• Optimization of membrane characteristics
• Chemical cleaning --Maintenance cleaning --Intensive cleaning
• Optimization of operating conditions --Aeration --Other operating conditions --Membrane module design • Modification of biomass characteristics -Aerobic granular sludge -Coagulant/flocculent -Adsorbent/flux enhancers
Figure 12 Reported membrane fouling mitigation strategies at a glance.
Membrane Biological Reactors
•
•
•
affects operating costs as, obviously, energy is required to achieve a pressure suitable for flow reversion. Certain studies are, therefore, devoted to optimization of backwashing (Smith et al., 2005). Air backwashing. Air, instead of permeate, can also be used as the backflushing medium (Visvanathan et al., 1997; Sun et al., 2004). The invention of air backwashing techniques for membrane declogging led to the development of using the membrane itself as both clarifier and air diffuser. In this approach, two sets of membrane modules are submerged in the aeration tank. While the permeate is extracted through one of the sets, the other is supplied with compressed air for backwashing. The cycle is repeated alternatively, and there is a continuous airflow into the aeration tank, which is sufficient to aerate the mixed liquor. However, air backwashing may also present potential issues of membrane breakage and rewetting (Le-Clech et al., 2006). Intermittent operation. Intermittent operation or membrane relaxation can significantly improve membrane productivity (Yamamoto et al., 1989). During relaxation, back transport of foulants is naturally enhanced as loosely attached foulants can diffuse away from the membrane surface (Ng et al., 2005). Although some studies found it more important than backwashing for fouling removal (Schoeberl et al., 2005), recent studies tend to combine intermittent operation with frequent backwashing for optimum results (Zhang et al., 2005; Vallero et al., 2005). The economic feasibility of intermittent operation for large-scale MBRs has been the focus of certain studies (Hong et al., 2002); however, it seems rather an established operation mode nowadays. Sonification and other energy-intensive processes. Although sonification would be difficult to apply at a large scale due to the focused nature of the sonic energy, laboratory-scale studies have explored sonification for breaking down cake layers in MBRs, especially in case of ceramic membranes. Certain studies have confirmed the efficiency of application of sonification alone or in combination with backwashing for removing the cake layer (Lim and Bai, 2003; Fang and Shi, 2005). However, other studies report that fouling may even worsen due to pore blocking (Hai et al., 2006a). Attempts have also been made to control fouling or modify sludge by using ozone and electric field (Chen et al., 2007; Huang and Wu, 2008; Sui et al., 2008; Wen et al., 2008).
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Maintenance cleaning with moderate chemical concentration (weekly) is applied to maintain design permeability and it helps to reduce the frequency of intense cleaning. This may be replaced by a more frequent
(e.g., on a daily basis) chemically enhanced backwash utilizing mild chemical concentration. Intensive (or recovery) chemical cleaning (once or twice a year) is generally carried out when further filtration is no longer sustainable because of an elevated TMP.
The MBR suppliers propose their own chemical cleaning recipes, which differ mainly in terms of concentration and methods, and often site-specific protocols are followed (Kox, 2004; Tao et al., 2005; Le-Clech et al., 2005b). Mainly, sodium hypochlorite (for organic foulants) and citric acid (for inorganics) are used as chemical agents. Some pitfalls of chemical cleaning are worth noting. The detrimental effect of cleaning chemicals on biological performance has been reported (Lim et al., 2005; Hai et al., 2007). It has also been mentioned that the level of pollutants (measured as TOC) in the permeate rises just after the chemical cleaning step (Tao et al., 2005). This raises concern especially in case of MBRs used in the reclamation process trains (i.e., e.g., upstream of RO) (Le-Clech et al., 2006). Chemical cleaning may also shorten the membrane lifetime and disposal of spent chemical agents causes environmental problems (Yamamura et al., 2007). The measures to limit fouling are discussed next. Recently, there have been a significant number of studies which focused on the ways to limit fouling. The proposed strategies include (1) improving the antifouling properties of the membrane, (2) operating the MBR under specific nonor-little-fouling conditions, and/or (3) pretreating the biomass suspension to limit its fouling propensity. They are discussed as follows: 1. Membrane modification.
•
2. Chemical cleaning. The effectiveness of physical cleaning tends to decrease with operation time as more recalcitrant fouling accumulates on the membrane surface. Therefore, in addition to physical cleaning, different types/intensities of chemical cleaning are applied in practice. A combination of the following types of cleaning is usually applied (Le-Clech et al., 2006):
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Optimization of membrane characteristics. Many studies have shown that chemical modifications of the membrane surface can efficiently improve antifouling properties. Recent examples comprise (1) increasing membrane hydrophilicity by NH3 and CO2 plasma treatments (Yu et al., 2005a, 2005b) and ultraviolet (UV) irradiation (Yu et al., 2007), (2) TiO2 entrapped membrane (Bae and Tak, 2005), and (3) applying precoating of TiO2 (Bae et al., 2006), GAC (Hai, 2007), ferric hydroxide (Zhang et al., 2004), polyvinylidene fluoride-graft-polyoxyethylene methacrylated (PVDF-gPOEM) (Asatekin et al., 2006), polyvinyl alcohol (PVA) (Zhang et al., 2008a), etc. Improved performance in case of precoated membrane has been attributed to the adsorption of soluble organics on the precoat, limiting the direct contact between the organics and the membrane. Self-forming dynamic membrane-coupled bioreactors, utilizing coarse pore-sized substrates and allowing cake and gel layers to deposit on the surface, have been reported to obtain high flux and good removal in certain studies, although stable performance cannot be expected with such a filtration barrier (Wu et al., 2004). Membrane module design. The membrane module design by optimizing the packing density of hollow fibers or flat sheets, the location of aerators, the orientation of
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fibers, and diameters of fibers (Chang and Fane, 2001; Chang et al., 2002b; Fane et al., 2002) remains another important parameter in the optimization of the MBR operation. In a specially designed module in which air bubbles were confined in close proximity to the hollow fiber (rather than diffusing in the reactor), higher permeability was obtained (Ghosh, 2006). Two major design approaches are adopted in case of the commercially available hollow-fiber bundles. One of these approaches relies on partitioning of bundles of fibers, which are fixed at both ends, to secure flow path of air bubbles introduced from the center of the bundle at the base, thereby leading sludge out of the module. In another approach, bundle of one-end free fibers are allowed to float freely under the scouring action of air bubbles introduced from the core of the bundle to avoid accumulation of sludge. In order to utilize high packing density without encountering severe fouling, a new approach to hollow-fiber module design was explored by Hai et al. (2008a). Spacer was introduced within usual hollow-fiber bundles with the aim of minimizing the intrusion of sludge into the module. The little amount of intruded sludge was then backwashed through the bottom end, while the sludge deposited on the surface was effectively cleaned by air scouring. In this way, efficient utilization of cleaning solution and air for backwashing and surface cleaning, respectively, were possible. Recent approaches such as novel fiber sheet (FiSh) membrane (Heijnen et al., 2009), multimodule flat-sheet concept (Kreckel et al., 2009), and vacuum rotation membrane (Alnaizy and Sarin, 2009; Komesli et al., 2007) are also noticeable.
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3. Modification of biomass characteristics.
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•
2. Optimization of operating conditions.
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Aeration. As mentioned earlier, bubbling is an established strategy to induce flow circulation and shear stress on the membrane surface. The aeration intensity (air/permeate ratio, m3/m3) applied by MBR suppliers may vary between 24 and 50, depending on the membrane configuration (flat sheet vs. hollow fiber) and the MBR tank design (whether the membrane and aerobic zone combined into a single tank or not) (Tao et al., 2005; Le-Clech et al., 2006). However, recent large-scale studies revealed these original ratios to be quite conservative (Tao et al., 2005). The specific design of bubble size, airflow rate and patterns, and location of aerators have been defined as crucial parameters in fouling mitigation. As the energy involved in providing aeration to the membrane remains a significant cost factor in MBR design, efforts have been focused on optimization of aeration both from the points of view of fouling mitigation and reducing energy requirement. Recent developments in aeration design include cyclic aeration systems (Rabie et al., 2003), intermittent aeration (Yeom et al., 1999; Nagaoka and Nemoto, 2005), air pulsing (Judd et al., 2006), air sparging (Ghosh, 2006), improved aerator systems (Miyashita et al., 2000; Cote, 2002; Hai et al., 2008), etc.
Other operating conditions. The overall performance of the MBR is closely related to the choice of SRT value. Further optimizations of operating conditions through reactor design have been studied and include the addition of a spiral flocculator (Guo et al., 2004), vibrating membranes (Genkin et al., 2005), helical baffles (Ghaffour et al., 2004), suction mode (Kim et al., 2004) and high-performance compact reactor (Yeon et al., 2005), novel types of air lift (Chang and Judd, 2002), porous and flexible suspended membrane carriers (Yang et al., 2006), and the sequencing batch MBR (Zhang et al., 2006d). A reasonable flux rate without significant fouling is ideally expected. The concept of sustainable flux in MBRs was introduced from this point of view (Ng et al., 2005).
•
Aerobic granular sludge. In order to obtain higher biological aggregates in the bioreactor, aerobic granular sludge has also been used in MBR systems (Li et al., 2005b). With an average size around 1 mm, granular sludge increased the membrane permeability by 50%, but lower cleaning recoveries were observed (88% of those obtained with a conventional MBR). Such granular sludge may also not be stable under long-term operation (Hai, 2007). Coagulant/flocculant. Due to back transport and shearinduced fouling control mechanisms, large microbial flocs are expected to have a lower impact on membrane fouling. Based on this expectation, studies have explored addition of coagulants such as alum (Holbrook et al., 2004), ferric chloride, zeolite (Lee et al., 2001b), chitosan (Ji et al., 2008), etc. and have shown permeability enhancement. Pretreatment of the effluent is also possible and studies based on the pre-coagulation/ sedimentation of effluent before its introduction in the bioreactor revealed the fouling limitation offered by this technique (Itonga and Watanabe, 2004; Le-Clech et al., 2006). Adsorbent/flux enhancers. Lower fouling propensity is observed in MBR processes when biomass is mixed with adsorbents in that addition of adsorbents into biological treatment systems decreases the level of pollutants, and more particularly organic compounds (Kim and Lee, 2003; Lesage et al., 2005; Li et al., 2005c; Ng et al., 2006). In view of saturation of PAC during longterm studies, researchers have suggested periodic addition of PAC (Ng et al., 2005; Fang et al., 2006). Certain studies have proposed pre-flocculation and PAC addition (Guo et al., 2004; Cao et al., 2005).
A cationic polymer-based membrane performance enhancer (MPE 50) has been commercialized by Nalco recently. The interaction between the polymer and the soluble organics was reported as the main mechanism responsible for performance enhancement (Yoon et al., 2005). The potential impacts of coagulants or adsorbents on biomass community or biomass metabolism need to be taken into account (Iversen et al., 2009), and the discharge of some chemicals that are used as coagulants or adsorbents might be a potential environmental
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risk. Such flux enhancers are probably best suited for solving occasional upsets rather than their continuous addition. Emerging fouling monitoring/control techniques such as interference of microbial intercellular communication by enzymatic degradation of signal molecules (Kjelleberg et al., 2008; Yeon et al., 2009), proteins and polysaccharides sensor for online fouling control (Mehrez et al., 2007), application of two-dimensional fluorescence for monitoring MBR performance (Galinha et al., 2009), etc., are worth noting.
4.16.5 Worldwide Commercial Application 4.16.5.1 Installations Worldwide The MBR process is an emerging advanced wastewater-treatment technology that has been successfully applied at an everincreasing number of locations around the world. MBRs were first developed 40 years ago and have been used commercially in Japan for almost 30 years. Since 1990, MBR technology has been adopted in North America and Europe, and it is now experiencing rapid growth in a wide variety of applications. In Asia, the drive in Japan was followed by an enthusiastic uptake in South Korea in the 1990s, and more recently by China. The highest growth rates are found in areas of greatest water stress for reuse applications, such as the southwestern US, China, Singapore, and Australia. The low footprint of the MBR is a significant driver for developed economies.
4.16.5.1.1 Location-specific drivers for MBR applications Howell (2004) stipulated the location-specific global drivers for MBR technology as follows: 1. Asia. MBR technology is being considered at many locations all over Asia, the main driver being water reclamation. Examples of settings vary from small-scale applications in Japan, where MBR product water is reused as toilet-flushing water in apartment blocks, medium-sized industrial applications in various countries, and large-scale municipal WWTPs in China. 2. Middle East. Clean-water shortages are the obvious driver for MBR applications in the Middle East, in treatment of both municipal as well as industrial (petrochemical) wastewater. 3. Europe. In Western Europe, water reclamation is not the main driver. In the UK, an important driver is compactness and strict discharge limits due to bathing wastewater requirements. In Germany and the Netherlands, important push factors are strict discharge requirements due to ecologically sensitive surface waters and the innovative character of the technological developments related to MBR. In Southern Europe, water reclamation can be considered as the main driver. 4. Northern America. In the US and Canada, MBR initiatives are predominantly driven by strict discharge requirements due to ecologically sensitive surface waters. At some locations, water reclamation is another important driver. In the US, where wastewater-treatment infrastructure lags behind population growth, MBRs are being increasingly implemented to make up the shortfall. Where there is
597
limited space to locate treatment plants, MBRs offer the potential to meet the needs of communities. 5. Australia. Stringent effluent-quality targets and water-reuse potential are obvious drivers for drought-stricken Australia.
4.16.5.1.2 Plant size Earlier MBR technology was favored in difficult applications or those applications where compactness was important and reuse was the target; and it usually involved smaller plants. As the demand for MBR technology grows globally, both the number of installations and the capacity of the installed plants are increasing dramatically. The most optimistic industry estimates suggest that up to 1000 new MBR plants will be built annually during the survey period. The size of the constructed plants has grown from facilities treating hundreds to thousands of gallons of wastewater per day to those treating tens of millions of gallons per day in just a few years. However, the most common capacity for current worldwide MBR installations ranges from the 50 000 gpd (200 m3 d1) to 500 000 gpd systems. The largest MBR plant in the world is set to be operational in 2010/11 in King County, Washington State. When completed, the facility will have an initial peak flow capacity of 495 000 m3 d1 (average 136 000 m3 d1), rising to a daily 645 000 m3 (average 205 000 m3) by 2040.
4.16.5.1.3 Development trend and the current status in different regions Figure 13 shows the regional share of total MBR plants as of 2003. Next, we discusss the trend of MBR growth in the three continents, Asia, Europe, and North America. 1. Asia. In the 1970s sidestream technology first entered the Japanese market. By 1993, 39 of such facilities had been reported for use in sanitary and industrial applications (Aya, 1994). The application of MBR in Japan concerned Europe 11%
N. America 16%
Asia 73% Figure 13 Regional share of total MBR plants (2003). Data from Pearce G (2008a) Introduction to membranes: An introduction to membrane.
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small-scale installations for domestic wastewater treatment and reuse and some industrial applications, mainly in the food and beverage industries where highly concentrated flows are common. The domestic application often consists of so-called Johkaso or septic-tank treatment and inbuilding (office or domestic) wastewater-collection systems. In the early 1990s, the Japanese Government launched an ambitious 6-year research and development (R&D) project which led to a major technological and industrial breakthrough of the MBR process: the conception of submerged membrane modules, working with low negative pressure (out-to-in permeate suction), and membrane aeration to reduce fouling. This paved the way toward a significant reduction of capital and operation costs, due to the reduction and simplification of equipment and the abandonment of the energy-demanding sludge-recirculation loop. Since then, commercial MBRs proliferated in Japan, which had 66% of the world’s processes in 2000 (Stephenson et al., 2000). In Japan, although MBRs have long been used for industrial wastewater treatment or for reuse of wastewater in large buildings and so on, the introduction of municipal MBRs has lagged behind compared with other water-related fields. The first MBR for municipal wastewater treatment with an installed capacity of 2100 m3 d1 (total design capacity 12 500 m3 d1) in Japan started operation in March 2005, and this accelerated the introduction of MBRs in Japanese sewerage systems. Nine MBR plants, mostly small scale, for municipal wastewater treatment, are in operation at present (Table 8). In addition, there are several MBR plants currently in the design or planning stage. The number of MBRs for municipal wastewater is expected to increase in the near future and the technology will also play an important role in retrofitting and upgrading of existing treatment plants. The MBR technology saw an enthusiastic uptake in South Korea in the 1990s following its introduction in Japan. By 2005, the number of MBR plants rose up to more than 1300 (Namkung, 2008). The plants are mostly small, with more than 60% of the total plants having a capacity of less than 50 m3 d1. The plants were predominantly built on the submerged membrane technology (hollow fiber, 79%; flat sheet, 12%), while a meager 9% facilities utilized the tubular membranes in sidestream format. China has recently emerged as a strong MBR market. Hence, it would be interesting to cast light on the specific
Table 8
mode of development in that country. While the first paper on MBR was published in 1991, the emergence of a number of local and overseas companies developing MBR market in China accelerated with the funding of R&D projects by the Ministry of Science and Technology (MOST) in 1996 (Wang et al., 2008a). Since then, much progress has been achieved both in research and in practical applications of MBR in China. This is evident by the recent yearly publication rate of 35–40 English articles on MBR in China and the construction of a total of 254 plants for municipal (137) and industrial (117) wastewater treatment by 2008. The Chinese MBR market has the presence of a total of 33 companies or institutes, including famous overseas companies such as GE–Zenon Environmental Inc., Mitsubishi–Rayon (Japan), Toray (Japan), NOVO Environmental Technology (Singapore), and XFlow (Netherlands). Among these, only three companies provide flat-sheet MBR, and, interestingly, the worldwide renowned flat-sheet membrane provider, Kubota (Japan), was not found to be very active in the Chinese membrane market. Most of the plants in operation are medium scale or small scale in terms of treatment capacity, the number of plants with treatment capacity below 1000 m3 d1 totaling 225. The largest MBR plant with a capacity of 80 000 m3 d1 for municipal wastewater treatment and reuse is located in Beijing. Several other large MBR plants are also in the planning stage. Wang et al. (2008a) contend that the increasingly stringent discharge standards and the great need of water reclamation and reuse will further push forward the application of everlarger municipal MBR plants in China, especially in North China which has severe water shortage. 2. Europe. A market survey of the European MBR industry was performed by Lesjean and Huisjes (2008). They identified MBR plants constructed up to 2005, and about 300 references of industrial applications (420 m3 d1) and about 100 municipal WWTPs 4500 p.e. were listed. In Europe, the first full-scale MBR plant for treatment of municipal wastewater was constructed in Porlock (UK, commissioned in 1998, 3800 p.e.), soon followed by WWTPs in Bu¨chel and Ro¨dingen (Germany, 1999, 1000 and 3000 p.e., respectively), and in Perthes-en-Gaˆtinais (France, 1999, 4500 p.e.). In 2004, the largest MBR plant worldwide so far was commissioned to serve a population of 80 000 p.e. (in Kaarst, Germany). The installations thus grew from small WWTPs to very large WWTPs within a few
Municipal MBR plants in Japan
Name of plant
Total design capacity (m3d1)
Capacity at commissioning (m3d1)
Membrane format
Start of operation
Fukusaki Kobuhara Yusuhara Okutsu Daito Kaietsu Zyosai Heta Ooda
12 500 240 720 580 2000 230 1375 3200 8600
2100 240 360 580 1000 230 1375 2140 1075
Flat sheet Flat sheet Flat sheet Hollow fiber Flat sheet Hollow fiber – – –
2005 2005 2005 2006 2006 2007 2008 2008 2009
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years. Nevertheless, the favored range for MBR systems still appears to be only 100–500 m3 d1 and 1000–20 000 p.e. for industrial and municipal wastewaters, respectively. The design capacity of the industrial units is more than an order of magnitude smaller than for the municipal WWTPs. Lesjean and Huisjes (2008) opined that, although the construction of very large MBR plants (4100 000 p.e.) were recently announced with much publicity, this will remain the exception in Europe, because of the lower lifecycle costs (Lesjean et al., 2004) of WWTP plants equipped with tertiary-membrane filtration (Figure 14). Although not representative of the market, the very large plants will attract much attention and thereby may contribute to the market expansion. The industrial market was the pioneer in the early 1990s, whereas the municipal market took off only in 1999. In 2002, 154 MBR units could be counted, among which 85% were for industrial applications. However, taking the installed membrane surface as an indicator of market share, for the period 2003–05, the municipal sector represented 75% of the market volume. Both municipal and industrial sectors saw a sharp increase in the following years, due to the commercial success and much lower capital and operating costs. By 2005, the market growth rate was linear with at least 50 industrial units and 20 municipal plants constructed per year. This progression rate is expected to sustain in the next years or may even further accelerate owing to the evolution and implementation of European and national regulations (Lesjean et al., 2006). The survey by Lesjean and Huisjes (2008) also demonstrated the predominance of the suppliers Kubota (Japan) and GE–Zenon. Their technologies based on submerged filtration modules have been outstandingly successful since 2002. In recent years, the European market can therefore be seen as a quasi-duopoly of two nonEuropean suppliers. In contrast, the most successful MBR technologies in the 1990s, based on sidestream configurations supplied by Wehrle, Norit X-Flow, Berghof, Rodia Orelis, etc., did not experience any significant market
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growth over the last 3 years. This could explain the recent move of companies such as Wehrle and Norit to develop and commercialize novel low-energy airlift MBR systems. They argued that the industrial market has become mature: the MBR is considered as the best available technology by many industries. On the other hand, the municipal market is expected to witness further growth over the next decade under the combined effects of the acceleration of plant construction and the capacity increase. 3. North America. Full-scale commercial applications of MBR technology in North America for treatment of industrial wastewaters dated back to 1991 (Sutton, 2003). In the early 1990s, MBR installations were mostly constructed in external configuration. After the mid-1990s, with the development of SMBR system, MBR applications in municipal wastewater extended widely. In the past 15 years, MBR technology has been of increased interest both for municipal and industrial wastewater treatment in North America. The hesitancy on the part of North American municipalities to consider alternative treatment systems to the well-established conventional treatment options delayed the introduction of MBRs into the municipal arena. Industrial applications, particularly for high-strength, difficult-to-treat waste streams, on the other hand, allowed for the considerations of alternative technologies, such as MBRs (Yang et al., 2006). Nevertheless, currently, commercial application in treating industrial wastewaters does not constitute a high percentage of total full-scale MBR plants. Zenon occupies the majority of the MBR market in North America. In 2006, the North American installations constituted about 11% of worldwide installations. As in other places, in North America too, although plant capacities of MBR systems for municipal wastewater treatment are becoming larger, most of the plants in operation are medium scale or small scale in terms of capacity. The largest capacity MBR plant in operation is in Traverse City, MI at 26 900 m3 d1, and the two largest capacity plants under construction are in Johns Creek, GA at 60 000 m3 d1 and King County, Washington State at 136 000 m3 d1.
Capacity, p.e × 104
8
4.16.5.1.4 Decentralized MBR plants: Where and why? 6
4
2
0 1996
1998
2000
2002
2004
2006
2008
Year of commissioning Figure 14 Plot of capacity of randomly selected European MBR plants showing predominance of medium size plants (similar trend prevails worldwide). Data from Schier W, Frechen FB, and Fischer S (2009) Efficiency of mechanical pre-treatment on European MBR plants. Desalination 236: 85–93.
MBR technology can also provide decentralized small-scale wastewater treatment for remote or isolated communities, campsites, tourist hotels, or industries not connected to municipal treatment plants. In small communities, houses are spread out, the population density is low, and hence the use of an on-site system for an individual home or for a cluster of homes could be a cost-effective option. For emerging nations with vast unsewered areas, the population has practically no access to water sanitation, whereby wastewater is directly discharged into water bodies or reused for irrigation without treatment, thus spreading waterborne diseases and causing eutrophication and pollution of water resources. MBR technology could provide a decentralized, robust, and cost-effective treatment for achieving high-quality effluent in such instances. MBRs also offer excellent retrofit capability for expanding or upgrading existing conventional WWTPs.
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Even when appropriate infrastructure for large-scale water recycling facility exists, the decentralized option may be preferable in some cases. This is because the cost of large-scale water-recycling applications remains high and often uneconomical due to the need to overhaul the existing waterdistribution systems. Large-scale water-recycling applications are, hence, currently somewhat restricted. Furthermore, there is a significant risk of cross-connection associated with the dual-reticulation network, which can seriously dampen public support. While the implementation of the large-scale water recycling is expected to take many years, decentralized water recycling can be applied much more readily. It is expected that MBRs can contribute to a significant increase in decentralized water reclamation and reuse activities. The discussion now centers on the limitations of traditional onsite treatment systems. A gradual but permanent reduction in per-capita water use through socially acceptable means is widely recognized by all stakeholders in the water industry as the strategic longterm sustainable solution to address the ongoing water shortage currently experienced by many countries (Tadkaew et al., 2007). Decentralized wastewater management is not a new concept. Tchobanoglous et al. (2003) defined it as the collection, treatment, and disposal/reuse of wastewater from individual dwellings, clusters of homes or isolated communities, industries, or institution facilities. Traditional decentralized treatment systems such as septic tanks were, in the past, widely used to treat small quantities of wastewater. Due to the likely toughening of environmental legislation in the near future, many of the currently operating wastewater treatments will no longer be acceptable and there will be a need to increase their efficiency significantly. Stricter regulations are found for especially sensitive areas, drinking-waterabstraction areas, and bathing waters. The problem of meeting existing and forecasted more-stringent new regulations affects especially small communities, hotels, and campsites in relatively remote areas without access to sophisticated WWTPs. A major obstacle of decentralized water recycling remains the lack of a suitable technology that can meet the strict and unique effluent criteria required for small-scale water treatment. Some essential requirements are high and reliable treated effluent quality, robustness, tolerance to variable contaminant loading, small footprint, and ease of operation and maintenance. We now discuss the advantages of MBRs in decentralized treatment. As discussed in Section 4.16.5.1.2, historically, the largest number of MBR applications was for a capacity of less than 100 m3 d1. This suggests that the application of MBRs for on-site decentralized system is possible and can offer the most advanced wastewater-treatment options in low-density areas at a cost lower than that of conventional large-scale pipeand-plant systems. Jefferson et al. (2000) argued that smallscale WWTPs constitute a potential growth market for the next millennium and urban sustainability through domestic water recycling is a major identified source for this development. Key advantages of MBRs for decentralized wastewater treatment and reuse are:
•
High and reliable treated effluent quality, small footprint, and high tolerance to variable contaminant loading.
•
•
Due to the robustness and modular nature of MBRs, smallscale MBRs can retain the superiority over conventional treatment methods such as septic tanks with regard to effluent quality, which has been very well documented in the literature (Fane and Fane, 2005). MBRs can be easily combined with other complementary treatment technologies such as UV disinfection and prescreening, which can further enhance the robustness of the treatment system and hence make it particularly suitable for water-recycling applications.
The MBRs for decentralized treatment are not without limitations. Besides the obstacles against widespread application of MBR, in general, the high capital cost can be seen as the key limitation of small-scale MBRs although currently there is very little information to substantiate this premise. Friedler and Hadari (2006) analyzed the economic feasibility of on-site graywater-reuse systems in buildings based on MBR systems. They found that on-site MBR systems became feasible when they were used for the treatment of wastewater incorporating several buildings together because cost was sensitive to building size. Therefore, the on-site MBR system for single building might be unfeasible. This could be a limitation of decentralized MBR systems. However, the true cost of water supply, which takes into account the externalities of resource depletion, was not used in their analysis. It is expected that as the demand for decentralized MBRs increases and membrane technology continues to develop, the use of on-site MBRs even for individual dwellings can be cost competitive in the near future. Some of the examples of worldwide decentralized MBRs are discussed next. The successful introduction of MBR systems into small-scale and decentralized applications has led to the development of packaged treatment solutions from most of the main technology suppliers. Sports stadia, shopping complexes, and office blocks are becoming typical end users, especially in areas of water stress (Stephenson et al., 2000; Melin et al., 2006; Tadkaew et al., 2007). The application of MBRs in Japan to date has predominantly concerned small-scale installations for domestic wastewater treatment. One of the earliest reported case studies is on graywater recycling facilities in the Mori building, Tokyo (Stephenson et al., 2000). The plant consists of a sidestream Pleiade MBR (Ubis) to treat the building flow of 500 m3 d1. The selection of an MBR over a traditional treatment process saved an area equivalent to 25 car-parking places. The treated graywater contained less than 5.5 mg l1 BOD and belowdetection level of suspended solids, colon bacilli, and n-hexane extract, enabling reuse of the graywater. Today, the main Japanese MBR providers such as Kubota or Mitsubishi Rayon offer commercial MBR package plants for on-site domestic water treatment. In Australia, small-scale MBR systems for graywater recycling at a single household level have been marketed by several companies such as AquaCell in New South Wales and BushWater in Queensland (Tadkaew et al., 2007). Commercially available systems in Europe include the package treatment plant Clereflo MBR (Conder Products, UK), designed to service populations up to 5000 and the ZeeMOD (Zenon Environmental Inc.) which is available for flow rates
Membrane Biological Reactors of up to 7500 m3 d1. Most of the manufacturers offer similar systems which means that effluent qualities of 5:5:5 (mg l1) (BOD: NH4-N:SS) are now routinely available to end users as standard treatment options (Melin et al., 2006). Households/ community units (4–50 p.e.) is a concept pioneered by Busse (Germany) in 2000 (Lesjean and Huisjes, 2008). This has become a very competitive market (at least eight products available in Germany). The units are mostly covered by maintenance contracts. The number of sales is expected to increase to address wastewater schemes of small and remote communities, although the revenue may remain marginal in the overall European MBR market. An example in USA is in eastern San Diego County, California, where expansion of an existing casino and development of a shopping mall required extension to the existing treatment facilities. The existing extended aeration system was converted to a ZeeWeed MBR allowing almost triple the capacity of the infrastructure (Melin et al., 2006). The scheme has been operational since July 2000 with the water quality meeting the California tertiary effluent standards for waterreclamation plants.
4.16.5.2 Commercialized MBR Formats As mentioned in Section 4.16.3.1, the first-generation MBRs in wastewater treatment used a sidestream format, in which feed was pumped from the bioreactor through an external membrane system. This approach was suitable for the early stage, small-scale applications for difficult-to-treat feeds. An alternative format was developed in the 1990s using modules submerged in the bioreactor tank, or in an adjoining compartment. This was much more cost effective for treating larger-scale flows with more easily treatable wastewater. The submerged format is available with modules either in a flat-sheet configuration or as hollow fibers or capillary membranes. Originally, the favored concept was to submerge the modules directly into the bioreactor for simplicity. However, in order to gain better control of the balance between the biological and filtration-treatment capacity, it is now more common to use the membrane in an external membrane tank (Brow, 2007). The external arrangement allows the size and design of the membrane tank to be optimized independently, with practical advantages for operation and maintenance. The sidestream approaches are also divided into two formats – the long-established traditional method of crossflow, now used only for the most difficult feeds, and the newer concept of airlift, which uses air to recirculate the feed and thereby significantly reduces energy demand. Both sidestream formats use tubular membranes.
4.16.5.3 Case-Specific Suitability of Different Formats The competing MBR formats based on submerged and sidestream configurations each have their own pros and cons for different application types and plant size. The energy cost for the aeration to control membrane fouling in the MBR is of an order similar to the microbiology aeration for an easy-to-treat feed, and increases by 2.5–3.0 times for the more difficult feed (Cornel and Krause, 2006).
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Crossflow is more energy intensive – very high cross-flow velocities (up to 5–6 m3 h1) may be necessary to control the fouling; but for the more difficult feeds, it may be the only option that works reliably. Airlift is a more cost-effective way of improving mass transfer through the creation of slug-flow conditions in the lumen of the membrane tubes (Laborie et al., 1997), but there is a limit to how much air flow can be used while retaining slug-flow conditions. Airlift technology has a power cost similar to that of the submerged technology. In general, submerged MBR formats based on hollow fibers have been found to provide the most cost-effective solution for large-scale, easy-to-treat applications. Technology has been developed with optimized packing density and aeration bubble size to achieve stable performance at minimum energy use (Fane et al., 2005). However, this format can experience operational difficulties due to fibers becoming matted close to the potted ends, and therefore pretreatment and removal of hairs and fibers is essential. Hollow-fiber technology hence requires more instrumentation and control. The submerged MBR formats based on flat sheets have been found to be cost effective for similar types of wastewater, but due to higher air use and lower compactness, tend to be selected for small- to medium-scale duties. The flat-sheet format has operational advantages in terms of plugging and cleaning, and has been used in somewhat more difficult feeds. Flat-sheet systems have the advantage of relatively low manufacturing cost compared to hollow-fiber systems. However, packing density tends to be significantly lower than a hollow-fiber system (e.g., by a factor of 2.5–3 times). Therefore, flat-sheet systems tend to have a cost advantage for smallto medium-scale systems, whereas hollow fiber becomes more attractive at large scale due to the footprint advantage (Pearce, 2008b). The comparison is made more complicated, however, since aeration costs for hollow-fiber systems are often lower. This means that the most cost-effective solution for total treatment costs at medium scale is closely contested, and both approaches are found across the size range due to site-specific circumstances, which could favor either solution. Lesjean et al. (2004), taking into account the current knowledge, anticipated a future market share as follows: for municipal applications, it is expected that the hollow-fiber submerged configuration would be competitive for mediumto large-size plants. For small to medium sizes, flat-sheet technologies would have an advantage. However, in case of larger plants, or a plant refurbishment, the alternative membrane scheme (secondary/tertiary treatment followed by an MF/UF membrane filtration) is very likely to be cost competitive, unless high-cost land has to be purchased for the construction. This multi-barrier scheme will also be easier to control and to optimize because of the disconnection of the treatment steps. It will also be associated with the lowest risk in relation to the membrane operation, as the membranes will be operated under smooth hydrodynamic conditions in terms of particle matter, turbulence, and backwash re´gime. In a recent paper, Lesjean and Huisjes (2008) reiterated this expectation despite the present trend of large MBR plant construction. The airlift format has been developed as a low-energy alternative to the energy-intensive cross-flow sidestream format,
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which has been used historically for the most difficult feeds. As mentioned earlier, the energy cost of crossflow prohibits it as a treatment option for any application other than small scale or where there is no other treatment option. However, the airlift has very low energy use, and may even undercut the energy requirements of the submerged options, due to the advantage of containment of the feed inside the tubular membrane (Van ‘T Oever, 2005; Futselaar et al., 2007). Since airlift eliminates operator contact and has good operational characteristics, it may as well make a major impact on the MBR market in the long run. Pearce (2008a, 2008b) argued that the airlift format may find applications throughout a broader range than the submerged formats. Figure 15 depicts the concept of airlift MBR.
4.16.5.4 MBR Providers 4.16.5.4.1 Market share of the providers The global market value of MBR is expected to rise up to US$500 million by 2013 from around US$300 million in 2008 (BCC Research, 2008). The MBR market is dominated by three companies, namely GE–Zenon, Kubota, and Mitsubishi Rayon Engineering (MRE). Only GE–Zenon and Kubota have a strong presence in Europe and North America, while MRE have until now mainly focused on sales in Asia. All these companies use submerged formats, with GE–Zenon and MRE Air release
Return to bioreactor
Permeate
Permeate backwash
Air injection
Airlift Feed supply Figure 15 The concept of airlift MBR.
using hollow-fiber membranes, and Kubota, flat-sheet membranes. Another three companies too have an international presence, namely Siemens–Memcor, Norit, and Koch-Puron, but the sales for these three companies makes up a small portion of the worldwide market. Among the latter three, Norit promotes the airlift format. Figure 16(a) shows the worldwide relative market share (in terms of installations numbers) for the three large players (Yang et al., 2006; Pearce, 2008b). The MBR market has several dozen regional or application specialists, quite a few of who use flat-sheet formats as adopted by Kubota: for example, Japan’s Toray and A3 from Germany. In addition to these international companies, there are a further 30 companies in the European Union (EU) market that have either significant regional presence, or an application focus, or a low-level international presence (Lesjean and Huisjes, 2008). Many of these companies are significant in the local markets, but individually, they have a small share of the international market. It is interesting to note that the MBR market has characteristics different from that of the UF/MF market. In UF/MF, there are 10–12 significant players with worldwide presence, with four market leaders, none of who dominate the market. Besides these companies, other regional players are relatively insignificant (Pearce, 2008a, 2008b). Zenon is long established in the market and has been one of the major companies promoting the MBR concept, and the use of PVDF membranes. The North American market is dominated by Zenon (Yang et al., 2006) as shown by the revenue share illustrated in Figure 16(b) and has many more opportunities in the municipal sector than in industry. Zenon leads the European market as well (Figure 16(c)). Kubota was one of the early pioneers of the MBR concept, encouraged by a Japanese Government initiative in the 1980s. They achieved a very large number of installations in small- to medium-scale systems, initially focusing on the residential/ commercial market in Japan and have approached export markets through exclusive partnerships. Kubota has a significantly greater number of plants than Zenon, with a slightly higher proportion of industrial plants. Many of Kubota’s installations in Japan and Korea are for small-scale municipal and domestic applications. Figure 17 shows the market characteristics of the two market leaders, Kubota and Zenon, illustrating the significantly different market strategies with regard to the size of plant targeted. Kubota is the strongest market player for industrial and small-scale municipal applications. MRE is a long-established supplier of MBR, with a very strong position in the relatively mature MBR market in Japan and Korea. There are a large number of references for this technology in Asia, but relatively few installations elsewhere. MRE also has a very large number of installations, with a higher proportion of industrial users, mostly with small flowrates. Koch Membrane Systems (KMS) is a long-established membrane manufacturer and membrane-systems company. In 2004, KMS acquired the MBR start-up company Puron, which had been founded in 2001. They introduced an approach to fiber potting different from that of the other hollow-fiber module providers.
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603
15
17
68
(a) Worldwide (relative installation numbers % in 2006)
2 3
6
10
20
33
65
61
(b) North America (revenue % in 2003)
GE−Zenon
Kubota
(c) Europe (installed membrane surface % in 2005)
Mitsubishi−Rayon
Siemens−Memcor
Koch−Puron
Others (N. America: Mitsubishi, Norit; Europe: Norit, Wehrle and other EU and non-EU suppliers) Figure 16 Market share of the suppliers. Data from (a) Yang Q, Chen J, and Zhang F (2006) Membrane fouling control in a submerged membrane; (b) Pearce G (2008 b) Introduction to membranes – MBRs: Manufacturers’ comparison: Part 1. Filtration and Separation 45(3): 28–31; and (c) calculated from Lesjean B and Huisjes EH (2008) Survey of the European MBR market: Trends and perspectives. Desalination 231: 71–81.
Memjet product is characterized by high permeability and packing density, providing a competitive position for capital and operating costs. However, worldwide market share for MemJet MBR is not very significant, since the company tends to focus on selected regional markets (Yang et al., 2006; Pearce, 2008b).
100 Plant capacity
80
60
No. of plants
% 40
20
0 Kubota
GE−Zenon
Figure 17 Relative market share (number of plants and capacity) showing distinct market strategies of the two market leaders.
Memcor have extensive experience in the use of their products in wastewater polishing. Their very fine polypropylene (PP) fibers developed in the 1980s were inexpensive and flexible, but unfortunately had low chlorine tolerance (Judd et al., 2004). In the late 1990s, Memcor developed a PVDF fiber, and now use the PVDF fiber for their MBR product range. The
4.16.5.4.2 Design considerations The design of the reactor (including membrane, baffle, and aerator locations) and the mode of operation of the membrane are key parameters in the optimization of the system. The leading MBR providers propose several MBR designs. In each case, the process proposed is very specific. Not only are the membrane material and configuration used different, but the operating conditions, cleaning protocols, and reactor designs also change from one company to another. For example, the flat-sheet membrane provided by Kubota does not require backwash operation, while hollow-fiber membranes have been especially designed to hydraulically backwash the membrane on a given frequency. The MBR industry first developed in Japan with the use of chlorinated polyethylene (PE) flat-sheet membrane by Kubota, and PE fibers by MRE (Stephenson et al., 2000). The modified PE is characterized by reasonable strength, flexibility, wettability, and resistance to chlorine. Although PE is normally made as an MF membrane, it has relatively low permeability, so process fluxes of PE membranes tend to be at the
Membrane Biological Reactors
Table 9
air-usage efficiency. In addition, the companies using hollow fiber use intermittent aeration, for example, based on a timer in the case of Zenon, or in proportion to flow in the case of Koch–Puron. Memcor introduced a novel cleaning method by using a mixture of air and mixed liquor, instead of using only air bubbles, to scour the membranes. The air bubbles effectively scour the membranes and the semi-crossflow of mixed liquor along the membranes continuously delivers the refresh mixed liquor to the membrane surface, minimizing the solidconcentration polarization at the membrane surface and therefore reducing filtration resistance. These enhancements have significantly reduced air usage and therefore power cost.
4.16.5.4.3 Performance comparison of different providers Few large-scale studies based on comparison of the commercially available MBR systems have been conducted. The city of San Diego, California, and the research consultant, Montgomery Watson Harza, have been evaluating the MBR process through various projects since 1997, including feasibility of using MBRs to produce reclaimed water (Adham and Gagliardo, 1998, 2000), optimization of MBR operation, and parallel comparison and cost estimations of the four leading MBR suppliers (Adham et al., 2004). MBRs were evaluated for their ability to produce high-quality effluent and to operate with minimum fouling. In terms of hydraulic performances, it (8.5−12)
500
400 (17−24) (50−60)
(17−24)
0
Toray
(29)
Norit
Siemens−Memcor
Koch−Puron
100
(30−34)
GE−Zenon
(17−24)
(14−26)
Mitsu. (PVDF)
200
Mitsubishi (PE module)
300
Kubota
low end of the range. Consequently, PE membranes are very cost effective at small scale, but struggle to compete in largerscale systems. In the 1990s, PVDF became established in MBRs through the reinforced capillary fiber in Zenon’s ZW 500 module (Yamato et al., 2006). PVDF has impressive performance in terms of strength and flexibility, but is significantly more expensive as a polymer. Nevertheless, PVDF membranes can achieve substantially higher flux, thereby overcoming price disadvantage. Recently, MRE also developed a PVDF-based membrane system. This membrane, designated as SADF, promises to be very competitive in both capital and operating costs, and despite it having a lower packing density than the PE product, it operates at much higher flux. With several companies now offering PVDF products in both capillary and flat-sheet formats, this is the dominant membrane polymer in the MBR market (Pearce, 2008c, 2008d). The third significantly used membrane polymer in MBR is a reinforced PES, used by Koch–Puron. Although PES is an important polymer in water treatment, in wastewater applications, its lack of flexibility limits the possibility of using air scour. Reinforcing the capillary does allow air scour, but at the expense of permeability. The Puron product uses reinforced PES rather than the PVDF, favored by its rivals. However, its main distinguishing feature is that the membrane fibers are potted at only one end. This overcomes the problem of fouling below the potting interface by hairs and fibers, which is a problem for the other hollow-fiber technologies (Vilim et al., 2009). Norit is the one major MBR company that offers a system based on a sidestream format with tubular membranes rather than a submerged format. Crossflow is only used for smallscale applications, with feeds that are difficult to treat, whereas airlift is cost effective for larger-scale municipal applications (Futselaar et al., 2007). Table 9 summarizes the specifications of the membranes used by different suppliers and Figure 18 compares the packing density and applicable flux of the membranes. Each of the suppliers makes regular improvements in air usage, since this has an important impact on total water cost. For example, the flat-sheet suppliers now use 1.5-m panels, which reduce air flow by up to 30% compared to the original 1 m panel (Pearce, 2008c, 2008d). In addition, they also use double-deck stacks wherever possible, which further improves
Membrane packing density, m2 m−3
604
Figure 18 Packing density (bar chart, m2 m3) and flux (values within parentheses, l m2 h1) of membranes from different suppliers.
MBR supplier specificationsa
Company
Membrane material
Pore size, mm
Membrane format
Fiber/tube dia (id,od),mm
pH tolerance
Kubota Mitsubishi Mitsubishi GE–Zenon Koch–Puron Siemens–Memcor Noritb Toray
Cl2 PE PE PVDF PVDF PES PVDF PVDF PVDF
0.4 0.4 0.4 0.04 0.05 0.04 0.03 0.08
FS HF HF HF HF HF TUB FS
– 0.37, 0.54 11, 2.8 0.8, 1.9 –, 2.6 –, 1.3 –, 5.2 or 8.0 –
1–13 1–13 1–10 2–10.5 2–12 2–10.5 1–11 1–11
a
All the membranes have moderate hydrophilicity and high chlorine resistance. All the companies except Norit use submerged format; Norit supplies airlift sidestream MBRs. FS, flat sheet; HF, hollow fiber; TUB, tubular.
b
Membrane Biological Reactors
was shown that all four processes were able to cope with flux rates exceeding 33 l m2 h1 and HRTs as low as 2 h. A 6-year development program has also been initiated for the introduction of MBR technology in the Netherlands market. Started in 2000, a comparative study of four 750 m3 d1 MBRs carried out by DHV water has been reported (van der Roest et al., 2002b). Three MBR plants, treating a design flow of 300 m3 d1 each, have been operated in parallel during 2003 and 2004 in Singapore (Le-Clech et al., 2006). A 4-year study, started in 2001, comparing the performance of Mitsubishi, Kubota, and Zenon MBR was conducted by the Swiss Federal Institute of Aquatic Science and Technology (EAWAG) (Judd, 2006). The Zenon MBR exhibited the most stable performance in the study. Although these studies have been conducted with the MBR systems running in parallel (with the same influent water), the MBR maximum flux, operating conditions and general design applied were those recommended by the suppliers, and therefore somewhat different for each system. This makes it difficult to make a fair comparison. Therefore, it is not possible to classify the MBRs as a function of their relative hydraulic performances, which need to be considered along with the cleaning protocols applied to each system. Mansell et al. (2004) performed measurements in which MS2 coliphage were seeded to the influent of a Kubota MBR (characteristic pore size 0.4 mm) and a Zenon MBR (characteristic pore size 0.04 mm). Permeate concentrations showed a log removal range of 3.2–7.4 for the Kubota installation and 5.32–7.5 for the Zenon installation. All of the heavy metals detected in the influent were removed to levels below detection limit, as well as the VOCs that were measured.
4.16.5.5 Standardization of Design and Performance-Evaluation Method The MBR market is very fragmented and exhibits many MBR filtration products with diverse geometries, module capacities, and operational modes (De Wilde et al., 2008; Lesjean and Huisjes, 2008). Although this situation promotes a competitive market, it is detrimental for the acceptance of the technology as a state-of-the-art process, and raises concern with potential clients or end users. From the point of view of the MBR operators, the possibility of interchanging filtration modules of different companies/suppliers would facilitate the replacement of the modules at the end of their life, and would reduce the risk of a supplier withdrawing from the market or releasing a new series of the product. In addition, the stakeholders in the industry employ various methods of membrane characterization and performance evaluation. This creates confusion among the users and prohibits fair comparison. Based on an extensive survey of the MBR industry, De Wilde et al. (2008) provided an overview of the market interests/expectations and technical potential of going through a standardization process of the SMBR technology in Europe. Due to the predominance of submerged filtration systems in municipal applications, the study focused only on this configuration. Two different aspects of standardization were considered:
•
standardization of MBR filtration modules toward interchangeable modules in MBRs and
•
605
standardization of MBR acceptance and monitoring test methods toward uniform quality-assessment methods of MBR filtration systems.
4.16.5.5.1 Standardization of MBR filtration systems In relation to the market expectations, about 20 potential technological, financial, economical, or environmental benefits/opportunities and drawbacks/threats of MBR module standardization for suppliers and operators were identified and mapped. It appeared that the number of advantages and disadvantages was quite balanced for both sides of the market, the main advantage perceived by the industry being that standardization should contribute to the growth of the MBR market. Other main advantages/opportunities are avoidance of vendor lock-in, price decrease, and increased trust and acceptance. Main disadvantages/threats for the end users are overdimensioning of civil constructions and supplementary works and costs to the peripherals during replacement. Main disadvantages for the module suppliers seem to be the higher competition, lower profit margins, and a limitation for innovative module producers to enter the market. From the technical point of view, the analysis showed that a standardization process common for both flat-sheet and hollow-fiber membranes/modules would not be realistic. In order to achieve interchangeability of filtration modules, not only should the prospect of pure dimensional standards for the module be considered, but also the design and mode of operation of the peripheral components, such as the filtration tank, pumps, blowers, and pretreatment should be borne in mind. More than 30 technical factors hampering or interfering with a standardization process were identified and quantified, and their relative potential for affecting the possible outcome was evaluated. For instance, four factors were grouped as the extremely high hindering factors: module dimensions, filtration tank dimensions, specific permeate production capacity, and specific coarse-bubble aeration demand. These factors are mainly the result of a completely different geometry and design of the filtration module and discussions for the standardization of MBR filtration systems should in essence focus on these factors. For each category, more or less the same number of obstacles lies ahead. Nevertheless, the nature of some of these obstacles or points of attention can be different. Some factors are specifically important for FS modules (e.g., flushing of air-supply pipes and design of a permeatecollection tank), and others for HF modules (e.g., type of prescreening, whether gravity filtration or any other type).
4.16.5.5.2 Standardization of MBR characterization methods The survey conducted by De Wilde et al. (2008) also revealed the respondents’ consensus in general on the positive impact of harmonization of membrane-acceptance tests at module delivery and monitoring methods on municipal MBR market growth. Some important parameters, for which a common definition and measurement protocol could be helpful, are mentioned below:
•
clearly defined and harmonized parameters to monitor membrane fouling, integrity, and aging;
606
• • • • • • • • •
Membrane Biological Reactors
a common definition of membrane lifetime for the guarantee clause; determination/definition of flux (operation and nominal design); common definition for sustainable peak hydraulic load; harmonized tests to check membrane performances over a defined period and under specific conditions; characterization method for membrane acceptance at module delivery; minimum requirements and technical methods to check membrane performance at plant commissioning; monitoring methods of normalized permeability in clear water, permeability in sludge, transmembrane pressure, and fouling rate; monitoring methods of sustainable flux and maximum flux; and operating conditions (biology and filtration systems) for warranty clauses.
It is interesting to note that, most of the newcomers in the market are developing their systems so that they can easily replace the products of the two main suppliers (Zenon–GE and Kubota). A standardization process driven by the end users could accelerate this evolution and contribute to the market development (Lesjean and Huisjes, 2008). Pearce (2008a, 2008b, 2008c, 2008d) also pointed out that, although the dimensions of the relatively newer Puron products are not identical to Zenon’s ZW 500d or MRE’s SADF, the elements are similar, and cassettes made from the elements could be used interchangeably. This begins to introduce retrofit possibilities into what hasuntil now been a fragmented market with no standardization.
4.16.6 Future Vision In addition to the alleviation of the technology bottlenecks illustrated in this chapter, a radical shift from the conventional concept of advanced wastewater treatment is deemed
Urine separation is also worthwhile to be considered
4.16.7 Conclusion MBR is a physicobiological hybrid process. The membrane provides a physical barrier for hygienically safe and clean water with the help of microbial–ecological treatment that can achieve good public acceptance. It is also well recognized by the experts that the clear membrane permeate makes post treatment easy; then, a variety of hybrid systems having the MBR as the core can be considered depending on the specific quality requirements of the reclaimed water . These advantages
A large amount of diluted organic wastewater (graywater)
To co-generation system A small amount of highstrength organic waste kitchen waste disposer-wastewater and toilet flushing)
imperative. In the context of sustainable water system, the advanced treatment must couple technologies to produce water of the required quality and realize material conversion from waste as well. The required quality does not always mean high quality. The quality comes from necessity. Membrane technology has the potential to be an on-demand quality provider just by separation. The conversion mainly comes from the biological reaction in the MBR. Three aspects of a sustainable society, namely, the low carbon society, sound material cycle society, and ecological society, are notable. From the point of view of sustainable water system, the advanced wastewater-treatment processes can be classified into the categories of energy saving (or productive), material productive, and ecologically oriented. The MBR technology might match more with the first two. However, present MBR technologies are still large energy consumers. Next-generation MBRs need to be developed to reduce the significant aeration requirement (by compact module design and sludge-concentration control techniques) and recover energy (e.g., by adding other organic wastes and combining anaerobic digestion for methane recovery). In line with the proposed definition of advanced treatment, the notion needs to be changed from organic wastewater treatment to water/biomass production by developing next-generation MBRs where the membrane acts as a separator of water and biomass and biomass is utilized for energy production. The concept is illustrated in Figures 19 and 20.
Anaerobic pretreatment Pretreatment Methane production
Biomass production from liquid organic waste (*)
Aerobic MBR
(*)
(A very small amount of residue) • Renewable energy utilization • IT-based maintenance service system • User participation in monitoring
(*) N,P recovery option
Figure 19 Next-generation MBR system: anaerobic combination for on-site small-scale advanced treatment.
Safe effluent
Membrane Biological Reactors
W.W.
Solid−liquid Solid-liquid separation
Solid concentration/ concentration/ Anoxic anoxic reaction reaction
Biosorption/ membrane separation/ aerobic reaction
607
Safe reclaimed water
Energy and/or material recovery process Other than biogas production, physicochemical treatments are also candidates for energy recovery, for example, supercritical water gasification of sludge−water mixture where the biomass sludge is utilized as energy source to produce hydrogen from water molecules (coupling clean energy production). Figure 20 Next-generation MBR system: renovation of existing wastewater-treatment plants.
make MBR a good device in water reclamation and/or advanced wastewater treatment. The continued push toward stricter discharge standards, increased requirement for water reuse, and greater than before urbanization and land limitations fuel the use of MBRs. However, there is room for improvement to utilize the potential of the MBR fully. The challenges will center on energy saving, ease of operation, simplified membrane cleaning and replacement strategies, and peak-flow management. The international adventure on R&D of MBR technologies continues.
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4.17 Anaerobic Processes DJ Batstone and PD Jensen, The University of Queensland, Brisbane, QLD, Australia & 2011 Elsevier B.V. All rights reserved.
4.17.1 4.17.1.1 4.17.1.1.1 4.17.1.1.2 4.17.1.1.3 4.17.1.1.4 4.17.1.2 4.17.1.3 4.17.1.4 4.17.1.5 4.17.2 4.17.2.1 4.17.2.1.1 4.17.2.1.2 4.17.2.1.3 4.17.2.1.4 4.17.2.1.5 4.17.2.2 4.17.2.2.1 4.17.2.2.2 4.17.3 4.17.3.1 4.17.3.2 4.17.3.2.1 4.17.3.2.2 4.17.3.3 4.17.3.3.1 4.17.3.3.2 4.17.3.4 4.17.4 4.17.4.1 4.17.4.2 4.17.4.3 References
Anaerobic Process Fundamentals Anaerobic Conversion Processes Hydrolysis Fermentation/acidogenesis Acetogenesis and methanogenesis from hydrogen Aceticlastic methanogenesis Physicochemical Processes and pH Temperature Inhibition and Toxicity Rate-Limiting Steps Selection and Design of Anaerobic Technology Anaerobic Digester Technologies High-rate anaerobic digestion Anaerobic ponds Fully mixed liquid digester Plug-flow liquid digesters Solid phase (leach bed) Digester Selection and Design for Specific Applications Domestic and industrial wastewater Sewage solids and activated sludge biosolids Interpretation and Operation of Anaerobic Systems Evaluating and Determining Controlling Mechanisms Performance and Process Indicators High-rate anaerobic reactors Sludge digesters Evaluating Substrate and Microbial Properties Activity testing Biological methane potential testing Advanced Model-Based Analysis Future Applications of Anaerobic Digestion Sewage Treatment and Nutrient Removal Nutrient Recovery Future Applications in Energy Generation and Transport
4.17.1 Anaerobic Process Fundamentals Anaerobic digestion is the biological conversion by a complex microbial ecosystem of organic and occasionally inorganic substrates in the absence of an oxygen source. During the process, organic material is converted mainly to methane, carbon dioxide, and biomass. Nitrogen released from converted organics is in the form of ammonia. Anaerobic processes for wastewater treatment have advantages over aerobic treatment in that there are no power requirements for air supply, production of sludges requiring treatment and disposal is much lower, and the methane production can be used for energy production. Aerobic processes are catabolically more favorable, yielding approximately 10 times the energy, with a correspondingly higher microbial yield (Madigan et al., 2009). For this reason, yields used for mixed heterotrophic processes are of the order of
615 615 616 618 619 620 621 623 624 625 626 626 626 626 627 627 627 627 627 628 631 631 632 632 632 633 633 634 635 636 636 636 637 637
0.63 gCODX gCODS1 (Henze et al., 2000) as compared to 0.05–0.1 gCODX gCODS1 for anaerobic processes. COD is the chemical oxygen demand and is a measure of organics. In this case, gCODX represents the biomass generated (in grams COD), while gCODS represents the substrate consumed (Batstone et al., 2002). This lower microbial yield results in decreased operating costs. The lower yield generally implies that extended solid-retention times are required to avoid washout of active biomass. This can be done either in parallel with an increased liquid retention time, or by separation of liquid and solid-retention times. Operation, design, and interpretation of engineered anaerobic processes have greatly advanced over the last 20 years. This improvement is based on a very good understanding of underlying concepts, which has allowed implementation of technology such that it will stably and reliably operate without intervention. The process itself has (1) multiple microbial
615
616
Anaerobic Processes
steps, mediated by different organisms; (2) different steps that can be rate limiting under specific conditions; (3) interaction with the physicochemical system, particularly weak acid and base inhibition of microbial processes, and (4) highly nonlinear behavior, particularly with respect to pH regulation and inhibition. Therefore, application of anaerobic technology needs careful thought, especially to achieve an optimally engineered process for a specific application. Fortunately, understanding of the underlying microbial and chemical processes is very good, both in the scientific and in engineering sectors. Good understanding of fundamentals, as outlined in this section, has allowed the use of anaerobic technologies in a wide variety of applications, as outlined in Section 4.17.2.
The different microbial groups mediating each step have been well characterized, and are from phylogenetically defined regions. As examples, all methanogenic organisms discovered so far are archaea, while acidogens and acetogens are largely bacteria. Aceticlastic methanogens belong to one of the two specific genera: Methanosaeta or Methanosarcina. As shown in Figure 1, under different conditions, different steps can be rate limiting. Specifically, for particulate or slowly degradable materials, hydrolysis is rate limiting. Under conditions of stress, or where the primary substrate is rapidly degradable, aceticlastic methanogenesis is normally rate limiting. The first condition normally results in decreased performance as undegraded substrate is washed out, while the second condition results in elevated, effluent organic-acid concentrations.
4.17.1.1 Anaerobic Conversion Processes 4.17.1.1.1 Hydrolysis Anaerobic digestion proceeds through a series of parallel and sequential processes by a variety of consortia as represented in Figure 1 (Batstone et al., 2002; Pavlostathis and GiraldoGomez, 1991). In contrast to aerobic digestion, where oxygen is an external electron acceptor, gaseous and dissolved products (largely methane and carbon dioxide) have the same combined carbon-oxidation state as the primary substrates. Thus, anaerobic digestion is largely constrained by the need to find appropriate internal electron acceptors. When this is impossible, hydrogen ions or bicarbonate must be used as electron acceptors via anaerobic oxidation to produce hydrogen or formate. This introduces thermodynamic constraints that bring in obligate syntrophic relationships between the electron producer and the methanogenic electron consumer (Schink, 1997). It is conceptually correct and convenient to group complex organics into carbohydrates, proteins, and lipids, and their soluble analogs of sugars, amino acids, and long-chain fatty acids (LCFAs). Any mixed organic stream can be represented by these components, while preserving full information of mass, energy density (or COD), and nitrogen content (Nopens et al., 2009). Anaerobic digestion processes consist of four main steps:
• •
•
•
Hydrolysis is an enzyme-mediated extracellular step which solubilizes particulates and substrates that cannot be directly utilized by the anaerobic organisms. Acidogenesis or fermentation is the conversion of soluble substrates such as amino acids and sugars, which can be converted largely without an external electron acceptor. The products are largely organic acids and alcohols. Syntrophic acetogenesis is the degradation of fermentation products to acetate using hydrogen ions or bicarbonate as an external electron acceptor. This process is coupled with hydrogen or formate utilizing methanogenesis, which maintains a low hydrogen or formate concentration. Acetoclastic methanogenesis is the cleavage of acetate to methane and carbon dioxide.
Processes such as homoacetogenesis (conversion of hydrogen and carbon dioxide to acetate), and its reverse, acetate oxidation to hydrogen and carbon dioxide, have not been included in Figure 1, but can be important in specific circumstances as outlined further in this chapter.
While the formal definition of hydrolysis is much stricter, as a digestion component, hydrolysis is a term that is used to refer to solubilization of complex particulate materials. The material can be regarded either as a mixture of the basic components (carbohydrates, proteins, and fats), or as a composite compound (e.g., homogeneous material such as activated sludge and yeast). Separate classification and analysis of composite material as a separate input was proposed in the International Water Association (IWA) Anaerobic Digestion Model No. 1 (Batstone et al., 2002), but this was found to be cumbersome, especially when representing both composites and primary aggregates (e.g., waste-activated sludge (WAS) and primary sludges), and the current trend is to represent all feed materials as a combination of carbohydrates, proteins, and fats (Nopens et al., 2009). There are three main pathways for enzymatic hydrolysis. 1. The organisms excrete enzymes into the bulk liquid where it adsorbs onto a particle or reacts with a soluble substrate (Jain et al., 1992). 2. The organism attaches to the particle and secretes enzymes into the vicinity of the particle. The organism benefits from the soluble substrates being released (Vavilin et al., 1996). 3. The organism has an attached enzyme which may double up as a transport receptor to the interior of the cell (Tong and McCarty, 1991). This method requires the organism to adsorb onto the surface of the particle. The actual mechanism used depends heavily on the nature of the material, reactor hydraulics, and solid concentration, but forms 1 and 2 in the list are variations on the same mechanism, and are the principal forms considered here. Steps in extracellular enzymatic hydrolysis include (Figure 2): 1.
2.
4.
Production of enzyme – production rate can decrease when there is excessive soluble substrate available (Ramsay, 1997). Steps 2, 3, and 6 are transport processes, which can be limited due to large particles, or in solid-phase systems due to inadequate carrier liquid. Adsorption processes that are limited by surface area.
Anaerobic Processes 5. 7.
•
Reaction rates that are limited by surface area and enzyme concentrations. Deactivation can be excessive when away from optimal temperature and pH.
•
While there have been complex models that include all of these functions (e.g., Humphrey, 1979), in practice, it is very difficult to properly validate these models, and the most commonly used model is the first-order one. The use of first-order models has been justified as ‘‘an empirical expression that reflects the cumulative effect of all the microscopic processes occurringy’’ (Eastman and Ferguson, 1981). First order (or slightly more complex) has also been found to be just as effective as more complex models (Vavilin et al., 1996). Hydrolysis commonly becomes rate limiting when
•
In a continuous mixed digester, without retained solids, hydraulic-loading rate becomes too high (there is not enough time to hydrolyze the solids). Mass-loading rate is generally not an issue, and higher concentrations allow higher loading rates. Normally, a minimum of 9 days of hydraulic-retention time is required for any significant degradation (see Section 4.17.2.2.2). Mixed carbohydrate feeds are among the slowest to degrade.
In a batch system, there is insufficient batch time. Batch digesters have a higher volumetric efficiency, due to kinetic considerations. In a plug-flow system, there is insufficient reactor volume. Plug-flow digesters are highly efficient on a volumetric basis. Time of contact with the active biomass can also be an issue if the system is not effectively mixed at the inlet.
Particularly for mixed systems (the most common form of digester), where hydrolysis is rate limiting, the hydrolysis rate determines the size of the digester. We now discuss the hydrolysis of various feed materials: 1. Hydrolysis of WAS. There has been a large amount of work investigating the rate and extent of WAS digestion, but only limited analysis of the actual mechanisms of cell solubilization as specific to activated sludge. It is a complex process, involving lysis of the cell, and subsequent degradation of both soluble and particulate cellular components (Aquino et al., 2008; Madigan et al., 2009). This is further complicated by the issue that microbial cells are naturally resistant to cell lysis by other cells, and that the cells are in flocs, with varying sizes. Degradability and hydrolysis rate have been extensively analyzed. As mentioned earlier, activated-sludge hydrolysis
Particulate carbohydrates, proteins, and lipids
Acidogens produce enzymes
Hydrolysis
Sugars and amino acids Fermentation acidogenesis
CO2 Alcohols and Long-chain organic acids fatty acids
NH3
Acetogenesis CO2
CO2 Hydrogen
Acetic acid
Hydrogenotrophic methanogenesis
Aceticlastic methanogenesis
Methane Figure 1 Key steps in anaerobic digestion processes.
617
Methane
CO2
May be rate limiting
618
Anaerobic Processes
4. Adsorption of enzyme onto surface
6. Transport of product to bulk
5. Reaction
2. Transport to bulk or local environment 1. Production of enzyme
7. Deactivation of enzyme 3. Diffusion from bulk to particle
Figure 2 Steps in enzymatic hydrolysis.
is an extremely complicated physical and chemical process that is, of necessity, represented as a first-order process (Eastman and Ferguson, 1981). Practical batch testing indicates that this complex material is well represented by first-order kinetics (Dwyer et al., 2008), while primary sludge (for example) has a far more complex kinetic profile, due to the presence of multiple primary substrates (Yasui et al., 2008). Extensive analysis also indicates that for untreated activated sludge, hydrolysis rates are relatively constant at approximately 0.1 d1 (Batstone et al., 2002; Eastman and Ferguson, 1981; Ge et al., 2010; Pavlostathis and Giraldo-Gomez, 1991). The degradability of activated sludge can be entirely related back to upstream sludge age, and longer sludge-age material will be less degradable (i.e., have a higher inert fraction; Ekama et al., 2007; Gossett and Belser, 1982). It is now widely accepted that material that is undegradable under aerobic conditions, is also largely undegradable under anaerobic conditions (Park et al., 2006; Speece, 2008). Therefore, material that is degradable under anaerobic conditions can be numerically calculated from the degradable fraction of the active aerobic biomass in the WAS (Ekama et al., 2007; Nopens et al., 2009). There is a wide range of pretreatment methods to increase sludgedegradability extent and rate (Aquino et al., 2008), and these are discussed further in the Section 4.17.2 of this chapter. 2. Hydrolysis of carbohydrates. Carbohydrates mainly originate directly or indirectly from plants. Generally, plant material is a mixture of cellulose (25–60%), hemicellulose (15– 30%), and lignin (15–20%) (Tong and McCarty, 1991). Straw, a commonly used feed material, consists of 70% cellulose and hemicellulose, 8% lignins, 15% mineral solids, and 7% other organic compounds (Hashimoto, 1986). The remainder is tannins, soluble sugars, and ash. The first two components are very similar and are digested anaerobically via similar mechanisms. Tong and McCarty (1991) list typical chemical compositions of lignocellulosic materials. Cellulose is made up of linear chains of D-glucose units. Hemicellulose is a branched polymer comprising several natural minor sugars. Ease of degradation depends on the nature (crystalline or amorphous) and chain length. Hemicellulose is of a shorter length (200 units), while cellulose can have a chain length of up to 10 000 units. Lignin is a dense three-dimensional polymer of aromatic molecules. It is hydrophobic and is linked by carbon as well as ether bonds. Conversion of lignin by anaerobic bacteria is unknown,
H
R
O
H
R O
H R O
N
C
C
N
C
N
H
H
C
C
C
H
Figure 3 Protein chain with amino acids linked by amide groups.
and high lignin contents (together with the presence of crystalline cellulose) generally restrict or prevent hydrolysis of the underlying cellulosic material (Yang et al., 2009). 3. Hydrolysis of proteins. Proteins are natural polymers of different amino acids joined together by peptide (amide) bonds. The backbone of a protein is a repeating sequence of one nitrogen and two carbon atoms (Figure 3). There are 20 amino acids found in nature. These are differentiated by the R group, which defines the function of the amino acid. A protein has three structural components: • Amino-acid composition and sequence (primary structure). • The three-dimensional shape as set by bond angles and hydrogen bonds forms a helical shape in complex proteins. This is the secondary structure. • The tertiary structure defines the macromolecular shape as set by bonding between di-sulfide groups and to a lesser extent, other inter-R bonding. There are two major areas of importance for hydrolysis processes. Amino-acid composition (primary structure) affects the products. The tertiary structure defines the proteins as either fibrous or globular. Fibrous proteins are structural materials such as keratin, which is protective, and collagen, which is connective. Globular proteins are often chemically functional and act as enzymes, hormones, transport proteins, or storage proteins. Hydrolysis of proteins can be rate limiting in the overall process, depending on ease of structure degradation (Pavlostathis and Giraldo-Gomez, 1991). Protein structure is one of the main factors affecting the rate of hydrolysis. Globular proteins are rapidly hydrolyzable, while fibrous proteins are difficult to hydrolyze (McInerney, 1988). In general, all proteins apart from the most rigid type of keratin (such as the outer layer of hair and fingernails) are hydrolyzable (Figure 4). There are three main groups of proteases: serine, metallo, and acid proteases which have alkaline (8–11),
Anaerobic Processes
619
The 1,3-specific lipases can only act at the outside bonds of the triglycerides, yielding 1,2-diacylglycerols and 2-monoacylglycerols. These glyceride esters are unstable and undergo acyl migration to 1,3-diacylglycerol and 1-monoacylglycerol. Subsequently, these can be degraded further by the 1,3-specific lipase to glycerol and free fatty acids. Fatty-acid-specific lipases catalyze the removal of a specific fatty acid, preferentially removing cis-D9-monounsaturated fatty acids. Other fatty acids are degraded very slowly, especially those containing an additional double bond between D1 and D9. Figure 4 Cow hair from an anaerobic reactor showing intact keratin (A) compared with degradation of interior by anaerobic organisms (B). Photograph by Dr Damien Batstone.
CH2 OH CH
OH
CH2-O-fatty acid CH-O- fatty acid
CH2 OH
CH2-O-fatty acid
Glycerol
Triglyceride
Figure 5 Glycerol and triglycerides.
neutral (6–8), and acidic (4–6) pH optimums, respectively (Ramsay, 1997). Enzyme production may be suppressed when readably biodegradable substrates such as glucose or amino acids are supplied (Patterson-Curtis and Johnson, 1989; Ramsay, 1997). 4. Hydrolysis of lipids. Lipids are glycerol bonded to LCFAs, alcohols, and other groups by an ester or ether linkage (Madigan et al., 2009). Fats and oils have all the alcohol groups esterified with fatty acids as shown in Figure 5 and these form the bulk of glyceridic material in mixed oils and fat with other glyceridic compounds, usually a result of processing. Hydrolysis is catalyzed by LCFA ester hydrolases, called lipases. These act at the lipid–water interface in enzymatic hydrolysis to degrade the insoluble reactant to soluble products. There is little work on degradation of lipids in anaerobic environments when compared with that on carbohydrate and protein substrates. Most of this has been focused on the rumen, reviewed by (McInerney, 1988). One particular characteristic of lipases is increased activity with insoluble rather than soluble lipids (Martinelle and Hult, 1994), indicating that the activity of lipases increases greatly when the concentration of triglycerides reaches saturation and forms a second phase. The lipases are adsorbed at the interface. As there is an adsorption mechanism, combined reaction and adsorption rate may be dependent on the surface area of the insoluble triglycerides. Bacterial lipases can be divided into three main types: nonspecific lipases, 1,3-specific lipases, and fatty-acid-specific lipases (Finnerty, 1988). Nonspecific lipases can hydrolyze any fatty acid triglyceride regardless of structure, acting at any of the fatty acids. These can completely hydrolyze the ester bonds acting equally at all alkyl sites.
4.17.1.1.2 Fermentation/acidogenesis Fermentation and acidogenesis refer to the same process of conversion of sugars and amino acids to simpler compounds (mostly acids and alcohols). Fermentation is commonly applied in biotechnology processes where the focus is on the product. Acidogenesis is applied in wastewater processes. In our opinion, fermentation is a more precise and preferred term. Fermentation is defined as the conversion of organics without an obligate external electron acceptor to produce both reduced and oxidized products. The two major groups of compounds subject to fermentation under anaerobic conditions are sugars and amino acids, which are discussed next. Fermentation of sugars. Anaerobic fermentation from sugars is likely the most widely applied biotechnology process worldwide. It is used to produce food products, renewable fuels, pharmaceuticals, and industrial chemicals. It is currently in focus for production of biofuels (e.g., ethanol and butanol). Historically, fermentation has been carried out by pure or specialized microbial cultures, which are constrained to produce specific products from sugars, based on their physiology and genetic capabilities. In anaerobic digestion processes, fermentation is mediated by mixed culture, and a wide range of potential products can be formed. Sugars ferment via the Embden–Meyerhof–Parnas (EMP) pathway to pyruvate, and subsequently to C3 products (propionate or lactate), or C2–C6 products via acetyl-CoA (Madigan et al., 2009; Figure 6). The most common products are shown in Figure 6, as determined in practical mixedculture fermentation tests (Ren et al., 1997; Temudo et al., 2008). Smaller amounts of additional compounds, including metabolic intermediates, are also often detected. Actual product mixes are regulated by a number of environmental conditions, including pH, gas-phase hydrogen concentration, temperature, and biomass retention time. It is reasonable to assume that hydrogen-rich reactions (e.g., production of acetate) would be enhanced at low hydrogen concentrations and production of alcohols enhanced at low pH (Ren et al., 1997). Regulation of mixed-culture fermentation is exciting, as it offers the possibility of producing fuels and industrial chemicals directly from raw feedstocks such as crop residues and straw. While a number of models have been proposed (Costello et al., 1991; Mosey, 1983; Rodrı´guez et al., 2006), none of these can effectively describe the mixture of products under dynamic conditions. The most promising current approach evaluates the thermodynamic driving forces under varying conditions (Rodrı´guez et al., 2006).
620
Anaerobic Processes 1glucose
4e−
2pyruvate
4e−
4e−
8e−
2lactate 2propionate
2CO 2
2e−
2e− 2acetyl-CoA 2e−=H2
6e−
2acetate
2CO 2 2e−
2e− 1butyrate
2ethanol
Figure 6 Major products from C6 monosaccharide fermentation. Excess electrons are removed as hydrogen as shown.
Fermentation of amino acids. There are 20 common amino acids, which can be divided based on the R group (Figure 3) into the following groups:
• • • • • • •
Alkyl R groups: glycine, alanine, valine, leucine, and isoleucine. Alcohol R groups: serine and threonine. Carboxyl R groups: Aspartic and glutamic acids. Nitrogen-containing R groups: lysine, arginine, and histidine. Sulfur-containing R groups: cysteine and methionine. Aromatic R groups: phenylalanine, tyrosine, and tryptophan. Proline, which forms an amide ring with the amide group.
Fermentation of amino acids can either be by direct oxidation, or by fermentation in pairs along a coupled pathway. The coupled pathway is termed ‘Stickland digestion’, and it has several properties: a. Amino acids are degraded as a pair. b. One of the pair of amino acids acts as an electron acceptor (i.e., it is reduced), and the other as the electron donor (i.e., it is oxidized). c. The donor amino acid is oxidized to NH3, CO2, and a carboxylic acid with a chain length one carbon atom shorter than the original donor amino acid. d. The acceptor amino acid is reduced to NH3, and a carboxylic acid with a chain length equal to the original amino acid. e. Amino acids can act as an electron acceptor, an electron donor, or as both, but there is no rule based on the R chain. f. In general, there is a 10% shortfall in electron-acceptor amino acids in commonly found proteins. Due to the properties of Stickland reactions, and because the amino-acid compositions of most commonly encountered proteins are known, it is possible to estimate the organic acids
produced from a given protein (Ramsay and Pullammanappallil, 2001). This, however, assumes that Stickland reactions are used. If the hydrogen concentration is low, uncoupled oxidation of amino acids can occur (Stams, 1994). Uncoupled degradation can also result in a higher energy yield and, as in Stickland reactions, energy is only produced from the oxidation reaction (during regeneration of carboxyl-CoA). Figure 7 shows coupled and oxidation reactions for alanine (which is always a donor acid), and glycine (which is always an acceptor). The degradation of alanine is the same in both cases, as it is oxidized during the coupled reaction. The only change is that electrons are wasted into hydrogen ions, rather than glycine.
4.17.1.1.3 Acetogenesis and methanogenesis from hydrogen Organic acids and alcohols are converted to acetate (oddchained organics to propionate also) by anaerobic oxidation. This process utilizes hydrogen ions or bicarbonate ions to produce hydrogen gas or formate, respectively. The thermodynamics of the oxidation reaction require that the electronacceptor end product (hydrogen or formate) be maintained at a very low concentration, and, hence acetogenesis is obligately linked to a hydrogen-utilizing reaction, such as methanogenesis (Batstone et al., 2006b; Boone et al., 1989). Hence, interspecies electron transfer (IET), in which hydrogen is the electron carrier, is vital to the growth of both microbes. Indeed, the only the syntrophic association is obligate, and other forms of electron carriers are possible – even direct electron transfer via microbial nanowires (Reguera et al., 2005). In anaerobic biofilms, the oxidizing organism is normally a bacteria, while the methanogen is an archaea, and can be directly observed in close relationship (Figure 8). Hydrogen (plus bicarbonate) and formate are functionally, and thermodynamically, very similar, with hydrogen having a higher diffusivity, and formate having a higher solubility. Advanced modeling has indicated that their microscopic characteristics will be similar in either of the electron carriers (Batstone et al., 2006b). In addition, the free energy of conversion between formate and hydrogen is relatively low (5.7 kJ mol1), and the two may exist in enzyme-assisted equilibrium (Thiele and Zeikus, 1988). Therefore, hydrogen can be regarded as the representative electron carrier. The thermodynamics of the reactions can be assessed by a free-energy calculation. For the reaction a A þ b B3c C þ d D (with stoichiometry a, b, c, and d), the adjusted free energy of reaction is (Madigan et al., 2009)
DG0 ¼ DG00 þ RT ln
½C c ½D d ½A a ½B b
ð1Þ
where DG0 is the adjusted free energy of reaction, DG0 is the standard free energy of reaction, and ½C c ½D d =½A a ½B b is the reaction quotient, or concentration of products divided by concentration of reactants. The adjusted free energy DG0 must be less than zero for the reaction to proceed. Based on standard concentrations in a digester (0.001 M organic acids and 0.1 M bicarbonate), the hydrogen thresholds for different acetogenic reactions can be calculated. These are shown in Table 1. These are thermodynamic thresholds, and the actual
Anaerobic Processes Coupled
Uncoupled
Oxidation C
COO− Alanine (donor)
NH 2 H3C
C
Oxidation
Reduction
H H2C
Glycine (acceptor) 2
H2C
COO−
Alanine (donor)
NH2
2e−
2e− COO−
O
Pyruvate, NH3
Pyruvate, NH3 CO2
Acetate H3C COO−
2e−
CO2
2e−
Acetyl CoA
Acetyl CoA
Acetyl phosphate
H3C Alanine
H2 2H+ H2
Energy Acetate
Acetate
Acetate + CO2 + NH3 + 4H
2H+
Acetyl phosphate
Acetate H3C COO− Energy
COO−
621
2glycine + 4H
2acetate + 2NH3
Alanine
Acetate + CO2 + NH3 + 2H2
Figure 7 Coupled and uncoupled conversion of alanine.
Figure 8 Syntrophic community of bacteria and archaea (anaerobic granule), engaged in acetogenesis and methanogenesis. Bar is 500 nm.
levels are higher. This indicates there is only a narrow region of hydrogen concentrations where these reactions may proceed. The mechanism of the reaction can be explained thus. Oxidation of butyrate and larger organic acids (C4þ) is by b-oxidation, a process in which larger organic acids are sequentially oxidized in a cyclic process. Two carbon atoms are removed as acetyl-CoA per cycle, and energy is recovered by substrate-level phosphorylation (Ratledge, 1994). The cycle continues until only acetyl, or propionyl-CoA, remains. This is converted directly to acetate or propionate. Unsaturated bonds are reduced directly with hydrogen (with the unsaturated bond as electron acceptor), in a favorable reaction (Ratledge, 1994). While common organisms oxidize a range of C4þ fatty
acids (McInerney et al., 1981; Roy et al., 1985), the kinetics, particularly of branched chain fatty acids, can vary substantially (Batstone et al., 2003). Propionate conversion is by a limited number of specialized organisms, with the carboxyl group being converted to carbonate, and the two methyl groups being randomly converted to either the methyl or the carboxyl group on the final acetate product (de Bok et al., 2004; Stams and Plugge, 1994). Ethanol was the first observed syntrophic methanogenic culture (Bryant et al., 1967), and due to favorable thermodynamics, it was found to accumulate substantial hydrogen before thermodynamic limitations set in. Degradation was found to be via acetyl-CoA. The major pathway of hydrogen or formate removal in mesophilic high-rate reactors is methanogenesis. This occurs by activation of a carbon dioxide molecule or formate molecule and successive hydrogenation of this complex. As a final step, methyl-CoM is formed and this is reduced to methane with a yield of 1 (adenosine triphosphate (ATP) mol1 methane formed. None of the methanogenic archaea can utilize energy from substrate-level phosphorylation and ATP is probably generated from a proton-motive force (Boone et al., 1993). While methanogenesis is the major sink for electrons in anaerobic systems, there are a number of alternative sinks, including nitrate reduction, sulfate reduction, iron reduction, and homoacetogenesis (formation of acetate from hydrogen). Alternative electron acceptors, such as nitrate, sulfate, and Fe3þ, are preferred substrates to hydrogen ions. Homoacetogenesis can occur whenever there is elevated hydrogen, but is commonly observed under lower temperatures, where the thermodynamics of this reaction are more favorable.
4.17.1.1.4 Aceticlastic methanogenesis This is the major methanogenic step, where acetate is cleaved to methane and carbon dioxide. Only a limited number
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Anaerobic Processes
Table 1
Acetogenic reactions and hydrogen thresholdsa
Reactant
Reaction
H2 threshold (Pa)
Propionate
CH3 CH2 COO þ 3H2 O-CH3 COO þ 3H2 þ HCO3 þ Hþ
10
þ
Butyrate
CH3CH2CH2COO þ 2H2O-2CH3COO þ 2H2 þ H
100
Valerate
CH3CH2CH2CH2COO þ 2H2O -CH3CH2COO þ CH3COO þ 2H2 þ Hþ
100
þ
Ethanol
CH3CH2OH þ H2O-CH3COO þ 2H2 þ H
Palmitate
CH3(CH2)14COO þ 14H2O -8CH3COO þ 14H2 þ 7H
H2, HCO 3
4H2 þ HCO3 þ H -CH4 þ 3H2 O
1000
þ
10
þ
0.2
a
For acetogenic reactions, concentrations must be below the threshold levels. For the last reaction, concentration must be above the threshold.
of methanogens within the archaea have been identified that are capable of cleaving acetate:
• •
Members from the genus Methanosaeta within Methanosaetaceae – these are obligate acetate cleavers. Members from the genus Methanosarcina within Methanosarcinaceae. Members of this genus can also utilize hydrogen, CO2, and methylated C1 compounds (Ferry, 1993).
Methanosaeta is more pH, and ammonia, sensitive and dominates at below 103 M acetate (Zinder, 1993), while Methanosarcina is found outside these conditions, generally in high-ammonia conditions where there is also higher-effluent organic acids (Karakashev et al., 2005). Recent work has indicated that Methanosarcina may, instead of cleaving acetate to hydrogen and carbon dioxide, oxidize acetate to hydrogen, with subsequent reduction by a syntrophic methanogenic partner to methane (Karakashev et al., 2006) (Table 2). Therefore, under these conditions, Methanosarcina does not act as a methanogen, but simply provides electrons to another methanogen via hydrogen or an alternative electron carrier.
4.17.1.2 Physicochemical Processes and pH Physicochemical processes are those that are not biologically mediated, and hence occur spontaneously in water systems. This research field is generally referred to as aquatic chemistry (Stumm and Morgan, 1996). Some important physicochemical reactions that occur in anaerobic digesters are shown in Figure 9, and include 1. Association and dissociation of weak acids and bases such as water, organic acids, carbon dioxide, and ammonia – this is a rapid process. 2. Gas–liquid transfer of carbon dioxide, methane, hydrogen, and hydrogen sulfide – this is a medium-rate process. 3. Metal-ion precipitation to form solid precipitates – this is a medium-slow process. Unlike biochemical reactions, almost all physicochemical reactions are spontaneous and reversible. Therefore, equilibrium calculations are an important issue to assess physicochemical systems. The physicochemical state is most commonly expressed by the pH, or negative log of the hydrogen-ion concentration (–log10[Hþ]). It expresses the net balance of strong
Table 2
Acetoclastic methanogenesis
Substrate
Reactiona
Acetate (cleavage)
CH3COO þ H2O - CH4 þ HCO 3
þ Acetate (oxidation) CH3COO þ 4H2O - 2HCO 3 þ 4H2 þ H
DG0 31 þ 105
a
Reactions for coupled acids are shown. DG0 was calculated for reaction at pH 7.
and weak acids present in the system, but not their individual concentrations or strength. The physicochemical and biochemical reaction system are strongly linked in anaerobic digesters, through the following mechanisms:
• • • •
•
Biochemical reactions produce weak acids and bases, including organic acids, LCFAs, ammonia, and carbon dioxide. Biochemical reactions produce gases. Low pH inhibits biological activity through disruption of homeostasis and denaturing of enzymes, though specialized organisms can operate at extremes. The free form of many weak acids and bases, particularly ammonia, organic acids, and hydrogen sulfide, is inhibitory to organisms (Batstone et al., 2002). This means that not only does the total concentration of the parent compounds (e.g., inorganic nitrogen, sulfides, etc.) have an impact, but the pH also has an influence by determining the concentration of the inhibitory form (e.g., ammonia and hydrogen sulfide). Weak acids and bases buffer around their characteristic acidity coefficient (pKa, see further). This means that bicarbonate, in particular, resists pH changes around 6.3, since that is its pKa.
Of the three key classes of physicochemical reactions – acid– base, liquid–gas, and metal-ion precipitation – only the first two have been extensively addressed in anaerobic digestion models (Batstone et al., 2002). This is a clear limitation in anaerobic-digestion modeling, since the behavior of more concentrated systems, and particularly the behavior of solids cannot be effectively described without describing metal-ion precipitation (Batstone, 2009). Acid–base reactions are characteristically rapid, and can hence be described by the equilibrium equation. For the
Anaerobic Processes
CO2
623
H2 Gas
H2O
CH4
Composites
Biochemical
Liquid Gas Inerts
Death/decay
Proteins
Carbohydrates
MS
AA +
Lipids
NH3
NH4
VFA−, HCO3−, NH4+
HVFA, CO2, NH3 HCO3−
Ca2+ Growth
H2
HAc
Microbes
−
CO2
CO32−
Gas
CH4
H2O
CaCO3
Physicochemical Figure 9 Biochemical (vertical) and physicochemical processes (horizontal) in an anaerobic digester. AA, amino acids; MS, monosaccharides; HVFA, associated organic acids; VFA, dissociated organic acids; HAc, acetic acid; Ac, acetate. Adapted from Batstone DJ, Keller J, Angelidaki I, et al. (2002) Anaerobic Digestion Model No. 1 (ADM1), IWA Task Group for Mathematical Modelling of Anaerobic Digestion Processes. London: IWA Publishing.
reaction acid 2 base þ Hþ, the equilibrium relationship is
½Base½H þ ¼ Ka ½Acid
ð2Þ
where Ka is the acidity coefficient, and is often expressed as pKa ¼ –log10Ka, in a similar way to pH. Most analytical methods measure or report the total species concentration:
½Totalmeas ¼ ½Acid þ ½Base
ð3Þ
These two equations can be combined to give either the acid concentration, or base concentration, as a function of the measured concentration, the pH, and the acidity constant:
Ka ½Totalmeas Ka þ ½Hþ
ð4Þ
½H þ ½Totalmeas Ka þ ½Hþ
ð5Þ
½Base ¼
½Acid ¼
The equation system is complicated when there are three reactive species (e.g., the inorganic carbon system containing
CO2, HCO3 , and CO3 2 ), or more (e.g., the phosphorous system containing four reactive species). Equations (4) and (5) are commonly used to produce acid–base speciation diagrams. An example is shown for the inorganic nitrogen acid– base system in Figure 10. This demonstrates the relationship between pKa, pH, and fractionation. The reason why fractionation is practically important is that many acids and bases are mainly inhibitory in their free or uncharged form. Ammonia is free as the base (NH3), which is why ammonia inhibition increases at elevated pH levels (discussed later in the chapter). Other acid/base pairs of importance are the organic acids (most volatile fatty acids (VFAs) have pKa levels of 4.6– 4.8), CO2/HCO3 pair (pKa ¼ 6.35), H2S/HS (pKa ¼ 7.05), NH4 þ /NH3 (pKa ¼ 9.25), and HCO3 =CO3 2 , (pKa ¼ 10.3) (Batstone et al., 2002). Gas–liquid transfer is normally described by equilibriumdriven dynamic gas–liquid transfer. Hydrogen and methane are relatively insoluble, while carbon dioxide and hydrogen sulfide are relatively soluble. This means that the latter two compounds have a substantial impact on the liquid system. Metal-ion precipitation is also generally described by equilibrium-driven dynamic relationships. The actual mechanism of crystallization is complex and includes a number
624
Anaerobic Processes 1 Fraction as acid (NH4+)
0.9 0.8
Fraction as base (NH3)
Fraction
0.7 0.6 0.5 0.4 0.3 0.2 0.1
pKa = 9.3
0 8
8.5
9
9.5
10
pH Figure 10 Inorganic nitrogen acid–base speciation vs. pH. Note that the total inorganic nitrogen is equally split between ammonia and ammonium at a pH of 9.3.
of different factors, including the presence of seed, presence of confounding compounds, such as inhibitors and promoters, and solution activity. While simple first-order relationships have been used in complex systems, these are normally ineffective (Batstone, 2009). The basic solubility of a precipitant is described by equilibrium. For the reaction aAbþ þ bBa–2AaBb, the equilibrium relationship is
KSP ¼ ½Aa ½Bb
ð6Þ
where KSP is the equilibrium constant, and is normally referred to as the solubility product. For convenience, it is also often represented as pKSP ¼ log10KSP. The higher the pKSP, the less soluble the compound. Examples include CaCO3 (KSP ¼ 8.25), FeS (KSP ¼ 18), and CaOH2 (KSP ¼ 5.3). Note that it is the pH-adjusted anion concentration that is to be used in Equation (6), and, therefore, CaCO3 precipitation is driven by the concentration of CO3 2 . This means that pH has a strong impact on metal-ion precipitation. Until recently, physicochemical models within anaerobic digestion models have been relatively simple, mainly consisting of acid–base equilibrium equations, a charge balance to determine hydrogen-ion concentration, and gas–liquid transfer (Batstone et al., 2002). More complex models have represented limited nonideality, including ion activity (Musvoto et al., 2000a), and precipitation (van Langeraak and Hamelers, 1997). Simpler models work well with dilute, ideal systems without metal-ion precipitation, but have poor predictive power in concentrated systems, or where precipitationaffected compounds exist. This has led to implementation of a more robust but more complex physicochemical framework for anaerobic digester systems (Batstone, 2009).
include 1. Reaction rates increase with increased temperature according to the Arrhenius equation (Siegrist et al., 2002). As a rule, anaerobic digesters are relatively sensitive to temperature, with temperatures below 30 1C causing a substantial loss in activity. 2. A rapid decrease in activity with abrupt temperature increases above the maximum (Van Lier et al., 1996). Normally temperature rises are maintained below 2 1C d1. 3. Decrease in microbial yields, and an increase in apparent saturation concentration (KS), with increased temperature (Van Lier et al., 1996) related to an increase in cell maintenance. 4. Shifts in reaction pathways due to changes in the free energy of reaction with temperature. This is particularly relevant for oxidative reactions, and acetate oxidation becomes more competitive as compared to aceticlastic methanogenesis at higher temperatures (Zinder and Koch, 1984), while the reverse reaction (homoacetogenesis from hydrogen and carbon dioxide) is more favorable at lower temperatures (Rebac et al., 1995). 5. Pathogen deactivation increases with temperature. These impacts occur across the temperature range, but operating modes have been split based on reactor operability and dominant microbial population into the following three temperature ranges:
• • •
Psychrophilic 10–30 1C. Mesophilic 30–40 1C. Thermophilic 40–70 1C.
4.17.1.3 Temperature
Psychrophilic conditions are largely environmental, while mesophilic and thermophilic conditions are largely in engineered systems. There are also a number of physicochemical impacts:
Temperature has a number of impacts on outputs and internal processes in anaerobic digesters, including both biochemical and physicochemical impacts. Biochemical impacts
1. Increased temperature causes decreased gas solubility. 2. Volumetric gas production increases with increased temperature due to thermal expansion.
Anaerobic Processes
3. A change in temperature changes the solubility of solids. This may increase or decrease depending on solid enthalpy of precipitation. 4. Gas transfer rates increase, due to increases in diffusivity. 5. Increased temperature increases the water-vapor fraction in the gas phase. 6. The acid–base pKa values change with temperature (generally decreases). The variation in this is enormous. Organic acid pKa is relatively unaffected by temperature, while ammonia pKa changes dramatically. 7. Liquid viscosity increases with increased temperature. This changes the energy required to pump and mix reactor contents. Overall, as temperature increases from mesophilic to thermophilic conditions, the combination of all of these impacts can be observed as follows:
• • • • •
•
Rates increase due to increased activity. This can be especially important in hydraulic limited systems. Effluent organic-acid levels increase due to increased maintenance and substrate-saturation levels. Gas quality drops as the water and carbon dioxide fractions increase. Gas production increases because of increased activity and thermal expansion. pH is normally relatively stable. It drops due to increased organic-acid concentrations and lower pKa values, but rises due to decreased CO2 solubility. The net effect can be an increase or decrease depending on the feed type and reactor performance. The system is more susceptible to ammonia inhibition, due to a decrease in ammonia pKa, and hence, there is a higher concentration of free ammonia (see next section).
4.17.1.4 Inhibition and Toxicity Speece (2008) uses two definitions within the area of general restriction of biological processes: ‘‘inhibition: an impairment of bacterial function’’ (p. 432) and ‘‘toxicity: an adverse effect (not necessarily lethal) on bacterial metabolism.’’ Commonly, inhibition is reversible, while the effects of toxicants are irreversible. That is, if an inhibitor is removed, bacterial function will return to normal levels, while if a toxicant is removed, a portion of the population will have residual effects (e.g., be dead). Inhibition is measured by the IC50, or concentration at which bacterial catabolic rate is reduced by 50%, while toxicity is measured by a LD50 or median dose – dose which will kill half the population. While there are mechanisms or chemicals that particularly influence specific functional groups, methanogenic archaea are generally more vulnerable to inhibition, and toxicity than bacteria. The order of the least-to–most-impacted processes is as follows: acidogenesis-hydrolysis-acetogenesis/hydrogenotrophic methanogenesis-aceticlastic methanogenesis. While it has not been well documented in the literature, propionate is an exception, in that it responds after acetate to initial overloads, but can remain in the effluent for long periods (1–2 weeks) after the initial overload. Inhibition is the more commonly observed phenomena in anaerobic digesters. The IC50 measure is directly applicable for
625
use in noncompetitive functions for dynamic modeling (Batstone et al., 2002), while toxicity is not commonly modeled, largely because modeling has limited capacity to address the impacts of toxicants. The mechanism of toxicants is often specific, acting on a particular mechanism of cellular metabolism. LCFAs are a common toxicant, which are thought to adsorb to the cell surface and block substrate and membrane proton transfer (Hwu et al., 1996). Other examples of toxicants include detergents, aldehydes, nitro-compounds, cyanide, azides, antibiotics, and electrophiles (Batstone et al., 2002; Speece, 2008). Inhibition can follow a number of different mechanisms, most of which either decrease the energy available from catabolism, or increase the amount of energy needed for maintenance. Common forms of inhibition are pH inhibition, ionic inhibition, product inhibition, and weak acid and base inhibition. They are discussed in detail in the following. pH Inhibition. pH inhibition is a combination of weak acid or base inhibition, disruption of cellular homeostasis, and reversible and irreversible protein denaturation. Most anaerobic organisms have a relatively broad pH optimum, with activity steady through the optimum. Anaerobic digestion operates best at a pH below 8.0, with activity of most organisms dropping above that pH, due to either free-ammonia inhibition, or other mechanisms. Lower pH is a combination of free-acid inhibition and pH inhibition. Since anaerobic digestion is mostly an acid-producing process, low pH inhibition is the most relevant form. Optimal pH levels for the different anaerobic biochemical functional groups are:
• • • •
Hydrolysis. Normally optimal above pH of 6.0, feasible up to 5.0. Acidogens. Optimal between 5.5 and 8.0, feasible up to 4.0 (Batstone et al., 2002). Acetogens/hydrogen-utilizing methanogens. Optimal between 6.5 and 8.0, feasible up to 5.0 (Batstone et al., 2002; Ferry, 1993). Aceticlastic methanogens. Optimal between 7.0 and 8.0, feasible up to 6.0.
As shown above, acid-producing microbes (acidogens and acetogens) have a higher tolerance for lower pH values than acid-consuming microbes (aceticlastic methanogens). An increase in load to a methanogenic digester will generally cause a decrease in pH, due to an increase in most acids, as well as the weak acid bicarbonate – even in an ideally operated digester. Where wastewater is poorly buffered, that is, where a lack of weak acids or bases causes poor resistance to pH changes, the pH can dip below 7.0 in response to substantial load increase. This can cause aceticlastic methanogens to be inhibited, which causes a further pH decrease due to accumulation of acetic acid. The overload is therefore self-reinforcing, and causes an acid overload. This can be difficult to recover from. This is mainly an issue in high-rate systems, where there is little or no ammonia release, a lower level of bicarbonate buffering, and a lower operating pH. Most high-rate anaerobic digesters operating on carbohydrate wastewaters require active base dosing to maintain a suitable pH, and this can form a major portion of the cost in these plants, though it is possible to reduce this by effluent CO2
626
Anaerobic Processes
stripping and recirculation (Ramsay and Pullammanappallil, 2005). Ionic inhibition. The mechanism of ionic inhibition involves increasing maintenance requirements, due to an increase in basic osmotic pressure. Sodium is the most relevant ion, with IC50 values between 5 and 30 g l1 depending on the level of acclimatization, function, and antagonistic or protagonistic ions (Feijoo et al., 1995). Acclimatization is possible and common. Product inhibition. Product inhibition occurs when products build up to the point where the catabolic reaction becomes unfavorable, that is, where the adjusted free energy of reaction as shown in Equation (1) becomes positive. The most common case is inhibition of propionate acetogenesis, caused by
Active transport of H+ (requires energy) CH3COOH (acetic acid) Passive transport
H+ CH3COO−
Cell; pH = 7.3
4.17.1.5 Rate-Limiting Steps
CH3COO− (Acetate) Bulk; pH = 7.0 H+ Active transport of H+ (requires energy) NH3 (Ammonia) Passive transport
H+ NH4+
Cell; pH = 7.3 NH4+ (Ammonium) Bulk; pH = 7.8 Figure 11 Mechanism of weak acid (top), and base (inhibition) by passive diffusion of the free form of the acid or base into the cell, and disruption of homeostasis.
Table 3
accumulation of hydrogen levels above those shown in Table 1, or substantial accumulation of acetate. Weak acid and base inhibition. Weak acid and base inhibition are caused by passive transport of uncharged acids (e.g., organic acids) or bases (e.g., ammonia) into the cell. These acids or bases then dissociate or associate within the cell to disrupt homeostasis (Figure 11). This causes increased maintenance requirements. Some important compounds causing free acid or base inhibition are listed in Table 3. While adjusting pH is a normal method to address free acid or base inhibition, it is an expensive exercise, due to the inherent buffering in most anaerobic digesters. Free-ammonia inhibition is likely the most commonly encountered form of inhibition, particularly in manure digesters and where the feed is proteinaceous, as the ammonia causes a high pH and acts as an inhibitory agent as well. This not only causes poorer overall performance, but can also cause more fundamental shifts, and (Karakashev et al., 2006) found that high-ammonia systems were dominated by Methanosarcina, oxidizing instead of cleaving the acetate. As the free form of ammonia is most important, and because temperature has a strong impact on the pKa, the entire system is heavily impacted by both temperature and pH. This is demonstrated in Figure 12, which shows that in a system with 2000 mg N l1 a thermophilic system (55 1C) will have a pH threshold of approximately 7.5 before strong inhibition occurs, while a 37 1C system will have a threshold of approximately 8.0.
This chapter outlines the key processes that occur in an anaerobic digester. In most cases, for a given wastewater type or reactor design, there is a rate-limiting step that needs to be managed in order to achieve optimal design and operation. The most common controlling mechanisms are hydrolysis and methanogenesis. Hydrolysis is normally the rate-limiting step for solid digesters (41% solids), where there are no other inhibitory factors present. For most solids, a retention time of 410 days is required (see Section 4.17.2.2.2), and at lower retention times, undigested solids will go to the effluent. Performance, as assessed by solid destruction, will decrease. Aceticlastic methanogenesis is normally the rate limitingstep in high-rate anaerobic wastewater-treatment systems, or where there is a higher level of inhibitors. Aceticlastic methanogenesis generally controls treatment systems where there are biomass limitations, or where the system is heavily loaded. Occasionally, (e.g., manure digesters), both hydrolysis can limit, due to slow solid degradation, and simultaneously, methanogenesis can cause elevated organic acids, due to ammonia inhibition.
Compounds causing free acid or base inhibition
Compound
Inhibitory concentration (free acid or base)
pKa
Condition at which inhibition occurs
NH3 (ammonia) H2 S (hydrogen sulfide) HVFA (organic acids)
1–2 mM (14–30 mgN l1) 2–3 mM (32–40 mgS l1) 0.2 mM (13 mg l1)
9.25 7.05 4.8
High pH Neutral and low pH Low pH
Anaerobic Processes
627
Free ammonia at total NH3/NH4+ of 2000 mg l−1 0.04
55°C
37°C
20°C
Free ammonia (M)
0.035 0.03 0.025 Inhibition strong 0.02 0.015
Inhibition significant
0.01 0.005
Inhibition starts
0 6.5
7
7.5
8
8.5
9
pH Figure 12 Free ammonia levels and ammonia inhibition.
In rarer cases, acetogenesis/hydrogenotrophic methanogenesis can be the rate-limiting step (e.g., hydrogen overload in a highly loaded high-rate system fed with soluble sugars), but increases in the higher organic acids may also be in response to an increase in acetic acid. In the following sections, these controlling mechanisms are discussed in context with technology selection, design, and operation.
4.17.2 Selection and Design of Anaerobic Technology 4.17.2.1 Anaerobic Digester Technologies Implementation of anaerobic digestion needs to address the two key issues of (1) maintaining sufficient retention time to allow for hydrolysis of particulate substrates and (2) providing beneficial conditions for aceticlastic methanogenesis, including maintenance of pH above 7.0. Technologies are split between wastewater-treatment technologies, which need to focus on goal 2, with extended sludge-retention times, but limited liquid-retention times, and those which need to focus on goal 1, with extended solid-retention times (Figure 13). Technologies except for high-rate systems are largely hydrolysis limited. Treatment technologies are summarized in, and described further, in the following sections (Table 4).
4.17.2.1.1 High-rate anaerobic digestion High-rate anaerobic digesters normally operate with extended solid-retention time, and short hydraulic-retention times, by integrating solid retention within the main digester (Figures 14 and 15). The most common type is an upflow anaerobic sludge blanket (UASB) reactor, in which liquid percolates through a partially settled sludge blanket. This operates with a flocculant sludge blanket, but relies on formation of anaerobic granular sludge (particles 4200 mm) for higher loading systems, especially if high effluent quality is to be maintained. High-rate digesters require a low solid feed, with relatively high amounts of soluble feed material, and are most often used for industrial wastewaters as well as domestic sewage
treatment (van Lier, 2008). Hydraulic-retention times are normally short with o48 h, while solid-retention times can be very long (4200 days, years). UASB reactors have a gas– liquid–solid separation in the upper part of the digester, while variations may include packing (Figure 14) in hybrid reactors, or extended super-high-rate/low footprint systems such as expanded granular sludge bed (EGSB) and internal circulation (IC) reactors. Other alternatives for high-rate anaerobic systems include anaerobic baffled reactors (multi-compartment reactors), fluidized bed or attached-growth systems, fixed-media anaerobic filters, anaerobic membrane bioreactors, and sequencing anaerobic batch reactors. UASB type systems are currently the market leaders in high-rate systems by a large margin (van Lier, 2008).
4.17.2.1.2 Anaerobic ponds Anaerobic ponds are a low-capital cost option, but they tie up land and require desludging approximately every 10 years, which can be excessively expensive (US$150 per dry ton). Anaerobic ponds are typically operated with very limited external control (e.g., temperature) and are therefore largely impacted by the local climate. This limits the effectiveness of ponds in colder regions. Overall costs are heavily driven by solid loading. Methane capture is relatively poor, and this results in an increase in greenhouse-gas emissions, and, generally, odors from the pond. Due to the large volumes, correction under failure can be extremely expensive or impractical. Anaerobic ponds have a depth of 5 m, with surface area determined by loading rates.
4.17.2.1.3 Fully mixed liquid digester Fully mixed digesters are most often applied to sewage sludge, activated sludge, and manure digestion (Speece, 2008). They are the most commonly applied configuration for anaerobic digestion. They operate as fully mixed reactors, with either gas recirculation or mechanical/liquid mixing systems. Mixing configuration is critical, and is reviewed further in (Tchobanoglous et al., 2003), particularly with respect to sludge
628
Anaerobic Processes 100
Plug flow
Hydraulic retention time (d)
Anaerobic ponds Liquid mixed digesters
10
Solid-phase leach bed
1
High-rate AD
0.1 0.01
0.1
1
10
100
Feed solids concentration (%) Figure 13 Anaerobic treatment technologies ranked by hydraulic-retention time (vertical axis) and solid concentration (horizontal axis).
digestion. Their configurations include cylindrical (normally with recirculated gas or liquid mixing) and egg-shaped (normally with mechanical mixing) systems. Maximum loading rate is heavily dependent on achievable solid levels, and performance can often be enhanced by pre-concentrating solids. Due to viscosity and heat-exchange consideration, the maximum in-reactor solid concentration is approximately 4% (feed concentration of approximately 8%). Costs are relatively high due to their engineered nature.
4.17.2.1.4 Plug-flow liquid digesters Plug-flow liquid digesters operate as a semisolid liquid (10– 20%) in a long polyethylene tube, vaulted brick, or concreteshaped reactor. Material is loaded at the front of the digester, and passes through to product at the end. As it is not mixed, contact with biomass is poor. These reactors have high kinetic efficiency, due to the plug-flow configuration, but are susceptible to lack of inoculation and topical souring. They are most often applied to agricultural solid digestion.
material removed). The latter is considerably more expensive due to solid handling and feed requirements. An alternative to in-reactor methanogenesis is recirculated leachate leach bed reactors (Figure 15). In this configuration, leachate is continuously percolated through a loop that includes the main solid phase leach bed, as well as a high-rate system to remove organic acids produced by the leach bed. This system has the advantage that overload and souring of the leach bed is far less likely, and gas production is steadier. The main disadvantage is susceptibility of the UASB reactor to solids. This type of system has been applied to municipal solid waste, and poultry litter (Rao et al., 2008).
4.17.2.2 Digester Selection and Design for Specific Applications Common wastewater types are shown in Table 5. As demonstrated in Figure 13, wastewater technologies are classified by their solid concentration. Specific considerations for application of technologies are given in the following sections.
4.17.2.1.5 Solid phase (leach bed)
4.17.2.2.1 Domestic and industrial wastewater
Solid-phase digesters are similar to an engineered, high-rate landfill, where material is loaded in a reactor, tumbler, or baskets, and leachate liquid is circulated through the reactor. Liquid percolates through the solid matrix and liberates organic acids, which are subsequently degraded to produce methane. It can be produced either in batches (where the system is reacted until no more methane is produced), or continuously (where material is continually added, and spent
The main criteria for application of high-rate granular anaerobic treatment technology are higher strength (4500 mg COD l1), low solids (2000 mg l1), and low oil and grease (o500 mg l1). Given these constraints, it is not surprising that 75% of applications of high-rate technology are on wastewater largely containing soluble carbohydrates and organic acids (e.g., cannery, brewery, confectionery, and distillery) (van Lier, 2008). High-rate anaerobic digestion has been
Anaerobic Processes Table 4
629
Anaerobic digestion technologiesa
Technology
Principle
Advantages
Disadvantages
Loading rate (kg COD m3d1)
High-rate digester/upflow anaerobic sludge blanket
Mainly liquid wastewater flows upward through a granular bed
Low footprint, low capital cost, very stable, produces good effluent
Intolerant to solids
10 (UASB) 20 (EGSB/IC)
Anaerobic pond
Large retention time mixed vessel
Low capital cost
Very high footprint Must be desludged Methane capture poor Can produce odors
0.1
Mixed tank
Dilution to 3–6%, and continuous feed in mixed tank. Retention of 20 days. Used across many industries
Established tech Easy to control Continuous gas production
Poor volumetric loading rate Expensive tanks Need dilution liquid Liquid (not solid) residue
1–3
Liquid plug flow
Dilution to 15%, and feed through a liquid plugflow reactor
Very high loading rates Continuous gas production
Need dilution liquid. Poor contact with active biomass. Liquid residue
5
Batch solid phase
Fill and react in a solidphase reactor. Can be an engineered landfill (but must be properly sealed). System is loaded, enclosed, and leachate/inoculum circulated intermittently
Can be very cheap Very high loading rates Good gas conversion due to retention of active biomass Easy to control via leachate No milling required
Non continuous system (gas–flow changes in quality and flow over time) Can be difficult to seal (gas seals) Needs loading and unloading
6–10
Continuous dry solid phase (plug flow)
Continuous feed of solid phase through a system. Recirculation of leachate around solid phase
Continuous gas and residue production Do not need dilution liquid Very good loading rates
Extremely high capital costs, and only really practical at very large scale. Very complicated mechanical system Potential solid handling issues
10
a
Note that the high loading of later options is achieved by high solids concentrations.
traditionally regarded as being less applicable to proteinaceous wastewaters, due to poor granule development (Fang et al., 1994). However, it is more likely that this is due to the particulate nature of these wastewaters (Batstone et al., 2004), and high-rate granular systems fed with soluble proteins (e.g., gelatine, casein) can be as effective as those fed with soluble carbohydrates (Moosbrugger et al., 1990). One of the main considerations associated with carbohydrate wastewater is buffering and pH. Carbohydrate wastewaters have no inherent buffering, which means that the acidity associated with carbon dioxide production needs to be offset by addition of a base. This can be a substantial cost consideration as outlined in the physicochemical section, although substantial savings can be achieved by effluent CO2 stripping and recycling. This is not as severe for protein-type wastewaters, as the weak base ammonia is produced during acidogenesis of proteins. Excessive ammonia release can cause free-ammonia inhibition. The flexibility of high-rate anaerobic digestion is illustrated by its applicability to domestic wastewater. Domestic
wastewater would normally be a poor feed source for high-rate anaerobic digestion, being low in strength (o1500 mg COD l1), relatively high in proteins, fats, and solids (often 4500 mg SS l1), and normally at lower temperatures. However, it has been successfully applied in both pilot and full-scale for removal of organics, and for sanitization (Seghezzo et al., 1998). This is further addressed in a later section.
4.17.2.2.2 Sewage solids and activated sludge biosolids Primary sewage solids (primary sludge) and activated sludge are the two main solid streams produced from activated sludge treatment plants. Primary sludge is material that can be settled out of raw sewage, and is relatively degradable, that is, 60– 100% can be anaerobically degraded, depending on the upstream catchment. Primary sludge has a relatively large lipid component (approximately 50% by COD; Siegrist et al., 2002; Speece, 2008). Activated sludge is a combination of microbial material produced during the activated sludge process
630
Anaerobic Processes Gas
Gas
Effluent
Effluent
Gas−liquid−solid Packing Granules
Inlet
Inlet
(a)
(b)
Figure 14 Upflow anaerobic sludge blanket (UASB) (a) and hybrid (b) systems.
Bleed stream
Gas Leach bed Makeup water
Overflow
Solid feed
High rate (UASB) Figure 15 Combined leach bed and high-rate system.
Table 5
Preferred technologies for different wastewater types
Application
Solids concentration (%)
Preferred technology
Design parameter
Nominal design parameter
Domestic or industrial wastewater
o0.2% (soluble solids may be up to 5) 2–7 2–7 10–30
High rate
Mass loading
10 kg COD m3d1
Mixed liquid phase Mixed liquid phase Solid phase
Retention time Retention time Retention time
10–15 days 20 days 30–50 days (batch)
Sewage solids, activated sludge Animal manure Organic solid wastes
(partially degradable), inert particulate material derived from influent material (not degradable), and undegradable cellular product (not degradable; Nopens et al., 2009). Activated sludge is a more homogeneous material than primary sludge, with a lower lipid content, and consequently higher protein and carbohydrate content. Overall, degradability is heavily dependent on sludge age (see further). The cost of biosolid handling and disposal can be a substantial fraction (30–50%) of overall wastewater-treatment costs, with cost being determined on a per wet ton basis. The
key considerations for sludge treatment are (1) volume and mass reduction, to reduce all costs associated with handling; (2) removal of unstable organics, to improve utility and storage options; and (3) pathogen removal, to increase utility and safety of the sludge product. The driving consideration is volume/mass reduction, since this determines eventual cost. Anaerobic digestion is effective in meeting all considerations, providing cost-effective solids and organics destruction, allowing essential pathogen destruction, and improving dewaterability of the final product.
Anaerobic Processes
70 60 50 40 30 20 10 0 5
10
15
20
25
30
0.9 0.8 0.7 0.6 0.5
Primary sludge khyd = 0.5 d−1
Activated sludge khyd = 0.3 d−1
Crop residues khyd = 0.1 d−1
10
15
40
Figure 17 Anaerobic degradability waste-activated sludge (WAS) vs. upstream sludge age.
0.4 5
35
Activated sludge age (d)
khyd = 1 d−1 Methanogenesis begins to fail
% of degradable fraction destroyed
1
temperatures, the degradability of activated sludge simply becomes too low for viable anaerobic digestion, as the biogas produced is insufficient to meet mixing requirements and provide sufficient energy for heating of the digester. In these cases, without a supplementary primary sludge stream, anaerobic digestion is no longer an option for sludge stabilization. For these poorly degradable streams, there are pretreatment options to improve both apparent hydrolysis coefficient and degradability. Lower-energy options, such as sonication, temperature-phased anaerobic digestion (TPAD), and enzymatic pretreatment, appear to largely act to increase apparent hydrolysis coefficient (Ge et al., 2010), and hence move the material upward as shown in Figure 16. High-energy options, such as thermal hydrolysis, increase the amount available and the hydrolysis coefficient, and hence can result in considerably enhanced performance (Batstone et al., 2009), though at higher capital and operating cost.
Activated sludge degradability (%)
Sizing of sludge digesters is driven by sludge hydrolysis rate coefficient (Speece, 2008; Tchobanoglous et al., 2003). Primary sludge is rapidly degradable, with first-order coefficients of the order of 0.3–0.5 d1 (Gujer and Zehnder, 1983; O’Rourke, 1968; Siegrist et al., 2002). Activated sludges are more slowly degradable, with hydrolysis rates of the order of 0.1–0.3 d1 (Ge et al., 2010). The impact that this has on digester sizing and performance is shown Figure 16, which demonstrates that as hydrolysis rate decreases, a longer retention time is required to achieve the equivalent efficiency. Apart from hydrolysis rate, the other major factor determining performance is degradability (fd). This represents the amount of material, either as COD, or as organic volatile solids (VS) that can be broken down to methane or biogas, respectively. For a perfect digester, it would be equivalent to the organic solids, or VS, destruction, but in most cases, the VS destruction is 70–90% of the degradable fraction. It has been extensively shown that the availability of material in both primary and activated sludge is the same across aerobic and anaerobic systems (Ekama et al., 2007; Gossett and Belser, 1982), and hence, degradability, or inert fraction can be directly translated between activated sludge and anaerobic models (Nopens et al., 2009). Primary sludge has a degradability of 60–100%, depending on the upstream catchment. A larger proportion of industrial input normally results in a lower net degradability. Activated sludge degradability depends heavily on the inert fraction remaining from the activated sludge process, and is hence heavily dependent on upstream activated sludge age. This was evaluated by Gossett and Belser (1982), and the results are summarized in Figure 17. Therefore, a WAS with an upstream sludge age of 15 days would be expected to have a degradability of 45%. With reference to Figure 16, a digester with a retention time of 20 days would be expected to have an efficiency of 85%. Therefore, feeding this digester with this sludge, an overall VS destruction of 45 85% ¼ 38% would be expected. As can be seen, WASs are generally poorer candidates for anaerobic digestion as compared to primary sludges. At higher sludge ages and/or higher activated sludge
631
20
25
30
Digester hydraulic retention time (d) Figure 16 Digester performance (% efficiency on degradable fraction) vs. hydraulic-retention time, and hydrolysis coefficient (khyd).
632
Anaerobic Processes
4.17.3 Interpretation and Operation of Anaerobic Systems Anaerobic reactor systems have been traditionally regarded as more difficult to control than aerobic wastewater systems. In practice, this is partially true, as commonly the only control handle is feed rate, which in the wastewater industry is largely determined by upstream considerations (Steyer et al., 2006). Aerobic wastewater treatment or digestion has a wider range of control handles, including aeration intensity and internal and sludge recycles and bypasses. Monitoring of anaerobic reactors and digesters is also more problematic, as the nonlinearity of the physicochemical process (see Section 4.17.1.2) means that simple sensors, such as pH and gas flow, have limited utility (Steyer et al., 2006). Conversely, anaerobic processes operate at higher loading rates and feed concentrations than aerobic systems, meaning that reactor sizes are smaller, and they can be readily overdesigned. In addition, anaerobic systems have large time constants (change relatively slowly) in the linear region, and it is suitable to choose a long-term optimal operating condition and aim for that set point.
4.17.3.1 Evaluating and Determining Controlling Mechanisms Optimal instrumentation, interpretation, and operation of digesters depend heavily on the controlling mechanism as discussed in Section 4.17.1.5, which is in turn dependent on configuration and feed type. Systems can be divided into 1. Methanogenesis controlled systems. Aceticlastic methanogenesis is the controlling mechanism in systems which are fed predominantly soluble wastewaters, or where pH buffering is poor. These are most often high-rate anaerobic digesters. Diminished performance is indicated by elevated acetate (and other organic acid concentrations), as well as liquidand gas-phase hydrogen concentrations (Pauss and Guiot, 1993), at mild overload conditions, and substantially decreased pH (o7.0) and gas flow during process failure. The kinetics are fast, and the process is highly nonlinear at the point between mild and severe overload. 2. Hydrolysis controlled systems. Hydrolysis is normally the controlling mechanism for systems fed with predominantly solids (41% solids), and performance is mainly determined by retention time of solids in the digester. The kinetics are relatively slow, and the overload mode is diminished performance in terms of gas flow, and unstabilized solids in the effluent. Poor performance is determined by analysis of these measures. Solid digesters are commonly well buffered due to release of ammonia from protein digesters, and pH is therefore a less-useful measure of process stability. Most of these systems are relatively linear and stable in response to changes in load. 3. Inhibited-hydrolysis controlled systems. These are relatively stable hydrolysis controlled systems (normally solid digesters), but which have a methanogenic inhibitor in the feed. The most common instance is manure digesters, where gas flow is largely determined by hydrolytic processes, and hence the retention time, while the presence of
ammonia causes substantially elevated organic-acid levels. In this case, the process itself is relatively stable, due to the ammonia buffering, but long-term performance may be poor both due to effluent organic acids (hence, lost biogas), and an inappropriate retention time for hydrolysis. Thus, the controlling mechanisms can be determined both from reactor and feed types (e.g., solid digesters vs. liquid-fed high-rate systems), and from direct analysis of solids destruction levels, and organic acid levels. This then leads to ongoing analysis of the most suitable performance indicators.
4.17.3.2 Performance and Process Indicators Suitable performance offline indicators or online sensors should provide an accurate reflection of process performance, related to process goals. If necessary, they should also offer potential for process correction, and in advanced cases, online process control. Again, selection of a suitable indicator or sensor depends heavily on the application or reactor type. As examples, the process goal for solid digesters is solid destruction (and hence gas production). A suitable indicator would be VS destruction, or gas yield. The process goal for high-rate anaerobic systems is good effluent quality, and process stability. Therefore, an indicator that provides early warning (prior to process failure) is desired. The process goal for fermentation systems is the extent of fermentation (or desired mix of organic acids). A good indicator would be product mix as measured by VFA concentration. Stability-state sensors for anaerobic digesters are difficult to apply online. As stated previously, there is a balance between the rate of hydrolysis and fermentation, and the rate of methanogenesis. When the production rate of organic acids exceeds the capacity of the digester to remove organic acids, the pH can drop, and the reactor sours. This is a hysteric process that can be extremely expensive, time consuming, and difficult to recover from. Start-up is a particularly hazardous period for this, as the biomass is nonacclimatized, and may vary in quality. The most simple sensors for anaerobic digesters are pH and gas flow. Due to system nonlinearity and stability characteristics, these measures are generally unsuitable (Steyer et al., 2006), only changing after the reactor sours. The best indicators are the intermediates, including volatile fatty acids (Pind et al., 2003), and measures such as bicarbonate alkalinity which indicate resistance to overload (Steyer et al., 2006). Overall, VFA concentration is by far the most widely applicable, direct, and meaningful measure of stability. However, the measurement of VFAs is generally an offline process involving measurement by gas chromatography-flame ionization detection (GC-FID), which is relatively slow and expensive, or titration, which is slow, and can be inaccurate in the presence of other buffers. While online methods for VFA measurement have been developed (Boe et al., 2007; Pind et al., 2003; Steyer et al., 2006), these are generally expensive, and/or require extensive sample preparation (e.g., online membrane filtration or gas-phase extraction), and there is still a need for a simple, relatively low cost online sensor to indicate anaerobic process stability state.
Anaerobic Processes 4.17.3.2.1 High-rate anaerobic reactors Most high-rate anaerobic digesters operate on mainly carbohydrate-based industrial wastewaters (van Lier, 2008), including agro-food, beverage, distillery, and pulp/paper. The objectives of the process are to remove organics, in order to reduce load to downstream treatment units, and potentially produce an effluent suitable for reuse or discharge to sewer. This requires good organic-acid removal, and process stability. The key performance measure is therefore VFA concentration, which indicates the level of acid-contributing compounds, as well as the bicarbonate, or partial alkalinity (PA – titration to pH 5.8). Titration to pH 4.2 indicates the total alkalinity (TA), which includes both bicarbonate, as well as organic acids. The contribution of the organic acids is termed intermediate alkalinity (IA ¼ TA PA). The current guideline for reactor stability is IA/TA r0.3, that is, the ratio of VFA contributed versus TA should be less than 0.3 (Steyer et al., 2006). Speece (2008) critiques this in influent analysis, pointing out that a large proportion of the total alkalinity must be also allocated to neutralize the CO2 produced during the digestion process, and suggests that reserve alkalinity (after production of CO2) is a better measure. For direct digester analysis however, the IA/TA measure is a reasonable indicator, though absolute values (of PA and VFA) should also be assessed. The two terms in this measure are contributed by bicarbonates and organic acids, and the IA/TA terms can either be measured directly (by offline, or online titration), or indirectly, and calculated by alternative methods, including Fourier transform-infrared (FTIR) (Steyer et al., 2006) and GC-FID, which are the standard analytical methods in commercial laboratories. For smaller systems, offline titration is simple, low cost, and relatively informative.
4.17.3.2.2 Sludge digesters As stated in Section 4.17.2.2.2, sludge and higher solid digesters can never achieve a high-quality effluent, due to the presence of inert solids. The main cost associated with primary and activated sludge digesters is disposal of the product, and the value of gas produced is relatively low compared to this. Therefore, the primary performance-related measure is organic solid destruction (or VS destruction). This is naturally related to methane production, since any solid destroyed must be created as methane. Secondary performance measures include stability (as measured by remaining degradable solids) and pathogen levels (Speece, 2008), and both of these are normally regulated in sludges, on the basis of vector attraction, usability, and disease control grounds. Additional performance measures may include mineral and nutrient content, odor, dewaterability, and texture, which are largely related to primary and secondary measures. As an example, a wellstabilized anaerobic biosolid product will generally have good dewaterability and low odor. Given that the primary measure is solid destruction, there are three ways to calculate this (Ge et al., 2010): mass-balance VS destruction, which assesses the flow of organic solids out, compared to the flow of organic solids in; Van Kleeck VS destruction, which is a modification of the mass balance to use VS fraction only; and apparent VS destruction of gas flow, which relies on the principle that organics destroyed must be
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converted into gas. All measures assume that the system is at steady state, or values are averaged over longer terms. If the system is not at steady state, mass balance VS destruction and gas flow VS destruction need to be adjusted for flow, and van Kleeck VS destruction cannot be used. Mass-balance VS destruction is calculated as follows:
VSdestroyed ¼ ðVSconc;in VSconc;out Þ=VSconc;in
ð7Þ
where VSconc is the concentration of organics as measured by the volatile solids method (g l1), and subscript in and out indicate concentrations in the inlet and outlet streams. Van Kleeck VS destruction is calculated as follows:
VSdestroyed ¼
VSfrac;in VSfrac;out VSfrac;in VSfrac;in VSfrac;out
ð8Þ
where VSfrac is the fraction of total solids that is volatile (VSconc/TSconc). Gas flow VS destruction is calculated as
VSdestroyed ¼
CODgas ðkg COD d21 Þ CODin ðkg COD d21 Þ
ð9Þ
where CODgas is the calculated gas flow COD in kg COD d1. It can normally be calculated as
CODgas ¼ 2:9Qgas pCH4
ð10Þ
where Qgas is the gas flow at standard temperature and pressure (N m3 d1), pCH4 is the partial pressure of methane (atm), and 2.9 is a conversion factor (kg COD N m3). CODin is the incoming COD, and can be either directly measured, and multiplied by flow for a kg COD d1, or calculated as
CODin ¼ 1:5VSconc;in Qin
ð11Þ
where 1.5 is the assumed COD:VS ratio for activated sludges (approximately 1.7–1.8 for primary sludges), VSconc,in is the influent organic solids (kg m3 or g l1), and Qin is the inflow/ reactor hydraulic flow (m3 d1). Each of these measures has specific advantages, and can be influenced by different systematic and random errors: 1. Mass balance VS destruction. It is sensitive to errors in flow measurement, and systematic sampling issues. For example, it is common to have differential settling around sample points, such that the solid concentration is not representative of the in-reactor, or the outlet concentration. 2. Van Kleeck VS destruction. It is sensitive to accumulation of minerals in the reactor (which will read as a false low destruction), or precipitation (which will read as a false high). It is not as susceptible to systematic sampling issues, as dilution of mineral and organic solids are normally consistent. It is not dependent on flow-rate measurement. 3. Gas flow VS destruction. It is the least reliable, and is dependent on correct flow measurement in both liquid and gas streams, as well as correct VS inlet measurement. It is also sensitive to the assumed COD:VS ratio, and this can vary significantly (e.g., longer sludge ages normally result
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in a higher COD:VS ratio). Finally, it assumes that the COD:VS ratio is the same across the digester. While it should not be used as a primary measure of VS destruction, it is a useful comparison to other terms. It is not sensitive at all on outlet flow measurement or outlet liquid-digester analysis. While VS destruction is the key indicator of performance, digester stability and health can be assessed as discussed in the previous section as either total organic-acid concentration, or a combination of organic-acid concentration and PA. Sludge and manure digesters normally have higher levels of inherent PA due to the pH rise produced by free-ammonia release during hydrolysis of proteins.
4.17.3.3 Evaluating Substrate and Microbial Properties
Methanogenic activity tests are common due to methanogenesis as a rate-limiting step, and because methane can be readily measured. In the case of methanogenic activity tests, the substrates (acetate, H2, and CO2) are direct precursors for methane, and activity may be determined from direct measurements of methane produced. This is not true for determinations of acetogenic, acidogenic, proteolytic, and hydrolytic activities where substrates are converted through several steps; thus, measurements of methane production are not sufficient to determine activity, as the methane-production rate will only reflect the slowest step of a more complex degradation process. Measurements of substrate depletion are more valuable in this type of test. Methanogenic activity testing is particularly important as methanogenesis is the final stage in any degradation process, and the slowest, and the most sensitive step. Methanogenic activity is estimated based on the initial rate of methane production during a controlled batch test. Only the initial
4.17.3.3.1 Activity testing Anaerobic activity tests are used to evaluate the performance of anaerobic sludge communities, and may be used to select an adapted sludge as inoculum, to estimate maximum applicable loading rates of certain processes or sludges, and to evaluate batch kinetic parameters. These tests can also be used to monitor possible changes in sludge activities over time due to the build up of toxic or inhibitory material or the accumulation of inert material. Activity testing cannot determine the presence or concentration of individual microbial species; however, relative activities indicate the balance of trophic groups within the community. The microbial activity of the different trophic groups determines the rate of each of the four main steps in anaerobic digestion and allows identification of the limiting step. The rate-limiting step will give information about the maximum organic load which can be applied to the system without causing a loss in its stability. Activity as identified by testing can be used together with other reactor indicators to promote stable operation. The quality of inoculum is important for prediction of degradation characteristics of novel waste materials. For example, in applications where the waste material is a complex organic solid, the hydrolysis step will limit the material available for fermentation. In applications where the waste is soluble or readily fermentable, the production of intermediates will be more rapid and a high methanogenic activity is required to balance this. For determination of activities of different trophic groups, model substrates should be used. The concentration of model substrate used in activity testing is a critical factor in the test set up. The initial concentration should be sufficiently high such that the biomass concentration is the limiting factor in the test, but sufficiently low to prevent inhibition of the microbial community as well. Model substrates used to determine the activity of the main trophic groups in anaerobic communities and recommended concentration ranges are shown in Table 5. A synthetic medium may be used in the assays to ensure that necessary nutrients/micronutrient/vitamins are available to allow optimal performance of anaerobic microorganisms. The composition of basic anaerobic (BA) medium recommended for anaerobic activity testing is given in Tables 6 and 7.
Table 6 Model substrates for determination of specific activities of trophic groups in anaerobic communities Trophic group
Substrate
Concentration range
Hydrolytic Proteolytic Acidogenic Acetogenic
Cellulose Casein Glucose Propionic acid n-butyric acid Acetic acid
1–10 g l1 1 g l1 1–2 g l1 0.5–1 g l1 0.5–1 g l1 1–2 g l1
H2/CO2 (80:20)
1 bar total
Methanogenic – acetoclastic Methanogenic – hydrogenotrophic
Table 7
Base anaerobic (BA) medium suggested for activity testing
Stock solution
Volume per l
Components (g l 1 of stock solution)
A
10 ml
B C D
2 ml 1 ml 1 ml
E
1 ml
NH4Cl, 100; NaCl, 10; MgCl2 6H2O, 10; CaCl2 2H2O, 5 K2PO4 3H2O, 200 Resazurin, 0.5 Trace metals: FeCl2 4H2O 2; H3BO3 0.05; ZnCl2 0.05; CuCl2.2H2O 0.038; MnCl2 4H2O 0.05; (NH4)6Mo7O24 4H2O, 0.05; AlCl3 0.05; CoCl2 6H2O 0.05; NiCl2 6H2O 0.092; EDTA, 0.5; conc. HCl, 1 ml; Na2SeO3 5H2O, 0.1 Vitamins: biotin, 2; folic acid, 2; pyridoxine acid, 10; riboflavin, 5; thiamine hydrochloride, 5; cyanocobalamine, 0.1; nicotinic acid, 5; P-aminobenzoic acid, 5; lipoic acid, 5; DL-pantothenic acid, 5
NaHCO3 Na2S 9H2O
2.6 g 0.5 g
From Angelidaki I and Sanders W (2004) Assessment of the anaerobic biodegradability of macropollutants. Reviews in Environmental Science and Bio/Technology 3: 117.
Anaerobic Processes
635
0.09 0.08
y = 0.2735 × −0.0191 R 2 = 0.9967
Methane (g COD)
0.07 0.06 0.05 0.04 0.03 0.02 0.01 0.00 0.00
0.05
0.10
0.15
0.20
0.25
0.30
0.35
Time (day) Figure 18 Methane generation during specific methanogenic activity assay.
linear methane-production rate is used to reduce the influence of biomass growth and adaptations or changes of the biomass characteristics. Environmental factors including substrate and nutrient concentrations and pH also vary during the tests. The methane-production curve for a specific methanogenic activity test is shown in Figure 18. This shows the slope of the curve, which in the linear region indicates the methanogenic activity. The specific methanogenic activity (SMA) is this slope divided by the VS present in the test vial. Methanogenic activity is generally higher than 0.2 g COD CH4 g VS1 d1 for digesters and industrial high-rate sludges, and may be far higher for laboratory-grown granules.
The ultimate methane yield represents the potential to recover energy during waste treatment; however, the value of energy produced is often only a small consideration in determining the feasibility of the anaerobic project. An example output from a methane potential test is shown in Figure 19. The key parameters used to indicate degradability of a complex feed are degradation extent (fd), the fraction of the substrate that may be converted to methane, and apparent first-order hydrolysis rate coefficient (khyd), an indicator of the rate at which conversion occurs. Determination of degradability parameters is critical in feasibility analysis, system design, troubleshooting, and competitive testing of inoculums. Hydrolysis is normally represented using a first-order model as discussed in Section 4.17.1 of this chapter:
4.17.3.3.2 Biological methane potential testing The biological methane potential (BMP) test is a simple batch assay used to determine the potential methane generated from anaerobic biodegradation of a mass of test substrate. In addition to potential methane (ml) generated per gram of substrate (wet, dry, and VS basis), the BMP assay is used in determination of parameters critical in process design, troubleshooting, and competitive testing of inoculums. The BMP test requires a test substrate to be mixed with a known good inoculum (containing a strong and balanced anaerobic community) in a controlled environment. BMP testing can be done at multiple scales ranging from several grams of test material up to tons (in pilot digesters). Extensive biodegradability testing of thousands of different materials in both aerobic and anaerobic conditions has been performed over the last 50 years; however, comparison of biodegradability data between studies in the literature has been limited by a lack of a common basis. Previously, factors including type of equipment, operating conditions, method of analysis, test compound, inoculum, and nutrient medium varied among studies and influenced the outcome of the batch assays (Rozzi and Remigi, 2004). However, this is improving with the publication of practical and standardized methods (e.g., activities of the IWA Anaerobic Biodegradation Activity and Inhibition (ABAI) Taskgroup; Angelidaki et al., 2009).
dS ¼ khyd S dt
ð12Þ
where S is the degradable portion of substrate, t is the incubation time, and khyd is the first-order hydrolysis rate constant. Determination of residual substrate, S, requires that the degradable fraction of the substrate is known. A simplified approach is achieved through the separation of variables and the integration of Equation (12):
ln
Pf P ¼ khyd t Pf
ð13Þ
where residual substrate at time t, is represented as the difference between the methane yield at that time P, and the ultimate methane yield Pf. Equation (13) will produce a linear curve when the degradation kinetics are of the first order, and the hydrolysis rate constant is represented by the slope of the curve. Alternatively, the first-order hydrolysis coefficient and degradability parameters can be estimated using a dynamic firstorder (single step) model. The first-order hydrolysis rate is used to estimate process-retention time and thus digester size, while the degradability fraction can be used to calculate the expected VS destruction during the process. Replicate testing
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Anaerobic Processes 250
Methane (ml CH4 .gVS−1)
200
150
100
50
0 0
2
4
6
8
10
12
14
16
18
Time (day) Figure 19 Example output from biological methane potential (BMP) test. Error bars indicate 95% confidence errors from triplicate batches. The line indicates the model used to return key parameters.
is essential to determine repeatability and variability in the test results. More advanced modeling methods, including nonlinear parameter estimation and parameter uncertainty evaluations, are now being used to determine a multidimensional parameter surface showing the confidence region of the parameter estimations (e.g., Figure 20; Batstone et al., 2009). Batch assays are used on the basis that they are appropriate for assessing the performance or potential of full-scale anaerobic processes. It is not possible to replicate and maintain equal environmental factors such as nutrient, buffer, pH, and gas-phase conditions between the full-scale reactors and batch tests. The mode of operation (batch or continuous, mixed or plug flow) also varies between full-scale reactors systems and the standard assay (batch, no mixing). Degradability is a characteristic of test material rather than the test conditions and, indeed, degradability estimated using BMP tests are similar to the degradability performance achieved in continuous full-scale anaerobic processes; however, the first-order hydrolysis rate is influenced by test conditions and therefore is not directly comparable (Batstone et al., 2009). However, results from batch tests represent a conservative estimate of parameters needed for system-feasibility analysis and design. Environmental conditions and substrate characteristics vary between the BMP test and reactor used as an inoculum source; as a result, the inoculum is rarely optimized for the test material and significant adaptation does not occur during the batch test. Inoculum should be collected from a reactor operating on a complex feed material to provide a diverse and balanced microbial population and ensure complete breakdown of the degradable portion of the test material. The issue of inoculum to substrate ratio has been evaluated in some detail (Fernandez et al., 2001; Neves et al., 2004; Raposo et al., 2006). The inoculum to substrate ratio must be sufficient to ensure that hydrolysis is limited by surface availability or substrate concentration, rather than microbial concentration. This would typically require that the inoculum volume is greater than 50% of the test volume.
4.17.3.4 Advanced Model-Based Analysis Dynamic modeling of anaerobic systems developed reasonably quickly from simple dynamic first-order models, largely reflecting only hydrolysis (Gossett and Belser, 1982; Pavlostathis and Giraldo-Gomez, 1991; Pavlostathis and Gossett, 1986) to more complex dynamic multi-step models that include all the steps shown in Figure 1 (Costello et al., 1991; Siegrist et al., 1993). These include complex interactions such as physicochemical models, ammonia inhibition, and the production of organic acids. The wide variety of multi-step models have been largely consolidated in the IWA Anaerobic Digestion Model No. 1 (Batstone et al., 2002), which was designed to be a broadly applicable generic model of anaerobic digestion processes. This has been adapted to a number of diverse applications (Batstone et al., 2006a), including high-rate, sulfate reducing, nitrate reducing, solid and manure digestion, fermentative, and solid-phase digestion. In particular, there has been substantial effort into including anaerobic digesters in whole-plant models that largely depend on the IWA activated-sludge models (Nopens et al., 2009). This has allowed relatively easy characterization of input streams to the anaerobic digestion model – something which has been classically challenging. Dynamic modeling has a number of very practical applications, as well as enables specific areas of research. Specific practical applications include
• • •
Scenario analysis prior to major process changes – particularly with respect to particular inhibitors (Batstone and Keller, 2003). Its use to determine degradability rate and extent properties of upstream materials in situ, rather than through BMP testing (Batstone et al., 2009). Dynamic and detailed assessment of caustic dosing requirements and optimization for alkalinity addition in comparison with static analysis.
Anaerobic Processes
637
Hydrolysis rate - khyd (d−1)
0.2
0.15
0.1 0.3
0.35
0.4
0.45
Degradability fraction (f d) Figure 20 Surface estimation of degradability parameters – for two-parameter estimates on BMP tests. The 95% two-parameter region is represented by the line, while confidence intervals represent uncorrelated, linear estimates of parameter confidence.
4.17.4 Future Applications of Anaerobic Digestion 4.17.4.1 Sewage Treatment and Nutrient Removal High-rate anaerobic digestion has now been evaluated extensively for treatment of low concentration and domestic sewage (Barber and Stuckey, 1999; Foresti et al., 2006; Seghezzo et al., 1998). It is generally suitable for removal of bulk organics, and to remove some pathogens (though not to standards). In comparison with conventional aerobic treatment, high-rate anaerobic treatment of domestic wastewater is relatively low in capital costs and distinctly lower in operating cost, does not require aeration energy (and can produce energy as methane), and is relatively low maintenance. The main disadvantages are (1) low removal of nutrients; (2) relatively poor removal of organics (60–90%), and (3) release of methane dissolved either in the liquid or directly from the digester surface. While it is a suitable alternative to no treatment, high-rate anaerobic treatment cannot produce an effluent suitable for direct discharge to inland watercourses, with minimal environmental impact. Phosphorus can be removed (and recovered by precipitation) and methane can be captured during treatment, or removed in aerobic or other post treatment, which can also be used to remove residual organics and pathogens (Barber and Stuckey, 1999; Chernicharo, 2006; Foresti et al., 2006; Seghezzo et al., 1998). However, the key issue is removal of nitrogen. Currently, the main method of nitrogen removal (nitrification–denitrification) requires carbon for denitrification, and anaerobic processes remove carbon. Partial nitration to nitrite reduces the carbon load. While there are processes that can remove ammonia, such as the biological process anammox (using nitrite as electron acceptor to remove ammonia), stripping, and adsorption (Foresti et al., 2006), most of these are applicable at higher ammonia concentration. The anammox process is probably the most promising for low-concentration ammonia removal, and ammonia concentration is possible, through both adsorption and
membrane processes. While there are challenges, anaerobic processes both at low concentration and in solids digesters offer sustainable, low-cost alternatives to conventional aerobic processes.
4.17.4.2 Nutrient Recovery Phosphorus and nitrogen are key components in many organic sources, including biosolids and manure. Phosphorus in particular is a key resource, since it is a nonrenewable resource. World reserves are substantial, and depletion of ready resources is not expected until later this century (Isherwood, 2000). However, demand is also increasing substantially, and this has led to dramatic price increases, particularly through 2008. Alternative, sustainable, and low-cost alternatives are therefore highly desirable, particularly where national reserves are low, or supply restricted. Anaerobic digestion is already used to stabilize organic biosolids and manure for agricultural applications. Anaerobically stabilized organic biosolid is an excellent fertilizer, generally with comparable impacts (per unit nitrogen) to mineral fertilizer (Warne, 2009), with the added benefits of carbon, water, and trace-compound addition. However, stabilized organic solids are bulky, with nitrogen content between 3 and 10% (dry basis), or 0.3 and 2% on a wet basis. Transport costs are normally in the same cost order of magnitude as the value of the nutrients, meaning that beneficial use is driven by disposal costs, rather than the value of the nutrients. However, there is substantial scope for recovery and concentration of both nitrogen and phosphorus (De-Bashan and Bashan, 2004). Aerobic microbes can be used to accumulate phosphorus via enhanced biological phosphorus removal (EBPR), which results in a high phosphorus WAS stream. Anaerobic digestion plays a key component in phosphorus recovery, as it can be used to re-mobilize ammonia and phosphorus, which can then be recovered as precipitated phosphorus. Struvite ((MgNH4PO4 6H2O) is probably the
638
Anaerobic Processes
best mineral, as magnesium is low cost, and struvite precipitation also allows for nitrogen recovery (Munch and Barr, 2001). At the moment, struvite precipitation is largely used as a phosphorus removal method, than as a recovery technique, but this is likely to change in the future, with anaerobic digestion the major component to mobilize and recover accumulated phosphorus and nitrogen.
4.17.4.3 Future Applications in Energy Generation and Transport Until now, anaerobic digestion has been mainly industrially applied in developed nations for large-scale organic solid stabilization and destruction. Economics are normally driven by solid destruction and stabilization rather than energy production. Energy is either utilized to produce low-quality heat, or in co-generation engines. Economies of scale and maintenance requirements mean that the optimal economic size of co-generation engines is approximately 500 kW. While newer technologies such as microturbines are making anaerobic digestion more attractive at smaller scale, for electricity production, anaerobic digestion is still a large-scale proposition. It is also widely applied at very small scales across Asia, South America, and Africa and the Middle East directly for methane generation and utilization. Methane is used directly and effectively as a natural-gas replacement. In fact, anaerobic digestion is one of the only renewable energy technologies which is fully mature, completely scalable, and generates an energy product that can be stored as produced. There is a particular application in intensive agriculture and food processing, where there is a need for water treatment and energy, and much of the organics are being emitted as methane, which is currently lost (with a consequent greenhouse-gas impact). It is very likely that anaerobic digestion will be implemented increasingly at smaller scale once the technology is standardized further as a partner to other scalable renewable options such as wind and photovoltaic solar cells. The applications of small-to-medium-scale anaerobic digestion are not only limited to methanogenesis. Fermentation can also be used to produce a number of alternative products, including organic acids, alcohols, and hydrogen. While we do not yet have the knowledge to fully direct mixed-culture fermentation to specific products (Rodrı´guez et al., 2006), this is an exciting research area that will likely challenge classical pure-culture fermentation on cost, conversion, and specificity. It has been further enhanced by the application of bioelectrochemical systems, which have the capacity to utilize electrical current to drive full conversion of low-value carbon feedstocks to valuable products such as hydrogen, organic acids, and alcohols.
References Angelidaki I, Alves M, Bolzonella D, et al. (2009) Defining the biomethane potential (BMP) of solid organic wastes and energy crops: A proposed protocol for batch assays. Water Science and Technology 59: 927--934. Aquino SF, Chernicharo CAL, Soares H, Takemoto SY, and Vazoller RF (2008) Methodologies for determining the bioavailability and biodegradability of sludges. Environmental Technology 29: 855--862.
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Ramsay IR and Pullammanappallil PC (2005) Full-scale application of a dynamic model for high-rate anaerobic wastewater treatment systems. Journal of Environmental Engineering 131: 1030--1036. Rao AG, Reddy TSK, Prakash SS, et al. (2008) Biomethanation of poultry litter leachate in UASB reactor coupled with ammonia stripper for enhancement of overall performance. Bioresource Technology 99: 8679--8684. Raposo F, Banks CJ, Siegert I, Heaven S, and Boria R (2006) Influence of inoculum to substrate ratio on the biochemical methane potential of maize in batch tests. Process Biochemistry 41: 1444--1450. Ratledge C (1994) Biodegradation of oils, fats and fatty acids. In: Ratledge C (ed.) Biochemistry of Microbial Degredation, vol. 1, 590pp. Dordrecht: Kluwer. Rebac S, Ruskova J, Gerbens S, vanLier JB, Stams AJM, and Lettinga G (1995) Highrate anaerobic treatment of wastewater under psychrophilic conditions. Journal of Fermentation and Bioengineering 80: 499--506. Reguera G, McCarthy KD, Mehta T, Nicoll JS, Tuominen MT, and Lovley DR (2005) Extracellular electron transfer via microbial nanowires. Nature 435: 1098--1101. Ren N, Wan B, and JuChang H (1997) Ethanol-type fermentation from carbohydrate in high rate acidogenic reactor. Biotechnology and Bioengineering 54: 428--433. Rodrı´guez J, Kleerebezem R, Lema JM, and van Loosdrecht MCM (2006) Modeling product formation in anaerobic mixed culture fermentations. Biotechnology and Bioengineering 93: 592--606. Roy F, Albagnac G, and Samain E (1985) Influence of calcium addition on growth of highly purified syntrophic cultures degrading long-chain fatty acids. Applied Environmental Microbiology 49: 702--705. Rozzi A and Remigi E (2004) Methods of assessing microbial activity and inhibition under anaerobic conditions: A literature review. Reviews in Environmental Science and Bio/Technology 3: 93--115. Schink B (1997) Energetics of syntrophic cooperation in methanogenic degradation. Microbiology and Molecular Biology Reviews 61: 262--280. Seghezzo L, Zeeman G, van Lier JB, Hamelers HVM, and Lettinga G (1998) A review: The anaerobic treatment of sewage in UASB and EGSB reactors. Bioresource Technology 65: 175--190. Siegrist H, Renggli D, and Gujer W (1993) Mathematical modelling of anaerobic mesophilic sewage sludge treatment. Water Science and Technology 27: 25--36. Siegrist H, Vogt D, Garcia-Heras J, and Gujer W (2002) Mathematical model for meso and thermophilic anaerobic sewage sludge digestion. Environmental Science and Technology 36: 1113--1123. Speece RE (2008) Anaerobic Biotechnology and Odor/Corrosion Control for Municipalities and Industries. Nashville, TN: Archae Press. Stams AJM (1994) Metabolic Interactions between anaerobic-bacteria in methanogenic environments. Antonie Van Leeuwenhoek International Journal of General and Molecular Microbiology 66: 271--294. Stams AJM and Plugge CM (1994) Occurrence and function of the acetyl-CoA cleavage pathway in a syntrophic propionate oxidising bacterium. In: Drake HL (ed.) Acetogenesis, pp. 557--630. New York: Chapman and Hall. Steyer JP, Bernard O, Batstone DJ, and Angelidaki I (2006) Lessons learnt from 15 years of ICA in anaerobic digesters. Water Science and Technology 53: 25--33. Stumm W and Morgan JJ (1996) Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters. New York: Wiley. Tchobanoglous G, Burton F, and Stensel H (2003) Metcalf and Eddy Inc. Wastewater Engineering, Treatment and Reuse. New York, NY: McGraw-Hill. Temudo MF, Muyzer G, Kleerebezem R, and van Loosdrecht MCM (2008) Diversity of microbial communities in open mixed culture fermentations: Impact of the pH and carbon source. Applied Microbiology and Biotechnology 80: 1121--1130. Thiele JH and Zeikus JG (1988) Interactions between hydrogen and formate producing bacteria and methanogens during anaerobic digestion. In: Erickson CE and DanielYee-Chak-Fung (eds.) Handbook on Anaerobic Fermentations, pp. 537--595. New York, NY: Dekker. Tong Z and McCarty P (1991) Microbial hydrolysis of lignocellulosic materials. In: Isaacson R (ed.) Methane from Community Wastes, pp. 61--100. London: Elsevier. Van Langerak E and Hamelers H (1997) Influent calcium removal by crystallization reusing anaerobic effluent alkalinity. Water Science Technology 36: 341--348. Van Lier JB (2008) High-rate anaerobic wastewater treatment: Diversifying from endof-the-pipe treatment to resource-oriented conversion techniques. Water Science and Technology 57: 1137--1148. Van Lier JB, Sanz Martin JL, and Lettinga G (1996) Effect of temperature on the anaerobic thermophilic conversion of volatile fatty acids by dispersed and granular sludge. Water Resources 30(1): 199--207. Vavilin VA, Rytov SV, and Lokshina LYa (1996) A description of hydrolysis kinetics in anaerobic degradation of particulate organic matter. Bioresource Technology 56: 229--237.
4.18 Microbial Fuel Cells B Virdis, S Freguia, RA Rozendal, K Rabaey, Z Yuan, and J Keller, The University of Queensland, Brisbane, QLD, Australia & 2011 Elsevier B.V. All rights reserved.
4.18.1 4.18.1.1 4.18.1.2 4.18.1.3 4.18.2 4.18.3 4.18.4 4.18.4.1 4.18.4.2 4.18.4.3 4.18.4.4 4.18.5 4.18.5.1 4.18.5.2 4.18.5.3 4.18.5.4 4.18.6 4.18.7 4.18.8 4.18.9 4.18.9.1 4.18.9.2 4.18.9.3 4.18.9.4 4.18.10 4.18.10.1 4.18.10.2 4.18.10.3 4.18.10.4 4.18.10.5 4.18.11 References
Resource Recovery from Wastewater Water Recovery Nutrient Recovery Energy Recovery Microbial Fuel Cells Thermodynamics of Microbial Fuel Cells Factors Determining the Decrease of Cell Voltage Losses due to Mass-Transfer Limitation Losses due to Bacterial Metabolic Kinetics Losses due to Electron Transfer to (and from) the Electrode Losses due to the Resistance of the Electrolytes (Including the Ion-Exchange Membrane) and of the Electrical Interconnection to the Charges Flow Materials and Architectures Design Compartment Separation Electrodes Cathodic Compartment Electrochemically Active Microorganisms and Extracellular Electron Transfer Oxidative Processes Reductive Processes Challenges toward Improving MFC Efficiency Minimizing Electrode-Potential Losses Respiration, Fermentation, and Methanogenesis Reducing pH Gradients Wastewater and Electrode Resistance Opportunities for Bioelectrochemical Systems Wastewater Treatment Nitrogen Removal Bioremediation H2 Production Bioelectrochemical Production of Value-Added Chemicals Outlook
4.18.1 Resource Recovery from Wastewater Fossil fuel exploitation has significantly affected the economic growth of developed countries within the past century. The world energy-consumption rate is projected to double from 13.5 TW (1 TW ¼ 1012 W) in 2001 to 27 TW by the year 2050 and to triple to 43 TW by 2100 (Lewis and Nocera, 2006). Although the rise in prices of liquid fuels (e.g., crude oil, natural gas plant liquid, biofuels, oil shale, and bitumen) and natural gas is expected to rationalize energy demand, world energy consumption is still projected to increase due to continuing rapid economic growth and expanding population, particularly in the developing countries. Fossil fuels are predicted to remain the dominant sources of primary energy, accounting for close to 83% of the overall increase in energy demand between 2004 and 2030 (Figure 1). Increasing awareness of the possible anthropogenic effects on climate change, in combination with the instability of the
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fossil fuel market, is motivating political consciousness to reduce greenhouse-gas emissions and to promote renewable energy. The greatest challenge in the future lies in catering to the world’s growing energy demands while simultaneously reducing emission of greenhouse gases. This is certainly predicted to provide serious challenges for fossil-fuel-based economies (Logan, 2008). Nuclear fission alone does not represent a feasible alternative, as known uranium reserves would be depleted within a few decades, not considering the environmental damage caused by the mining and disposal of radioactive material (Lewis and Nocera, 2006). Solar energy is an attractive energy source as it is both renewable and available in large amounts (Seboldt, 2004); however, a society completely dependent on solar energy is not realistic for the short term due to technological and economical difficulties. Other renewable energy technologies must be developed in conjunction with solar energy. About 200 TW of the 170 000 TW solar-radiation flux is continuously transformed
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Microbial Fuel Cells 250 History
Quadrillion kJ
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Projection
Oil Coal Natural gas Renewable Nuclear
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100
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0 1990
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Figure 1 World energy use detailed by fuel type, 1990–2030. Data from Energy Information Administration (EIA).
into wind power, whereas 100 TW is stored as biomass through photosynthesis, and 6 TW is transformed into hydropower through the water cycle (Niele, 2005). These indirect forms of solar energy are already exploited to some extent for electricity production through wind turbines, biomass gasification/combustion, and hydroelectric dams. Nevertheless, the extent of this exploitation must be further developed in the future in parallel to elevating societal energy demands. Securing water resources to match increasing demand is also becoming challenging. Global trends such as urbanization and migration, often combined with frequent drought periods, even in areas traditionally rich in water resources, have increased the demand for water, food, and energy, and put at risk the sustainability of current living standards. The pressure of water availability has also affected other waterconsuming sectors, such as public water supply, agriculture, industry, and, of course, power generation. In this global picture of fading resources, we can therefore no longer afford to waste any potential sources of all these three key resources. For example, domestic and industrial wastewaters are ubiquitous and represent a potential source of energy, water, and nutrients. The development of technologies capable of simultaneously recovering energy, water, and nutrients from wastewater is crucial to resource management in the future.
4.18.1.1 Water Recovery Wastewater represents a valuable recyclable water resource. Although containing compounds dangerous to public health and to the environment (e.g., pathogens, chemicals, organics, and nutrients), at least 99.9% of wastewater is in fact water and, as such, it should by no means be considered as waste. Engineered technologies for the reintroduction of treated wastewater to water-supply grids appear to be an essential
priority, given the increasingly limited water resources in both quantity and quality. Wastewater treatment plants are a crucial part of the overall water-recycle process, being an important pre-treatment step for the advanced treatment processes, which are currently almost nonexistent worldwide, but which can generate water qualities suitable for reuse even in potable water applications. Wastewater-treatment processes aim to reduce the relevant concentration of pollutants by means of separation, destruction, and disinfection (Tchobanoglous et al., 2003). The efforts into improving the quality of wastewater-treatmentplant effluents have achieved levels of pollutant elimination well beyond the standards of environmental protection. An activated sludge treatment can reduce the influent biological oxygen demand (BOD) concentration from 4300 mg l1 to o5 mg l1 when upgraded for biological nutrient removal, while reducing the influent total nitrogen (N) concentration from 460 mg l1 to o3 mg l1; and influent phosphorus (P) concentrations from 412 mg l1 to o1 mg l1. (BOD is a measure of the concentration of biodegradable material present in wastewater expressed as the amount of oxygen consumed by microorganisms in breaking down the organic matter during a certain period of time. It normally represents a fraction of the chemical oxygen demand (COD), which is the total oxygen consumption consumed during chemical breakdown of organic and inorganic matter.) Further disinfection treatment can achieve up to 99.9999% removal of bacteria where membrane filtration is used (Foley and Keller, 2008). Advanced water treatments (AWTs) can further improve the efficacy of the disinfection process, by removing recalcitrant organics that are not metabolized in the biological nutrientremoval process and by reducing the content of total dissolved solids. AWT consists of a multi-barrier system against various acute and chronic risk factors, such as micropollutants and pathogens, that remain even after regular wastewater
Microbial Fuel Cells
treatment, and prior to the addition to ground or surface water for reuse. The most common forms of AWT are microfiltration and reverse osmosis, which can be followed by advanced oxidation processes (ozone and H2O2/ultraviolet (UV)) to remove recalcitrant contaminants. If an energy-recovery process is also included, the wastewater can be, for example, initially treated through anaerobic digestion, which would produce biogas that can power gas turbines for electricity generation. The effluent would then require further aerobic polishing to remove the slower biodegradable material, while achieving drinking water standards would require AWT. The water can subsequently be collected in the environment (e.g., in a dam) where time and environmental buffers will ensure that the higher-quality standards are met, even prior to any treatment processes already in place to produce safe drinking water. In this scheme, a large fraction of the wastewater is thus recovered as clean water to be reused for domestic, agricultural, and industrial purposes.
4.18.1.2 Nutrient Recovery For many years, wastewater treatment methods have been improved to achieve environmental protection from nutrient overload in receiving water bodies. Removal of carbon, nitrogen, and phosphorus from wastewater requires large amounts of energy and produces potentially useful resources of minerals and water that are normally disposed of. The most important minerals for living organisms are considered to be nitrogen and phosphorus, although potassium and sulfur should also be included as essential. Nitrogen and phosphorus for agriculture are produced from natural resources. Phosphorus is currently entirely derived from highly geographically concentrated geological reserves (mainly in North Africa, USA, China, and Russia). Phosphate rocks are finite nonrenewable resources and are therefore limited, with an estimated 50–100 years until depletion is reached under current extraction rates (Larsen et al., 2007). Moreover, extraction and production of good-quality phosphorus require energy, and produces a waste (the production of 1 kg of phosphorus produces up to 2 kg of gypsum, inclusive of heavy metals and radioactive elements). It is thus essential that phosphate is recovered efficiently in the future. Recycling of phosphorus contained in sewage is currently very limited, even though several techniques are available to incorporate it into the excess activated sludge. Nitrogen in fertilizers is almost completely supplied by atmospheric N2, which makes the source virtually infinite. However, to be accessible to living organisms, atmospheric nitrogen has to be in the form of ammonia or nitrate. Industrial processes for the conversion of N2 gas to ammonia require a large energy investment using the Haber–Bosch process (10.3 kW h1 kg1 nitrogen produced; Maurer et al., 2003). Additional energy is invested to obtain the opposite process during wastewater-treatment processes to achieve low nitrogen levels in the treated effluent. With this in consideration, wastewater treatment for nitrogen removal can be regarded as an indirect and inefficient method for nitrogen recovery (recycling over the atmosphere) since it engineers biological nitrification and denitrification to N2 gas that is
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returned to the atmosphere. Therefore, any direct nitrogen recovery process will have to be more energy efficient than the indirect route via the atmosphere to be environmental favorable. As pointed out by Maurer et al. (2003), direct recovery of nitrogen-rich wastewater into reusable forms can be more sustainable than indirect recovery by biological nitrification/ denitrification and ammonia production through the Haber– Bosch process, as the total net amount of energy required is lower. For instance, the energy demand for nitrification and denitrification, together with ammonia production through the Haber–Bosch process, would be 42.8 kW h1 kg1 nitrogen using a denitrification process with methanol addition; 25 kW h1 kg1 nitrogen in case of a preanoxic denitrification system; and 17.8 kW h1 kg1 N for nitrogen removal through the Sharon–Anammox process (Maurer et al., 2003). This considerable energy requirement for indirect nitrogen recycling makes some direct recovery techniques such as thermal volume reduction of urine (requiring about 8.1–9.4 kW h1 kg1 N), or even struvite production, economically and environmentally interesting (Maurer et al., 2003). (Struvite is a phosphate mineral with formula NH4MgPO4 6H2O. Its production requires 28.3 kW h1 kg1 N (Maurer et al., 2003), which is higher than the energy demand for alternative N recovery processes like thermal volume reduction of urine. However, together with nitrogen, phosphorus is also recovered (struvite contains about 2.2 kg phosphorus per each kilogram of nitrogen).)
4.18.1.3 Energy Recovery Many recovery processes can provide bioenergy or valuable chemicals from relatively concentrated biomass streams, such as from wood and agricultural by-products (Hatti-Kaul et al., 2007; Petrus and Noordermeer, 2006; Ragauskas et al., 2006; van Wyk, 2001). Yet, not many conversion processes exist for energy and chemical production from diluted aqueous streams, such as industrial, agricultural, and municipal wastewater. Wastewater contains significant amounts of renewable energy in the form of chemical bonds. For example, domestic wastewater could potentially yield energy up to 2.2 kW h1 m3. (This is considering the energy content of glucose as 4.4 kW h1 kg1 COD, and a wastewater with 500 mg COD l1.) If properly recovered, the chemical energy daily wasted with sewage can potentially cover up to 7% of the energy consumption used for residential purposes in developed countries. (This is assuming an energy recovery of 1.2 kW h1 kg1 COD and considering a total residential energy consumption of 649.8 kg of oil equivalent capita1, equal to 7556 kW h1 capita1, and a total water withdrawal of 948 m3 yr1 capita1, assuming that it all ends up in sewage. Energy-consumption data include all energy used for activities by households except for transportation. Data on energy and water withdrawal are available at the World Research Institute.) This figure is expected to be much higher in developing countries, since the energy consumption tends to be far lower. Even though technologies such as anaerobic digestion have been long known and implemented for many years to recover energy from wastewater, the activated sludge process is by far the most widely applied process for wastewater treatment. The
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process relies upon aeration of the wastewater, which allows microorganisms to convert the available organics into carbon dioxide. The solids are then separated from the treated water through sedimentation. Although this process yields valuable purified water, the high energy requirements for aeration, typically 0.5 kW h1 m3 treated water (Rabaey and Verstraete, 2005), makes anaerobic processes far more energy efficient than aerobic treatment. Methanogenic anaerobic digestion of organics has been shown to be advantageous over aerobic processes due to its high organic removal rates, low sludge production, low energy inputs, and, ultimately, for its energy production (Angenent et al., 2004). The methane produced thus has traditionally been used as the on-site fuel source for heat/electrical applications, or used to power gas turbines, with net energy efficiencies up to 35–40%. More recently, methane has also been converted into other products by catalytic conversion to syngas, a mixture of hydrogen and carbon monoxide, or into methanol for use in production of biodiesel (Angenent et al., 2004). However, whether methane is used in a gas turbine or to produce syngas, it is necessary to purify the biogas from impurities such as hydrogen sulfide also produced during anaerobic digestion, which therefore equates to an additional treatment process. Dark fermentation represents an alternative to biological methane production, which shares with it much of the same process reactions involved, except that during dark fermentation, the hydrogen-metabolizing organisms (methanogens) are inhibited through heat treatment of the initial inoculum while retaining only spore-forming fermenting bacteria in which hydrogen-forming bacteria are included. However, due to the limitations imposed by the thermodynamics of hydrogen formation through the hydrogenase reaction, the conversion yields of the total electron equivalents present as carbohydrate in wastewater does not normally exceed B15% (Angenent et al., 2004). As such, the process appears less appealing in comparison with the more reliable and mature biological methane production. Microbial fuel cells (MFCs) have been gaining increasing attention in recent times as devices able to produce electric power while simultaneously treating industrial, agricultural, or municipal wastewater (Rozendal et al., 2008a). Compared to treatment technologies, MFCs have the advantage of being able to theoretically achieve efficiencies. The underlining rationale is that fuel cells do not use heat as an intermediate form of energy for electricity production. As such, the process efficiency is not limited by the Carnot cycle, according to which, for a reversible process, the theoretical maximum conversion efficiency of heat to work is determined by the absolute temperature Th (K) of the process and the absolute temperature Tc (K) of the cold sink (i.e., the environment):
Zideal ¼ 1
Tc Th
ð1Þ
As heat-resisting properties of construction materials are limited to a certain maximal temperature, the theorem implies that the overall yield of combustion processes is usually no higher than 35–45% (Carnot, 1824). Since fuel cells do not operate on a thermal cycle, they are not constrained to thermodynamic limitations such as the Carnot’s theorem.
Therefore, they can theoretically convert the entire free energy of the fuel oxidation into electric energy (Schroder and Harnisch, 2009).
4.18.2 Microbial Fuel Cells Although the existence of a bioelectrical phenomenon was first observed by Italian physicist Luigi Galvani in 1790 (Piccolino, 1997), the principle that microorganisms could generate voltage and current was put forth by Michael Cresse Potter, a professor of botany at the University of Durham, UK, at the beginning of the twentieth century (Potter, 1911). This occurred a few years earlier than the discovery of the activated sludge process and shortly after the invention of the Imhoff tank, an early form of anaerobic digester. In 1931, Barnett Cohen confirmed Potter’s observations reporting a stacked biological fuel cell delivering 35 V at a current of 2 mA (Cohen, 1931). However, it was not until the US National Aeronautics and Space Administration (NASA) became interested in exploiting opportunities for recycling organic wastes into electricity during long space flights that MFCs regained popularity, and by the year 1963, were already commercially available as a power supply for small electrical devices (Shukla et al., 2004). Despite these early successes, the rapid advancement of alternative technologies, such as solar photovoltaic systems, and the fact that the complexity of the underlying biochemical processes became more evident, MFCs suffered an inevitable setback. However, the growing awareness to reduce society’s dependency on fossil fuel and the emerging environmental consequences of their usage has triggered the revival of MFC research in the last 10–15 years. In an MFC, the chemical energy contained in soluble organic molecules, such as carbohydrates and volatile fatty acids (VFAs), can be directly recovered as electric energy. MFCs are galvanic cells that couple the oxidation of an electron donor at an anode with the reduction of an electron acceptor at a higher redox potential at the cathode. Power output is generated as the overall reaction is exergonic. MFCs are the most extensively described bioelectrochemical system (BES) which, more generically, refers to a device where microorganisms interact electrically with electrodes (Rabaey et al., 2007). Microbial electrolysis cells (MECs) are another category of bioelectrochemical systems where the oxidation reaction at the anode is coupled to the reduction of an electron acceptor at a lower potential at the cathode (i.e., water to produce hydrogen). Since the process is endergonic, a certain voltage needs to be applied. In its standard configuration, an MFC consists of two chambers: the anode and the cathode compartments (Figure 2). Bacteria growing at the anode catalyze the electron transfer from an organic (or inorganic) molecule to the anodic electrode. The reduction of the terminal electron acceptor takes place at the cathode, generally separated from the anode by an ion-selective membrane and electrically connected to it via an external circuit containing a resistor or power user that harvests the energy liberated by the reactions. Several electron acceptors can be used, for example, oxygen (O2), potassium hexacyanoferrate (also known as ferricyanide, K3Fe(CN)6), and nitrate, (NO3 ). The cathodic reaction can be of an
Microbial Fuel Cells
e-
A
The maximal work that can be derived from such processes can be measured by means of the Gibbs free energy of the general redox reaction nA A þ nB B- nC C þ nD D:
e-
H2O
CO2
O2
COD
Figure 2 Schematic representation of a microbial fuel cell (MFC). The substances (organics represented as chemical oxygen demand (COD)) are oxidized to CO2 by microorganisms, which transfer the gained electrons to the anode. At the cathode, the electrons are used to reduce oxygen abiotically or biotically, producing water. To maintain electroneutrality within the system, positive charges have to migrate from the anode to the cathode through an ion-permeable separator (a cation exchange membrane (CEM) in this representation).
electrochemical or bioelectrochemical nature. In the latter case, bacteria are involved as catalysts at the cathode as well, promoting electron transfer from the electrode to the final electron acceptor. If glucose is taken as an example of electron donor and oxygen as electron acceptor, Equations (2) and (3) characterize the reactions occurring at the anode and cathode, respectively:
C6 H12 O6 þ 6H2 O- 6CO2 þ 24H þ 24e
ð2Þ
6O2 þ 24Hþ þ 24eþ - 12H2 O
ð3Þ
þ
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The aqueous media in the anodic and cathodic compartments are called the anolyte and the catholyte, respectively. The role of the ion-exchange membrane (IEM) is to allow the transport of charges between the two compartments, thus maintaining the electroneutrality of the system, and also to physically separate the two redox processes, thus preventing the electron acceptor from reacting directly with the electron donor. (IEM is a type of membrane that allows the selective diffusion of certain ions. Different types of IEMs exist, depending on the species that is transported, including cationexchange membranes (CEMs), proton-exchange membranes (PEMs), and anion-exchange membranes (AEMs).)
4.18.3 Thermodynamics of Microbial Fuel Cells Chemotrophic organisms fulfill their energy requirements by transferring electrons from a low redox potential molecule (primary electron donor) to a high redox potential molecule (primary electron acceptor). MFC electrodes virtually interpose within the electron-transfer process that would naturally occur in bacteria between the electron donor and acceptor.
DGr ¼ DGr0 þ RTln
anCC anDD anAA anBB
ð4Þ
where DGr is the Gibbs free energy of a reaction at specific conditions, measured in Joules (J), DG0r (J) is the Gibbs free energy at standard conditions (usually defined as 298.15 K, 1 bar pressure, and 1 M concentration of the species), R is the universal gas constant (8.3145 J mol1 K1), T is the absolute temperature (K), and ai is the activity of reactant i, and ni the respective stoichiometric coefficient. (The Gibbs free energy represents the maximum amount of useful work that can be obtained from a reaction.) In diluted systems, the relation can be simplified by replacing the activities with the concentrations, and Equation (4) can be rewritten as
DGr ¼ DGr0 þ RTln
nC ½C ½D nD ½A nA ½B nB
ð5Þ
In order to generate a current, the overall process in an MFC needs to be thermodynamically spontaneous. This requires the Gibbs free-energy change of the process to be negative. For a bioelectrochemical conversion, it is useful to evaluate the reaction in terms of electromotive force (Eemf), which is expressed in volts (V). The electromotive force and the Gibbs free energy are related according to
DGr ¼ QEemf ¼ nFEemf
ð6Þ
where Q is the charge transferred in the reaction in coulombs (C), which is also equal to the number n of electrons exchanged in the reaction (mol) multiplied per the Faraday’s constant F (9.64853 104 C mol1). Equation (6) can therefore be rearranged, yielding
Eemf ¼
DGr nF
ð7Þ
At standard conditions, DGr is equal to DG0r , and Equation (7) can be written as
E0emf ¼
DGr0 nF
ð8Þ
where E0emf represents the electromotive force at standard conditions. Equations (4) and (8) can be combined to calculate the total electromotive force for a given redox reaction occurring at certain conditions, yielding Equation (9), which is known as the Nernst law:
Eemf ¼ E0emf
nC RT ½C ½D nD ln ½A nA ½B nB nF
ð9Þ
Positive values for Eemf refer to spontaneous processes, whereas negative values indicate a nonspontaneous reaction. MFC technology is a galvanic process characterized by positive
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values for the electromotive force. Microbial electrolysis is instead an electrolytic process where the electromotive force assumes negative values. The Eemf in an MFC can be evaluated by considering the generic redox reaction occurring as the sum of an oxidation and of a reduction. The Eemf is the result of the difference between the reduction potential of the reactions occurring at the cathode and at the anode (according to Equation (6)), each of them evaluated through the Nernst equation applied to the half reaction. Half-cell potentials are reported under the International Union of Pure and Applied Chemistry (IUPAC) convention as reduction potentials in comparison with the standard hydrogen electrode (which has a reduction potential conventionally set to zero at pH2 ¼1 bar, [Hþ] ¼ 1 M); therefore, the reaction is always written as an electron-consuming reaction (reduction). Table 1 lists a series of half-reaction reduction potentials important in MFCs and biological systems in general (Thauer et al., 1977). For biological purposes, the redox potentials are
Table 1 systemse
Summary of redox reactions important in biological
EAn ¼ E0An
ECat ¼
RT ½C6 H12 O6 ln 24F ½CO2 6 ½Hþ 24
E0Cat
RT 1 ln 4F pO2 ½Hþ 4
!
E0 0
6CO2 þ 24Hþ þ 24e-Glucose þ 6H2O 2Hþ þ 2e-H2 NADþ þ Hþ þ 2e-NADH 2CO2 þ 8Hþ 8e-Acetate þ 2H2O S þ 2Hþ þ 2e-H2S SO4 2 þ 10Hþ þ 8e - H2 S þ 4H2 O Pyruvate þ 2Hþ þ 2e-Lactate FADþ þ 2Hþ þ 2e-FADH2 Fumarate2 þ 2Hþ þ 2e-Succinate2 Cytochrome b (Fe3þ) þ e-Cytochrome b (Fe2þ) Ubiquinone þ 2Hþ þ 2e-Ubiquinone H2 Cytochrome c (Fe3þ) þ e-Cytochrome c (Fe2þ) NO2 þ 2Hþ þ e - NO þ H2 O FeðCNÞ6 3 þ e - FeðCNÞ6 4 Cytochrome a (Fe3þ) þ e-Cytochrome a (Fe2þ) NO3 þ 2Hþ þ 2e - NO2 þ H2 O NO2 þ 8Hþ þ 6e - NH4 þ þ 2H2 O NO3 þ 6Hþ þ 5e - 0:5N2 þ 3H2 O Fe3þ þ e-Fe2þ O2 þ 4Hþ þ 4e-2H2O NO þ Hþ þ e-0.5NO þ 0.5H2O 0.5N2O þ Hþ þ e-0.5N2 þ 0.5H2O
0.43 Va 0.42 Va 0.32 Va 0.28 Va 0.28 Va 0.22 Va 0.19 Va 0.180 Vd þ 0.03 Va þ 0.035 Va þ 0.11 Va þ 0.25 Va þ 0.350 Vb þ 0.36 Vc þ 0.39 Va þ 0.433 Vb þ 0.440 Vd þ 0.74 Va þ 0.76 (pH ¼ 2)a þ 0.82 Va þ 1.175 Vb þ 1.355 Vb
From Madigan MT, Martinko J, and Parker J (2000) Brock Biology of Microorganisms. Upper Saddle River, NJ: Prentice Hall. b From Thauer RK, Jungermann K, and Decker K (1977) Energy-conservation in chemotropic anaerobic bacteria. Bacteriological Reviews 41: 100–180. c From He Z and Angenent LT (2006) Application of bacterial biocathodes in microbial fuel cells. Electroanalysis 18: 2009–2015. d From Rabaey K and Verstraete W (2005) Microbial fuel cells: Novel biotechnology for energy generation. Trends in Biotechnology 23(6): 291–298. e The standard redox potentials are measured at pH 7 and 25 1C. Redox couples are arranged from the strongest oxidant (more positive reduction potential) at the bottom, to the strongest reductants (most negative reduction potential) at the top. Electrons naturally flow from lower to higher redox potentials. The larger the difference in reduction potential between electron donor and electron acceptor, the larger is the energy released.
ð10Þ
! ð11Þ
The difference between ECat and EAn would then give
Eemf ¼ ECat EAn
Redox reaction
a
generally referred to at pH 7 and 25 1C (in which case they are indicated with the symbol E0 0). These reactions include not only oxidations of organics and reductions of terminal electron acceptors, but also redox reactions of intermediate metabolites. Based on the values in Table 1, if glucose is the electron donor (–0.43 V) and oxygen is the electron acceptor ( þ 0.82 V), an electromotive force of 1.25 V would develop across the MFC at standard conditions. Under more general conditions, the application of the Nernst law on Equations (2) and (3) would yield the following potentials for the two half-cell reactions:
ð12Þ
4.18.4 Factors Determining the Decrease of Cell Voltage Although the electromotive force represents the upper limit for the total voltage that the MFC can generate under certain conditions, the actual voltage will always be lower under practical conditions, due to a number of losses of either purely electrochemical and/or of biological nature. An ideal MFC would deliver any amount of current while maintaining a constant voltage, as determined by thermodynamics. In practice, the actual voltage output would be lower due to irreversible losses. These potential losses increase with increasing currents and can have a dramatic effect on the performance of the MFC, as the loss of voltage would result in a lower power output, accordingly to Equation (13)
P ¼ Vi
ð13Þ
where P is the power density (W cm2), V is the voltage (V), and i is the current density (A cm2). (In order to permit the comparison between different systems, current and power are usually normalized to some characteristic of the reactor, such as the projected surface area of anode or cathode, or alternatively to the compartment total volume or liquid (net) volume.) Maintaining a high voltage under high current production is therefore critical for successful MFC operations. Polarization curves represent the cell’s voltage as a function of the current. They are regarded as a useful tool for the measurement of the MFC performance (Figure 3). They are performed by periodically modifying the applied load (external resistance) and recording the resulting voltage and current, the latter evaluated through Ohm’s law (V ¼ R i). They can be performed manually or automatically by means
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647
Theoretical Eemf Voltage at open circuit (OCV)
Cell voltage (V)
P-i curve (b) Activation region
Ohmic region
Mass transport region
Vopt
Current density (A.cm−2)
iopt
Power density (w.cm−2)
Pmax
V-i curve (a)
isc
Figure 3 Typical polarization (a) and power (b) curves for an MFC. The point of maximal power (Pmax) corresponds to the optimal voltage (Vopt) and the optimal current density (iopt.). The maximal current density at short circuit (iSC) is reached when the external resistance is zero.
of a potentiostat. In this case, an appropriate scan rate (e.g., 0.1 mV s1) should be chosen (Velasquez-Orta et al., 2009). Polarization curves should be recorded from high to low external load and vice versa. While the Eemf as defined earlier represents the thermodynamic potential difference achievable in an electrochemical system, its value is not normally reached in real systems. The open circuit voltage (OCV) is the maximal voltage that can in fact be measured under conditions at which there is infinite resistance (i.e., at open circuit). There is a series of limitations imposed by the specific bacterial communities catalyzing the anodic reaction (and cathodic, in case of biocathode) that reduce the overall potential difference attainable (Logan, 2008). Three zones defining as many different operating regimes can be identified in a polarization curve (Benziger et al., 2006): 1. At open circuit there is no flow of electric current. However, when the current starts flowing, the voltage drops rapidly as a result of the activation-energy barrier of the reactions occurring at the electrodes; this zone is referred to as ‘activation polarization region’. (The voltage at open circuit measures the activity of reactants at anode and cathode electrode surfaces.) 2. At medium currents, the voltage decreases almost linearly with the current; this is referred to as ‘ohmic polarization region’, as it is dominated by ohmic losses, which arise from the resistance opposed by electrolytes and the IEM to the transport of ions as well as by electrodes and interconnection circuit to the transport of electrons. 3. At higher currents, the voltage drastically drops as a result of the insufficient mass transport of reactants or reaction products to and from the electrode, which limits the
reaction. This is known as ‘concentration polarization region’. The ratio of the cell voltage (V) and the cell voltage at open circuit (OCV) gives the potential efficiency (PE) (Lee et al., 2008). It is essentially the portion of the total potential difference between electron donor and acceptor that is captured as useful electric energy. The coulombic efficiency, or charge transfer efficiency (CE, or eC), is defined as the ratio of the charge that is transferred to the anode and the maximal charge that would be yielded if all the converted substrate generate electricity (Logan et al., 2006). It represents therefore the fraction of electrons recovered as electricity from the substrate converted. The energy conversion efficiency (ECE, or eE) is obtained by multiplying the PE and the CE (Lee et al., 2008). It represents the ratio of the power delivered from the system and the power that would be delivered in the absence of internal resistances (Benziger et al., 2006). For an MFC, the objective is to maximize the power output (represented by the peak of the P vs. i curve in Figure 3) and the ECE. Maximal power is obtained when both current and voltage are maximized, whereas maximal ECE is obtained when the potential efficiency and the coulombic efficiency are both maximized. However, the potential efficiency is negatively affected by the current density, which is in turn needed to maximize the power. This aspect is very important in engineering MFC systems (and fuel cells systems in general) as it means in other words that ECE and power output cannot be simultaneously optimized. At maximal power output, the ECE is 50% (Benziger et al., 2006). Higher efficiencies are achievable but with lower power outputs. Understanding the nature of the losses is of fundamental importance for successful operations of MFCs. Electricity generation in MFC is in fact the result of several steps that
648
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A
CO2
H2O Microorganism
Link
Link
Microorganism O2
Voltage (V)
COD
OCV 1
2 3 4 3 2
Cell voltage
1
Figure 4 Potential losses during electron transfer in an MFC including a bioanode and a biocathode. 1: losses due to mass transfer limitation. 2: losses due to bacterial metabolic kinetics. 3: losses due to electron transfer to/from the electrode. 4: losses due to the resistance of the electrolytes (including the ion exchange membrane) and of the electrical interconnection to the flow of charges. These losses result in the reduction of the cell’s voltage from its value at open circuit (i.e., at infinite external resistance).
necessarily need to proceed at the same rate, as they occur in series. These steps are: (1) mass transfer of reactants and products from the bulk liquid to the electrode attached biofilm, and vice versa, (2) losses due to bacterial metabolic kinetics, (3) electron transfer from microbial cells to the electrode (and vice versa), and (4) transfer of charges through the electrodes and through the electrolyte and IEM. Some of these steps may limit the overall rate of electron transfer thus causing a larger voltage loss than others (Figure 4). The majority of the investigations carried out in the field of MFCs have been aimed at the improvement of power outputs by acting on one of the limiting steps to electricity generation.
4.18.4.1 Losses due to Mass-Transfer Limitation Electricity production in MFC relies on the flux of the reaction reactants in and out the biofilm. The flux of substrates is controlled by the diffusion in the biofilm as well as by the rates of utilization or production. If the reactions involving the substrates occur at a rate that is faster than that at which the reactants or products diffuse in or out of the biofilm, accumulation or depletion of one of the components occurs within the biofilm. As a result, the electrode potential becomes modified as depicted by the Nernst equation (Equation (9)).
If we consider, for instance, the anodic oxidation of glucose (Equation (2)), per mole of glucose that diffuses and is consumed within the biofilm, 6 mol of CO2 and 24 mol Hþ are produced and have to diffuse out of the biofilm. Protons are particularly important as their accumulation may lead the acidification of the biofilm. Torres et al. (2008) have shown that current density is largely determined by proton transport out of the biofilm. Current densities higher by more than 4 times were achieved when the phosphate buffer was increased from 12.5 to 100 mM. The authors also concluded that only in systems in which low COD concentrations are required in the effluent, substrate mass transport limitation may be more important than proton transport. However, it is important that the MFC compartments receive a proper loading rate of substrate to support the biomass attached to the electrodes. In continuous systems, organic loading rates at the anode of MFCs can vary between 0.5 and 4 kg COD m3 of anode liquid volume per day, with an optimum close to 3 kg COD m3 d1 (Rabaey et al., 2003). The hydrodynamic patterns are also important in order to provide homogeneous conditions on the biofilm/liquid interface.
4.18.4.2 Losses due to Bacterial Metabolic Kinetics Bacterial metabolic losses result from the rates of substrate uptake and utilization during the microbial metabolic activity,
Microbial Fuel Cells
which depends on both the specific microbial consortium catalyzing the reactions and the biomass density on the electrode surface, which in turn depends on the specific surface area of the electrode accessible to bacteria. For instance, the main limitation in MFC anodes is often not the specific uptake rate by the bacteria, but the bacterial density at the anode. Measurements have revealed that biomass concentrations at MFC anodes are 30 times lower compared to anaerobic digesters (Aelterman et al., 2008). Improvement of current and power outputs in MFCs requires the achievement of denser microbial colonization of the electrodes, while maintaining thin biofilms and open structures to facilitate diffusion and reduce mass-transfer limitation. While bacteria attach well to graphite electrodes, plain graphite may not be satisfactory if high power outputs are desired. Extensive research has been done on anode materials to maximize surface affinity with microorganisms (Cheng and Logan, 2007, Liu et al., 2007) and to facilitate electron transfer by the immobilization of mediators (Park and Zeikus, 2003) or conductive polymers (Schroder et al., 2003) on the anode surface. However, regardless of the material or design adopted, the anode biology does not currently constitute the main bottleneck of MFCs, unless competing populations such as fermentative bacteria outgrow the anodophilic population, driving the process to a failure (Rabaey et al., 2003). In addition, microorganisms themselves require energy for growth and maintenance purposes. Therefore, an anodic biofilm, for example, would take part of the energy available from the organic substrate and release the electrons at a slightly lower energy level, thus reducing the total voltage. As the anode is virtually the final electron acceptor, its potential would affect the total energy available for the microbes. The higher the difference is between the redox potential of the substrate and the electrode, the higher is the theoretical energy gain for bacteria growing on its surface, per electron-mole transferred. To maximize the voltage, anodic and cathodic electrode potentials should be kept as negative and as positive as possible, respectively, accordingly to the limits imposed by the redox potentials of the substrates used. Nevertheless, when the anode potential becomes very low, competitive processes such as fermentation or even acetoclastic methanogenesis may be favored, as the energy gain would be comparable in that case, as was shown in some recent studies (Aelterman et al., 2008; Finkelstein et al., 2006; Freguia et al., 2007b; Virdis et al., 2009). Furthermore, while higher cathodic potentials would maximize the voltage, the lower driving force pushing electrons from the cathode to the final electron acceptor may lead to the accumulation of intermediates as has been shown in the case of cathodic denitrification (Virdis et al., 2009).
4.18.4.3 Losses due to Electron Transfer to (and from) the Electrode Voltage losses due to electron transfer to (and from) the electrode are caused by the finite rate of electron transfer between microorganisms and the solid phase of the electrode (and vice versa). At an anode, the result is an accumulation of positive charge on the electrode and negative charge in the form of anions in the adjacent liquid layer. This double layer thus established causes the development of a potential
649
difference across it, called activation overpotential, which results in a reduction of cell voltage equal to its value. Bioelectrochemical reactions in BESs differ significantly from conventional electrocatalytic reactions by the fact that while in the latter the electron-transfer step from the electron donor to the electrode proceeds only at one particular point (e.g., the catalyst particle), the oxidation of the substrate in the former occurs throughout a more complex series of enzymatic reactions, the last of which is the electron transfer to the electrode in the case of the anodic reaction. Only this last step influences the activation polarization (Schroder and Harnisch, 2009). Activation overpotentials are extensively described in the electrochemistry literature. A detailed explanation of the origin of overpotentials can be found in Rieger (1994). Activation overpotentials are mathematically described by the Butler–Volmer equation, which dictates the logarithmic increase of the overpotential with the current density:
bFZ ð1bÞFZ i ¼ i0 e RT e RT
ð14Þ
where Z (V) is the overpotential at the electrode, R is the universal gas constant (8.3145 J mol1 K1), T is the absolute temperature (K), b is the symmetry factor (unitless), which is a constant that represents the dependence of the activation energy on the electrode potential, F is the Faraday’s constant (9.648 53 104 C mol1), i is the current density (mA m2), and i0 is exchange current density (mA m2), which depends on the activation energy of the reaction at equilibrium conditions, in such a way that higher activation energy results in lower exchange currents. At overpotentials that are sufficiently high (greater than 80–100 mV at 25 1C, according to Freguia et al. (2007c)), the second term between brackets becomes negligible and Equation (14) can be rewritten in its simplified version, best known as Tafel equation:
ln
i bFZ ¼ i0 RT
ð15Þ
Equation (15) can be used to experimentally estimate the parameters i0 and b using the so-called Tafel plots (ln(i) vs. Z), generated from polarization-curve measurements (Freguia et al., 2007c). The parameters i0 and b (and thus the activation overpotential) strongly depend on the activation energy of the reaction at the electrode. Electrodes with high specific surface area can not only support increased biomass densities but also decrease activation losses by reducing the current densities at the electrode surface (Chaudhuri and Lovley, 2003; Freguia et al., 2007c).
4.18.4.4 Losses due to the Resistance of the Electrolytes (Including the Ion-Exchange Membrane) and of the Electrical Interconnection to the Charges Flow It was described earlier (Section 4.18.2) that an equimolar amount of positive and negative charge is produced during the oxidation reaction at the anode (Equation (2)). While the electrons need to travel along the electrodes and the electrical circuitry to reach the cathode where the reduction reaction
650
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takes place, ions needs to travel through the electrolyte and the IEM to ensure electroneutrality (cations in case a CEM is used; anions in the case of an AEM). The resistance of the different conductors (e.g., electrodes, current collectors, wires, IEM, and electrolyte) toward this charge flow introduces a loss of voltage, which is referred to as ohmic loss. According to Ohm’s law, the voltage loss due to charges migrating through a medium is proportional to the current and to the ohmic resistance of the medium. In turn, the ohmic resistance depends on the medium resistivity r(mO cm) (or equivalently its inverse, the conductivity s, mS cm1), the average distance traveled by the ions (L, cm), and the cross-section area over which the charges move (A, cm2, often identical to the nominal surface area of the electrodes for two-dimensional configurations), according to the following equation:
L R¼r A
ð16Þ
The intrinsic resistance of the electrode could become an important limitation to the generation of electricity as the electric resistivity of the conductors can be very high (the resistivity of graphite and carbon is 1000 times higher than that of iron; Rozendal et al., 2008a). The resistance offered by the electrolytes as well as by the IEM toward this transfer of charge is often a major limitation in MFCs.
4.18.5 Materials and Architectures Although typical MFC architecture consists of an anodic chamber and a cathodic chamber separated by an IEM, as depicted in Figure 2, different materials and reactor configurations have been implemented for lab-scale studies, depending on the scope of the study itself. Designs may vary from two-compartments to single-chambered MFCs, from tubular to stacked configurations, and with or without a membrane. Sediment MFCs have also been constructed by placing one electrode into marine sediments and the other in the overlying oxic water (Reimers et al., 2001; Tender et al., 2002).
4.18.5.1 Design As explained above (Section 4.18.4), the performance of MFC is strongly affected by a number of factors, particularly the resistivity of material used for the electrodes, the resistance offered by the electrolytes toward the charge transport, and the nonperfect selectivity of the IEM, which creates pH gradients between the compartments. It is therefore not surprising that the system performances are dictated by the design and the materials used. The H-shape two-chambered design is an inexpensive and easy-to-handle laboratory design that has been widely adopted in early MFC research. It simply consists of two bottles connected by a tube that can interpose an IEM or a salt bridge between anode and cathode (Bond et al., 2002; Park and Zeikus, 1999; Min et al., 2005). H-shape systems typically produce low current densities due to the high internal resistance, which limit their use to basic parameter research, such as
examining new materials or studies of microbial communities (Logan et al., 2006). Better performance can be obtained by the two flatchamber designs first developed by Delaney et al. (1984), which offer lower internal resistance due to the proximity at which anode and cathode can be put over a generally larger IEM. This compact configuration resembles that of traditional chemical fuel cells. This strategy was adopted by Min and Logan (2004) while designing their flat-plate MFC that comprises of two polycarbonate plates bolted together and contains a carbon-cloth cathode hot-pressed to an IEM also in contact with a carbon paper that serves as an anode, obtaining up to 7271 mW m2 of power density. The flattened design MFCs can easily be stacked together and electrically connected in series or in parallel in order to increase the overall system voltage (Aelterman et al., 2006b). Tubular shapes have also been designed (Rabaey et al., 2005b), or upflow types with anode below and cathode above (He et al., 2005), with the liquid sequentially passing through the two compartments. More complex designs have also been implemented to allow more complex measurements such as gas production and consumption, pH, and dissolved oxygen (Freguia et al., 2007b). When oxygen is used as electron acceptor, the cathode can be placed directly in contact with air, thus circumventing the need for a second chamber. In the single-chamber configuration, the cathode consists either of a catalyzed electrode open to the air, or is assembled with the anode within the same unit. Park and Zeikus (2003) used an MFC made of one compartment consisting of an anode coupled with a porous air-cathode directly exposed to air. In Liu and Logan (2004), an anode and a cathode were placed on opposite sides of a Plexiglas cylindrical chamber of length 4 cm and diameter 3 cm. The anode was made of carbon paper without wet proofing, while the cathode was manufactured by bonding the IEM directly on a carbon cloth (with platinum as catalyst). Liu et al. (2004) implemented the tubular shape within a single-chamber configuration for the treatment of wastewater. The anode, consisting of several graphite rods, surrounded the cathode made of carbon/Pt/IEM layers bolted together to a plastic support through which air was blown. Rabaey et al. (2005b) manufactured a tubular MFC with an inner cylindrical anode consisting in granular packed-bed graphite and an outer cathode.
4.18.5.2 Compartment Separation In MFCs, the function of the IEM is not only to provide a physical barrier to prevent fuel crossover between the compartments, but also to create a way for the ions to selectively diffuse to ensure electroneutrality. For example, as shown earlier, for every negative charge that is transferred to the cathode through the electrical circuitry, an equal amount of positive charge needs to flow through the electrolyte to prevent charge build-up. Finally, it also prevents the electrolytes from large pH fluctuations due to proton production at the anode and proton consumption at the cathode. Nafion (DuPont Inc., USA) and Ultrex CMI-7000 (Membranes International Inc., USA) are largely applied CEMs.
Microbial Fuel Cells
Although Nafion CEMs have been widely used in fuel-cell research, they do not perform as well under typical conditions at which MFCs work, for example, neutral pH, and in the presence of other cations in concentrations that can be 105 times higher than the proton concentration. Rozendal et al. (2006a) showed that under these conditions, Nafion membranes mainly transfer other cations rather than protons, thus lacking specific selectivity for protons. Ultrex is a more general CEM with larger mechanic strength compared to Nafion (Harnisch et al., 2008). It is considered a more cost-effective alternative to Nafion. The reader can refer to the works of Rozendal et al. (2008c) and Harnisch et al. (2008) where alternative types of membranes are compared in MECs and MFCs. Several attempts have been made by researchers toward the development of membrane-less MFC, in which the IEM is absent (e.g., sediment MFCs), or is replaced by different types of separators. Liu and Logan (2004) studied how the performances are affected by the presence or the lack of a Nafion membrane. Their results showed that increased power densities were possible without the IEM. Nevertheless, the enhanced oxygen diffusion led to a decrease of Coulombic efficiency as a higher portion of the carbon source was oxidized without electrons transferring to the anode. In an attempt to increase the oxygen-diffusion resistance, Park and Zeikus replaced the IEM with a porcelain septum (100% kaolin) and despite obtaining higher power outputs, the Coulombic efficiency was fairly low (Park and Zeikus, 2002, 2003). The addition of successive layers of polytetrafluoroethylene to the cathodic air-side of a single-chamber MFC resulted in increased coulombic efficiency and increased maximal power density (Cheng et al., 2006a). Jang et al. (2004) designed a membrane-less MFC in which anode and cathode were physically assembled within the same reactor unit and separated by glass wool and glass bead layers. Anode and cathode (made of graphite felt) were placed at the bottom and the top of the reactor and an upflow was imposed through the cylinder. Oxygen was bubbled in the cathode and its back diffusion was avoided simply by the stream flow. The protons formed during the anodic reaction were transported to the cathode by the same liquid stream. The results showed that the internal resistance was excessively high (several kO) due to the large anode to cathode distance, which resulted in low power generation. Moreover, most of the COD was removed in the cathode compartment by direct reaction with oxygen rather than by bioelectrochemical oxidation at the anode.
4.18.5.3 Electrodes MFC anodes and cathodes are typically made of graphite, which can be in the form of rods, felt, carbon paper, or cloth. Reactions occurring at the electrodes are subject to activation energies that need to be reduced by the use of appropriate catalysts. In the anode compartment, bacteria normally accomplish the role of catalysts. The electrodes therefore need to provide a suitable surface for the bacterial growth. Rough surfaces may provide several opportunities for adhesion, as well as decrease the current densities and therefore the potential losses, as is further described later (see Section 4.18.9.).
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Increasing the surface area of the anode is also the simplest way to increase the loading rates that can be processed in an MFC as higher quantities of biomass can grow in the same reactor volume. Granular graphite is considered a convenient material for MFCs due to its cost (approximately US$ 0.5 kg1) and its high surface area and roughness. The porosity of the graphite is also an important aspect that needs to be taken into account when calculating the specific active surface of an electrode, as bacteria can grow in pores with a size larger than the bacteria themselves, but the smallest pores cannot be colonized and therefore do not contribute to the active electrode surface area. In addition, graphite granules have a high internal volume, which takes up about half the total reactor volume. Thus, carbon fiber brushes are increasingly considered as promising for future applications (Logan et al., 2007). Metal electrodes made, for example, of stainless steel can also be used (Tanisho et al., 1989), but despite being suggested as a good cathodic material, it has been shown to be less effective when used at anodes (Dumas et al., 2007). Moreover, metal electrodes do not normally offer a high specific surface area and their higher cost when compared to graphite limits their application, especially with regard to larger-scale use (Rozendal et al., 2008a). Uncoated titanium was also proposed by ter Heijne et al. (2008), although, based on DCvoltammetry and on electrochemical impedance spectroscopy (EIS), it was concluded that uncoated titanium was not suitable as an anodic material. Kargi and Eker (2007) proposed the use of copper and copper–gold electrodes, obtaining current and power production largely comparable with other studies. Increased performances have been obtained by adopting chemical–physical strategies, like incorporating Mn(IV) and using neutral red covalently linked to mediate electron transfer to the anode (Park and Zeikus, 2003). The use of materials such as polyanilines was also shown to improve current generation (Niessen et al., 2004; Schroder et al., 2003).
4.18.5.4 Cathodic Compartment Oxygen is by far the most suitable electron acceptor for MFC operations, due to its high redox potential (see Table 1), low cost, availability, and the fact that it does not produce any unwanted reaction product. However, its slow reduction kinetic on plain graphite requires in most instances the use of an appropriate catalyst. Platinum has been largely used in chemical fuel cells as an abiotic catalyst of the cathodic reaction. Nevertheless, platinum is not likely to be suitable for most of MFC applications because of its poisoning sensitivity toward some components in the substrate solution, especially to H2 S. The above-mentioned ferricyanide commonly used at MFC cathodes, but it cannot be considered a mediator for oxygen reduction, as its oxidation rate is much slower than its reduction (Pham et al., 2004). It acts therefore as an electron acceptor on its own, thus needing periodical replenishment. Yet, ferricyanide has the advantage of having a very low overpotential on plain carbon electrodes and operates at a potential close to its open circuit value. In spite of its sensitivity, it has been adopted for dissolved oxygen or open-air cathodes (Liu et al., 2004; Reimers et al.,
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2001), with 0.5 mg cm2 Pt loading being used in several studies (Liu and Logan, 2004). Platinum content as low as 0.1 mg Pt cm2 was also shown as effective (Cheng et al., 2006b). In an attempt to reduce the cost of the catalyst, some researchers have tested alternative electrode compositions where redox mediators were bound to the surface of the electrode, avoiding at the same time the use of soluble mediators. Park and Zeikus (2003) described a technique to bind ferric sulfate to woven graphite surfaces for improved oxygen reduction in an air-cathode MFC. Fe(III) was reduced to Fe(II) by the electrons generated at the anode and Fe(II) was subsequently re-oxidized by oxygen. An inexpensive cobalt-based material, cobalt tetramethylphenylporphyrin, was tested by two research groups (Cheng et al., 2006b; Zhao et al., 2006), both concluding that the performances were comparable with that of platinum but using a material less susceptible to poisoning. Zhao et al. (2005) found that transition metals phthalocyanines and porphyrins exhibit catalytic activity comparable to platinum. Other compounds have been employed to enhance cathode catalysis on active carbon or titanium electrodes, including cobalt oxide and molybdenum/ vanadium (Habermann and Pommer, 1991). Recently, the possibility of biocatalyzing the cathodic reaction has opened up a number of new opportunities. Biocathodes have a number of advantages compared to conventional chemical catalysts, such as their low cost, self(re)generation capacity, and the fact that they are less sensitive to the components typically present in the wastewater. In most of the biocathode studies in which oxygen was the final electron acceptor, microorganisms were used to transfer electrons from a reduced form of the metal compounds to oxygen itself. Manganese and iron have been used to transfer electrons from the electrode to oxygen by means of biological processes (Bergel et al., 2005; Rabaey et al., 2008; Rhoads et al., 2005). Ter Heijne et al. (2007) developed an oxygen cathode mediated by the couple Fe3þ/Fe2þ at very low pH with biological reoxidation of ferrous ions with oxygen by a culture of Acidithiobacillus ferrooxidans. Compounds other than oxygen can be also used as terminal electron acceptors. Nitrate, sulfate, iron, manganese, uranium, selenate, arsenate, urinate, fumarate, and carbon dioxide are all possible candidates for MFC applications (He and Angenent, 2006). Examples also exist of the use of oxygenase enzymes such as the multi-copper oxygenase laccase, as catalysts for oxygen reduction (Schaetzle et al., 2009). Although, the high costs together with the limited lifetime and stability of the enzymes are important drawbacks of enzymatic electrodes that need to be addressed.
4.18.6 Electrochemically Active Microorganisms and Extracellular Electron Transfer The underlying working principle of an MFC is extracellular electron transfer (EET; It refers to a mechanism by means of which bacteria donate or accept electrons to and from an electrode; Chang et al, 2006). Microorganisms use EET in order to utilize insoluble electron acceptors (or donors) that cannot enter the cell (Rabaey et al., 2007). Bacterial interaction with an insoluble electron acceptor has been first studied for
microorganisms that respire on Fe(III) and Mn(IV) or oxidize large humic substances that cannot enter the bacterial cell (Lovley et al., 1996, 1987; Myers and Nealson, 1988). Two pathways of EET are currently assumed to be used by microorganisms (Figure 5):
• •
through electron through electron
mobile components (also referred to as mediated transfer pathway) or immobilized structures (also referred to as direct transfer pathway).
Redox mediators (or shuttles) are soluble compounds that can transfer electrons between the microbial cells and the electrode surface. Reactions involving redox mediators can in principle occur outside or inside the cells (Gralnick and Newmann, 2007). In the first MFC prototypes, soluble redox mediators were added to the media to aid EET from bacteria to an electrode. The characteristics of redox mediators are: (1) the ability to be reversibly oxidized and reduced, (2) the resistance to biological degradation, (3) fast kinetics of oxidation at an electrode, (4) ease of diffusion through bacterial membranes, and (5) nontoxicity toward microbial consortia. Substances such as neutral red, hexacyanoferrate, thionin, or quinones were used to promote EET (Kim et al., 2000; Park and Zeikus, 2000). Delaney et al. (1984) and Allen and Bennetto (1993) developed and improved MFCs using different combinations of microorganisms and mediators. They showed that the use of suitable mediators could enhance both the efficiency and the rate of electron transfer. More recently, live–dead staining and confocal microscopy analysis showed that even in systems were no exogenous mediators were added, microorganisms could anyhow contribute to electricity generation. This means that bacteria growing at a certain distance from the electrode can also demonstrate EET. It was reported by Rabaey et al. (2005a) and Hernandez et al. (2004) that redox active compounds such as pyocyanin and phenazine-1-carboxamide were self-produced by Pseudomonas species. In particular, these compounds were essential for electricity production by Pseudomonas aeruginosa. The production of endogenous mediators was thereby identified as an additional strategy enabling mediated EET. Although redox mediators were long thought to be essential to enable EET, and were therefore extensively used in MFCs, the finding that bacteria have the ability to reduce insoluble electron acceptors such as Fe(III) and Mn(IV) in oxide forms by Lovley and Phillips as early as 1988 (Lovley and Phillips, 1988) already suggested that mediated EET was not necessarily the only mechanism for EET. This discovery indeed represented a landmark in MFC research and opened the door to the development of mediator-less MFCs. If neither endogenous nor exogenous redox mediators are used, a direct contact between the outer membrane of the bacterial cell and the electrode surface must be established in order to promote electron transfer. Direct electron transfer requires that the microorganisms rely on a transport structure that enables electron crossover to the outside of the cell where they can be delivered to a solid electron acceptor (a metal oxide or, more pertinently, to an MFC anode). Unusually high content of c-type cytochrome in Shewanella putrefaciens outer membrane during anaerobic growth was
Microbial Fuel Cells
A Anode
Substrate
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CO2
e−
Medox
B Medred
Bacterial cell C
Bulk
Figure 5 Possible microbial interactions with the anode. Microorganisms can transfer electrons through immobilized structures such as membrane-bound proteins (A) or electrically conductive pilus (B), or through mobile components (redox mediators or shuttles) that are alternatively oxidized and reduced at the electrode (C).
reported in as early as 1992 (Myers and Myers, 1992). Kim et al. (1999) were the first to measure the electrochemical activity of Shewanella putrefaciens when grown under anaerobic conditions without nitrate, as revealed by cyclic voltammetry. Cyclic voltammetry is an electrochemical technique where an electrode is immersed in a medium and its potential is changed cyclically by a potentiostat while the current is measured. The current versus potential diagram (called cyclic voltammogram) shows peaks in correspondence of the potential of each reversible redox couple in contact with the electrode. A double peak in the cyclic voltammogram revealed that a redox active compound (possibly a cytochrome) was responsible for the electrochemical activity of the cells of Shewanella putrefaciens. The study also showed that the cells lost their electrochemical activity when grown aerobically. In a further study, the electrochemical activity of Shewanella putrefaciens was demonstrated by current production in a mediatorless MFC (Kim et al., 2002). Bond et al. (2002) showed that some bacteria are capable of transferring electrons from anoxic marine sediments to an anode, connected to a cathode in the overlying aerobic zone. Community analysis of these bacteria showed that many of them belonged to the d-proteobacteria phylum. In particular, Geobacter metallireducens and Desulforomonas acetoxidans were identified. Another bacterium,
Geobacter sulfurreducens, was successfully tested by Bond and Lovley (2003) as a pure culture in a two-chamber fuel cell with acetate as substrate. Chaudhuri and Lovley (2003) demonstrated that the bacterium Rhodoferax ferriducens can perform electron transfer to an anode when fed with glucose. Recently, it has been reported that outer-membrane proteins are not always sufficient for the reduction of Fe(III) oxides. Geobacter and Shewanella species were shown to produce conductive appendages that were referred to as ‘nanowires’ (Gorby et al., 2006; Reguera et al., 2005). In Gorby et al. (2006) the conductivity of these pilus-like nanowires produced by Shewanella oneidensis was measured through conductive-scanning tunneling microscopy. A similar observation was done for nanowires produced by Geobacter sulfurreducens using conducting-probe atomic-force microscopy. Results show that while Shewanella species appear to produce rather thick bundles of conductive wires, Geobacter seems to produce more thin structures. Evidence also exists of interspecies electron transfer between bacteria in mixed communities, as suggested, for instance, by the presence of filaments connecting the propionate-fermenting Pelotomaculum thermopropionicum with the methanogen Methanothermobacter thermautotrophicus in Ishii et al. (2005). Whether interspecies electron transfer occurs
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through wired connections or the use of mediators, it may significantly impact BES engineering as it will broaden the array of organisms able to survive on an electrode. Although the nomenclature for microorganisms that perform EET is far from being uniform (Lovley, 2008), the term electrochemically active microorganisms has been introduced to refer to microorganisms that demonstrate the ability to perform EET on an electrode without the addition of exogenous mediators (Chang et al., 2006). More specifically, the term ‘electrode-reducing microorganisms’ refers to organisms that can use an electrode as electron acceptor (thus reducing the electrode), whereas ‘electrode-oxidizing microorganisms’ refers to the organisms that use the electrode as electron donor (and thus oxidize the electrode) (Lovley, 2008). Community analyses of mixed anodic cultures enriched in electrochemically active microorganisms were undertaken by several investigators (Holmes et al., 2004; Kim et al., 2007b; Rabaey et al., 2004a). The striking conclusion from these studies was that EET capacity is widespread in nature and it is found among most phyla of bacteria. The ubiquity of these microorganisms in nature is also confirmed by the diversity of inocula that can be used to start up lab-scale MFCs. Raw sewage (Liu et al., 2004), activated sludge (Lee et al., 2003), anaerobic and methanogenic sludge (Rabaey et al., 2003), river sediments (Gregory et al., 2004), and seawater (Bond et al., 2002) were all successfully used to develop bioelectrochemical activity. Pure or enriched mixed cultures have shown the ability to use anodes to oxidize a variety of organic substrates, including acetate (Bond and Lovley, 2003), propionate (Bond and Lovley, 2005), butyrate (Liu et al., 2005c), ethanol (Kim et al., 2007b), lactate (Kim et al., 1999), glucose (Chaudhuri and Lovley, 2003), domestic wastewater (Gil et al., 2003), and beer-processing wastewater (Wang et al., 2007). Whereas respiratory flexibility in mammalian mitochondria is rather poor, it can be extremely broad in Bacteria and Archaea, as a diverse range of electron acceptors can be used, including nitrogen oxyanions and nitrogen oxides, elemental sulfur and sulfur oxyanions, halogenated compounds, transition metals such as Fe(III) and Mn(IV), as well as radionuclides such as U(VI) (Richardson, 2000). It is widely agreed upon that the first respiratory processes to evolve on Earth over 3.5 billion years ago would have used Fe(III) or S(0) as electron acceptors. The fact that the ability to electrically interact with an insoluble electron acceptor is widespread in nature is not therefore particularly surprising.
4.18.7 Oxidative Processes As we have seen previously in Section 4.18.3, the difference in reduction potentials between primary electron donor and terminal electron acceptor determines the net energy change of the reaction, and thus the energy gain for chemotrophic microorganisms. Energy is conserved by production of adenosine triphosphate (ATP) molecules. Depending on the availability of electron acceptor, two possible pathways are possible: fermentation or respiration. In the case of fermentation, the electron acceptor is a compound internally generated from the initial substrate, whereas in the case of respiration, the electron acceptor is externally provided. When
oxygen is the electron acceptor, it is referred to as oxic respiration. When the electron acceptor is oxygen linked with other compounds, it is referred to as anoxic respiration. The basic respiratory process involves the transfer of electrons from a low redox potential electron donor such as nicotinamide adenine dinucleotide (abbreviated as NADþ in its oxidized form, and as NADH in its reduced form), to the terminal electron acceptor at a high redox potential. The transfer occurs through a chain of intermediate redox complexes. The case where an insoluble electron acceptor rather than a soluble compound is oxidized is still a form of respiration, with similar mechanisms for energy generation to other forms of respiration. Anodic oxidation relies on the tricarboxylic acid (TCA) cycle, which together with glycolysis and pyruvate oxidation before the TCA cycle and electron transfer chain after it, permits the chemical conversion of the organic substrates into carbon dioxide and water, and generates energy in the form of ATP (White, 1995). Many bacterial species have been shown to produce electricity in MFCs using compounds such as acetate, lactic acid, and ethanol which enter the TCA cycle through pyruvate or acetyl-CoA (Bond and Lovley, 2003; Kim et al., 1999, 2007b), whereas more complex carbohydrates such as glucose require glycolysis before entering the TCA cycle. NADH, nicotinamide adenine dinucleotide phosphate (NADPH), and flavin adenine dinucleotide (FADH2) are generated through the TCA cycle. These are reduced molecules that represent the primary electron donor for the electron transport chain, through a series of membrane-associated electron carriers, including flavoproteins, iron–sulfur proteins, quinone pool, and a series of cytochromes. The electron transport chain has two basic functions: (1) to accept electrons from an electron donor and to transfer these electrons to the next electron acceptor and (2) to conserve the energy released during electron transfer for the synthesis of ATP. The electron carriers are arranged in the membrane in such a way that the electrons are transferred from one complex to the following at a higher potential (Figure 6). Hydrogen atoms removed from carriers such as NADH are separated from the electrons. While the electrons are transferred to the following carrier, the protons are pumped outside the cell (or to the periplasm in Gram-negative bacteria). This generates a proton motive force across the membrane, which extrudes up to approximately 10 protons for each electron pair derived from 1 NADH. The proton motive force drives ATP generation through a process called phosphorylation, which involves a large membrane complex called protontranslocating ATP-synthase, an enzyme that exploits the electrochemical potential liberated by the protons as they return to the cytoplasm. It is assumed that about three protons are required to generate one molecule of ATP, although recent research suggested that this stoichiometry may vary from 3 to 5 protons per ATP (Nakanishi-Matsui and Futai, 2008). It is generally accepted that maximally three molecules of ATP are generated by prokaryotes per NADH molecule (White, 1995). Bacteria in MFCs establish a direct contact with the electrode through cytochromes or nanowires or via soluble redox mediators (see Section 4.18.6). In any case, the electrontransfer chain as explained earlier cannot be entirely exploited, as the potentials of the electron carrier used to transfer electrons to the electrode (soluble mediator or cytochrome) are
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Periplasm 4H+
4H+
3H+
2H+
Complex I
Complex IV cyt c
Q
ATP synthase
Complex III NADH
+
NAD
Quinone pool
Cytoplasm
O2 + 4H+
2H2O ADP+P
ATP
Figure 6 Respiratory electron transport chain of an organism such as Paracoccus denitrificans, a model for studies of respiration. Electrons are sequentially transferred through a series of membrane-associated electron carriers, which are embedded in the lipid bilayer of membranes in such a manner that most have access to both the inside and the outside of the cell. Hþ atoms removed from carriers such as NADH are separated from electrons and pumped outside the cell or in the periplasm. The reduction of O2 to H2O plus the extrusions of Hþ during electron transport generate a pH gradient and an electrochemical potential across the membrane (proton motive force, expressed in volts). This potential energy is used to drive the formation of high-energy phosphate bonds in ATP. Complex I: membrane-spanning complex comprising of flavoproteins and Fe–S proteins. Quinone: lipid electron carrier. Complex III: comprises cytochrome b, Fe–S proteins, and cytochrome c1. Cyt c: cytochrome c. Complex IV: cytochrome aa3 oxidase. From Madigan MT, Martinko J, and Parker J (2000) Brock Biology of Microorganisms. Upper Saddle River, NJ: Prentice Hall.
normally not electronegative enough to receive the electrons from the next step in the electron transport chain. Therefore, the maximal ATP yield for electroactive microorganisms is limited by the redox level at which the transfer chain is interrupted. As bacterial growth depends on the availability of intracellular ATP, the growth yield for bacteria in MFC will ultimately depend on the mechanism of electron transfer. Furthermore, as the potential of the anode determines the last step of the electron transfer to the electrode, the microbial growth will ultimately depend upon the anodic potential (Aelterman et al., 2008; Finkelstein et al., 2006; Freguia et al., 2008b; Schroder, 2007). A broad range of biodegradable materials has been shown to serve as electron donors for electricity generation in MFCs. Volatile fatty acids (e.g., acetate, formate, and butyrate), alcohols (e.g., ethanol and methanol), as well as more complex carbohydrates (e.g., glucose, sucrose, cellulose, and even starch), and even amino acids and proteins were used as organic electron donors (Freguia et al., 2007b; He et al., 2005; Heilmann and Logan, 2006; Ishii et al., 2008; Liu et al., 2005b; Logan et al., 2005; Min and Logan, 2004; Rabaey et al., 2003). Inorganic compounds such as sulfide (Rabaey et al., 2006) and synthetic acid-mine drainage (Cheng et al., 2007) have also been reported. Thus far it is still not clear which role the different types of substrates play. Acetate was reported to be the preferred substrate when compared to buyrate (Liu et al., 2005b) or to wastewater (Rabaey et al., 2005b), suggesting that MFC organisms prefer rapidly biodegradable substances to more complex compounds. Coulombic efficiency of 100% (i.e., a stoichiometric conversion of the substrate into current) was
reported in Freguia et al. (2007b) in their acetate-fed anode, whereas much lower efficiencies were obtained when glucose was used instead. Acetate has therefore been the substrate of choice in a large number of studies. Nevertheless, even if acetate is considered inert to alternative biochemical conversions such as fermentation in MFCs (Aelterman, 2009; Freguia et al., 2007b, 2008b), acetoclastic methanogenesis has recently been reported as an important anodic electron sink when the operating conditions favor the establishment of a methanogen community alongside electrochemically active microorganisms (Virdis et al., 2009).
4.18.8 Reductive Processes Bioelectrochemical oxidation of organics at the anode of MFCs has to be coupled with a reduction reaction at a counter electrode (cathode). Several electron acceptors have been used, depending on the scope of the BES. When power generation is the goal, oxygen appears to be the preferred choice due to its high availability and its high redox potential (see Table 1). As seen in Section 4.18.5, hexacyanoferrate has been extensively used in laboratory studies focused on the anodic reaction, due to its ability to provide a constant potential. More recently, the demonstration that compounds such as nitrate (Clauwaert et al., 2007a), nitrite (Virdis et al., 2008), hexavalent uranium (Gregory and Lovley, 2005), perchlorate (Thrash et al., 2007), and trichloroethene (Aulenta et al., 2009) can be reduced at the cathode, has broadened the array of applications of MFCs on nutrient removal and bioremediation as well.
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While oxygen has been studied extensively for conventional chemical fuel cells, MFCs face specific limitations that substantially differentiate them from their hydrogen, methanol, or methane counterparts. In particular, they must operate at ambient pressure and moderate temperatures; moreover, the relatively low current outputs that are achieved in MFC do not justify the use of expensive chemicals as catalysts and, on the other hand, the nature of wastewater environments exposes most metal catalysts to irreversible poisoning. Three major bottlenecks can be identified for oxygen utilization at the cathode of MFCs: (1) the low solubility of oxygen in water limits its delivery to the electrode surface; (2) direct oxygen reduction at graphite electrodes exhibits large overpotentials; and (3) oxygen can diffuse (to some degree) through most membranes causing coulombic losses by direct oxidation of the organic electron donor. As oxygen reduction on plain carbon was found to happen at limited rates (Oh et al., 2004; Zhao et al., 2005), oxygen cathodes were developed with a platinum coating, based on the knowledge acquired from years of research in the chemical fuel cell area. It was found that the use of platinum enhanced oxygen-reduction rates when applied at a surface concentration of at least 0.1 mg Pt cm2 (Cheng et al., 2006b), with best performances at a load of 0.5 mg Pt cm2. Platinumcoated graphite electrodes have thereby set a benchmark for oxygen cathodes in MFCs, despite the fact that encouraging results had previously been obtained by immobilizing Fe(III) on graphite (Park and Zeikus, 2002, 2003). If a current of 1000 A m3 is expected to be delivered by the bioanode (Rozendal et al., 2008a), the aeration capacity that would have to be provided at the cathode can be estimated to be as high as about 0.0035 m3 O2 min1 per cubic meter of liquid, which is in the same order of magnitude as that provided by aeration systems normally applied for aeration tanks of activated sludge-treatment plants (Tchobanoglous et al., 2003). Oh et al. (2004) showed that aqueous oxygen cathodes (with Pt as catalyst) are the limiting step to electron transfer in two-chambered MFCs: a cathodic overpotential of B0.5 mV implied that about half of the total electromotive force expected was lost entirely due to oxygen reduction. The same investigators found that oxygen reduction behaves accordingly to Monod-type kinetics, with a half saturation constant of 1.74 mg O2 l1, which indicates that a further reduction of the electron-transfer rates is expected at low aeration rates. The reasons for the low performance of platinum cathodes in MFCs are not well understood. It can be speculated that the relatively mild conditions at which they operate, such as pH of around 7 in most cases and ambient temperatures, may affect the reaction rate. Conventional chemical PEM fuel cells normally sustain much higher current densities but they also typically operate at very low pH values (lower than 1). Given that protons are reactants in the cathodic reaction (Equation (3)), low pH values guarantee that protons are available in high concentration. In addition, chemical fuel cells operate at temperatures ranging from 50 to 100 1C in the case of PEM fuel cells, higher than that used for microbial fuel cells, to enhance the reaction rate. While the solubility of oxygen in water is a physical property and as such, cannot be increased, the thickness of the liquid film across which O2 has to diffuse can be minimized to reduce mass-transfer resistance. These
constraints have led to the development of open-air cathodes (Liu et al., 2004), where the cathodes have a two-dimensional structure and are open to the air, letting oxygen diffuse to the electrode surface directly from the air. This passive aeration process is more sustainable for scale-up applications as it does not entail large energy requirements for air pumping. In order to increase oxygen supply to the cathode surface, rotating cathodes have also been developed (He et al., 2007). Despite the widespread use of platinum as cathode catalyst, its high cost and energy-intensive production technology make this metal usually unsuitable as a catalyst for wastewater applications. Considering that the economic feasibility of MFCs is strongly dependent on the cathodic compartment (almost half of the capital costs of MFCs are associated with the cathodic compartment when platinum is used as catalyst; Rozendal et al., 2008a), platinum needs to be replaced by alternative catalytic materials. Three strategies have been explored thus far: (1) the use of a material with increased surface area; (ii) alternative chemical catalysts; and (iii) biocathodes. Freguia and co-workers (2007c) have shown that by using a noncatalyzed material with a high surface area it was possible to decrease the overpotential for cathodic oxygen reduction. As shown by Equation (15), the activation overpotential Z at the cathode increases with the current density i. The use of a better catalyst has the effect of reducing the overpotential as it increases the exchange current i0, which is a characteristic of the material used. The approach of the researchers instead was to reduce the current density by using a material with a higher surface area (coarse highly porous industrial-grade granular graphite) rather than modify the exchange current using a catalyst. The current generated with this configuration was able to sustain COD removals up to 1.46 kg COD m3 d1, which is similar to that of a conventional aerobic process based on activated sludge. Among other strategies for enhancing the cathodic reaction, the use of bacteria as catalyst has attracted particular interest in the MFC field. Similarly to bioanodes, biocathodes utilize the electrical interactions that can be established between the microbes and the cathodic electrode. Biological catalysis has been shown to enhance the rate of cathodic oxygen reduction at a stainless-steel cathode in seawater sediments (Bergel et al., 2005). The existence of such a consortium of bacteria was also later demonstrated for freshwater applications (Clauwaert et al., 2007b; Freguia et al., 2008a). The involvement of microorganisms in the catalysis was unequivocally established by a pure culture study (Rabaey et al., 2008). Freguia et al. (2008a) observed increased current production in an MFC system operated with recirculation of the anode effluent to the cathode, and correlated this result to the increased microbial activity at the cathode. Also, Rozendal et al. (2008b) reported of a cathodic microbial consortia catalyzing hydrogen production at a graphite cathode. The concept of using denitrifying bacteria to reduce nitrate in the cathode of an MFC was first proposed by Lewis more than 40 years ago (Lewis, 1966). However, it was only recently that the presence of reactions involving nitrogen at the cathode was confirmed. Catalytic reduction of nitrate and nitrite driven by electric current has been explored by Mellor et al. (1992). Nitrate-enriched water was pumped into the anode chamber of an electro-bioreactor and recirculated into the
Microbial Fuel Cells
cathode chamber where purified NADH:nitrate reductase, nitrite reductase, and N2O reductase enzymes were immobilized on the surface of the cathode. The applied electric current provided the reducing power needed to carry out the process. Hydrogen is an excellent electron donor and it can be easily produced by electrolysis of water. When a denitrifying biofilm is growing on the surface of a hydrogen-producing electrode, it has the advantage of having a continuous supply of an electron donor to carry out the process. Several studies attempted therefore to obtain nitrate reduction by applying a potential difference to form hydrogen at the cathode (Kuroda et al., 1997; Sakakibara et al., 1994; Sakakibara and Kuroda, 1993). Kuroda et al. (1997) extended the concept to obtain simultaneous COD removal and denitrification. However, hydrogen was still the actual electron donor for nitrate reduction. Although bacteria were considered as able to directly use the electrode as the sole electron donor, it was only very recently that this microbial capability was experimentally verified. Gregory et al. (2004) showed that a bacterial culture enriched in Geobacter species could reduce nitrate (NO3 ) to nitrite (NO2 ) using the cathode as the sole electron donor, without producing hydrogen as redox mediator. As nitrate reduction did not occur in the absence of bacteria, the researchers concluded that the process was biochemically activated by the biofilm, showing for the first time that bacteria were using the cathode as the sole electron donor. A later study confirmed their hypothesis and showed that complete denitrification to nitrogen gas could be achieved (Park et al., 2005). More recently, full denitrification with simultaneous carbon removal was reported in MFCs by Clauwaert et al. (2007a) and Virdis et al. (2008). Besides denitrification, other cathodic reactions have also been reported. Perchlorate, a compound extensively used in industry, was shown to be reduced with the help of 2,6anthraquinone disulfonate (Thrash et al., 2007). Shea et al. (2008) coupled a perchlorate-reducing biocathode with an acetate-oxidizing bioanode in an MFC configuration. Recently, an as yet unknown self-produced redox mediator appeared to be involved in the reduction of trichloroethene to more reduced compounds such as vinyl chloride and ethane (Aulenta et al., 2009). Finally, the recent demonstration of methane production from carbon dioxide reduction (Cheng et al., 2009) and biocathodic alcohol production from VFAs (Steinbusch et al., 2008, 2009), has opened up new possibilities for BES applications to biofuel production.
4.18.9 Challenges toward Improving MFC Efficiency Currently, several bottlenecks of both microbiological and technological nature limit the efficiencies of MFCS. It implies that despite the fact that laboratory MFCs already produce current densities suitable for practical applications, full-scale implementations are not necessarily straightforward (Rozendal et al., 2008a). The presence of alternative electron acceptors in the anode compartment (e.g., due to crossover of electron acceptors from the cathodic compartment to the anode), competitive processes such as fermentations and methanogenesis, and bacterial growth, are recognized as responsible for diverting a part
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of the total electrons provided by the electron donor from the electricity-generation process, thus reducing the coulombic efficiency. The coulombic efficiencies reported in literature range from 1% (Kim et al., 2002) to about 100% (Freguia et al., 2007b). Besides, the presence of overpotentials at the electrodes, ohmic losses due to wastewater and electrode conductivity, and pH gradients due to imperfect ion selectivity of IEMs, are responsible for reducing the ECE as they reduce the cell voltage, described previously in Section 4.18.4. Their practical implications are described in this section.
4.18.9.1 Minimizing Electrode-Potential Losses Overpotentials at the electrodes can significantly limit the performance of MFCs as they decrease the actual voltage attainable, and therefore the energy efficiency. These losses are due to the electron-transfer kinetics from the microbial cells to the electrode (and vice versa) and to the bacterial metabolic kinetics. Moreover, as all heterotrophic bacteria retain a portion of their carbonaceous substrate to produce more biomass, a coulombic (and energetic) loss from this activity has to be taken into account as well. Interestingly, the potential losses at the anode have been reported to be much lower than that at the cathode. If, for example, an MFC can theoretically produce up to 1.1 V, less than 0.1 V is typically lost at the anode and more than 0.5 V can be lost at the cathode under working conditions (Logan et al., 2006). This would leave 0.5 V for power generation, without taking into account other losses such as ohmic losses. It is therefore obvious that any strategy intended to reduce the MFC overpotentials would have to pay particular attention to the cathodic reaction. As discussed above, the reasons for the high cathodic overpotential are mainly due to the slow kinetics of oxygen reduction. The use of biocathodes may be a valuable alternative to improve this catalytic process.
4.18.9.2 Respiration, Fermentation, and Methanogenesis In a complex environment such as wastewater, a multitude of other processes may occur alongside the conversion of organic molecules to electrons, thus competing with electricity generation in MFCs. When electron acceptors other than the electrode are present in the anode chamber, the organic electron donor can be oxidized using alternative pathways, thus reducing the electron transfer efficiency. In particular, nitrate and sulfate are commonly found in wastewater and are likely to divert the substrate electrons from the anode, as their redox potential makes them often more favorable electron acceptors than the anode. The presence of oxygen in the anodic compartment depletes substrate electrons through aerobic oxidation. Oxygen may be present due to possible pre-treatment of the MFC influent or through diffusion from the cathodic compartment in the case of oxygen cathodes. The potential for oxygen crossover to the anode is considerable in membraneless configurations (Liu and Logan, 2004). Methane is the end product of most anaerobic processes and it is regarded as one of the major bottlenecks of anode operations in MFCs, because methanogens compete with electrochemically active microorganisms for the organic material in the wastewater. It was recently shown that notable
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amounts of methane were produced with glucose as substrate, consuming between 15% and 50% of the fermentable substrate electrons (Freguia et al., 2007a, 2007b, Lee et al., 2008). Methane was also observed in ethanol-fed MFCs (Torres et al., 2007). Fermentable substrates such as ethanol and glucose yield hydrogen when fermented. Hydrogen can be used by electrochemically active microorganisms for electricity production, or it can be further converted into methane through hydrogenotrophic methanogenesis. Observations suggest that electrochemically active microorganisms cannot completely outcompete methanogens for hydrogen (Freguia, 2008). One possible explanation for this behavior is that contrary to electroactive microbes, methanogens can grow at any distance from the electrode, as they do not need it as an electron acceptor. Therefore, they can grow on the top layer of the anodic biofilm, where they can scavenge the hydrogen that is formed by fermentation before it reaches the underlying electrochemically active biofilm. In the perspective of a real application, pre-fermentation would be required whenever fermentable substrates are present in the wastewater, in order to convert fermentable substrates into nonfermentable substrates such as acetate, thus providing electrode-reducing organisms with better chances to compete with methanogens. However, it is worth noting that when a bioelectrochemical system is operated at a controlled potential, or when a low anodic potential is the result of a low-current-producing process, the energy gain for electrochemically active organisms may become too low to successfully compete with low-energyyielding biological processes (e.g., acetoclastic methanogenesis). To conclude, controlling competitive processes requires a complex synergy of operational strategies in order to avoid the conditions at which methanogens are likely to scavenge the electrons away from electrochemically active microorganisms.
4.18.9.3 Reducing pH Gradients As extensively recalled throughout the text, anodic reactions produce protons while cathodic reactions (such as oxygen reduction) consume them. IEMs are typically used to provide physical separation between the two compartments while enabling the transport of charge at the same time (Section 4.18.2). The bottleneck that typically draws from the use of IEMs in MFCs derives from their lack of selectivity. Rozendal et al. (2006a) showed that Naþ, Kþ, and NHþ 4 normally account for most of the ionic-charge transfer across Nafion CEMs, due to their typically much higher concentrations in wastewaters compared to the Hþ concentration. Further research revealed that limited proton transfer occurs with most types of membranes, including AEMs and charge mosaic membranes (CMMs) (Rozendal et al., 2008c). The inefficient proton transport causes a pH gradient across the membrane, which results in an acidic anolyte and an alkaline catholyte, accordingly to the stoichiometry described in Equations (2) and (3). The consequence of the membrane gradient is a significant decrease in performance due to the reduction of the electromotive force. From the Nernst equation (Equation (9)), an increase in proton concentration at the anode results in a higher anodic potential and, similarly, a
reduction in the cathodic potential, causing an estimated loss of B0.06 V per pH unit (Rozendal et al., 2007). Both domestic and industrial wastewaters are characterized by limited alkalinity, which during MFC treatment has to approximately match the quantity of protons produced by the anodic reaction. According to Equation (2), 24 mol of Hþ are produced per mole of glucose, which translates to 4 mol Hþ per mole of COD. The alkalinity should therefore be about 4 times the influent COD molar concentration. This is normally not a problem in lab-scale MFCs that work on highly buffered synthetic media (in the range of 60–100 meq l1). Nevertheless, it would represent an important limitation when treating real wastewater. A domestic wastewater with 500 mg COD l1 would in fact require a 62.5 mM buffer, which is already much higher than the typical alkalinity reported (50–200 mg l1 as CaCO3 (Tchobanoglous et al., 2003), equivalent to 1–4 meq l1 buffer). Membrane-less designs would partially solve the issue. Yet, as mentioned previously, the lack of physical separation between anode and cathode would lead to the crossover of electron acceptor to the anode with significant reduction of the electron recovery. Bipolar membranes (BPMs, the twolayer combinations of a cation and an anion exchange membrane) can partially solve the problem by splitting water into Hþ and OH in the liquid space between the two membranes. However, they do so at the expense of a larger membraneinternal resistance, resulting again in a reduced power output. The sequential loop operation of anode and cathode (Freguia et al., 2007a, 2008a; Virdis et al., 2008) partly alleviates the problem by enabling convective transfer of protons from anode to cathode together with the liquid stream. Proton production by the anodic process may also negatively affect the performance of MFCs at the biofilm level, when the protons do not leave fast enough and accumulate within the biofilm. The Nernstian effect discussed earlier would occur at a smaller scale with the same effect of reducing the total electromotive force attainable. Increasing the specific surface area of the electrode may provide a significant benefit when it promotes an increased biofilm/liquid contact interface area as it would increase the proton flux (Torres et al., 2008).
4.18.9.4 Wastewater and Electrode Resistance Ohmic losses derive from the resistance of materials to the transfer of charged particles (see Section 4.18.4). Ohmic losses can be considerable in real wastewaters as they typically have low conductivity (of the order of only 1–4 mS cm1). As noted by Rozendal et al. (2008a), in full-scale MFC applications delivering 10 A m2 anode surface area, the ohmic loss that is encountered would be B1 V cm1 distance between anode and cathode for a wastewater with conductivity of 1 mS cm1, which is already B90% of the theoretical maximal voltage attainable. As increasing the ionic strength of wastewater by salt addition is not economically feasible in practice, the design and the operation of a full-scale MFC can significantly affect the extent of the ohmic losses. Researchers have attempted to reduce the ohmic voltage losses by testing different types of IEMs or even by completely removing them from the system. These investigations have revealed that the internal ohmic resistance is not controlled by
Microbial Fuel Cells
the membrane but by the electrolyte. Kim et al. (2007a) tested several membranes in identical MFCs. These included cation and anion exchange membranes as well as different kinds of ultrafiltration membranes. All MFCs tested exhibited similar internal resistances, confirming that the membranes did not control the overall resistance. Liu and Logan (2004) showed that the complete removal of the CEM did increase the power output, but in that case the increase was attributed to a higher cathodic potential and not to a reduction of the internal resistance. Moreover, the coulombic efficiency dropped to 12% as no barrier for oxygen diffusion was present in the system. Liu et al. (2005a) observed that the current output increased with the ionic strength, which was set by the salt concentration. This result further confirmed that the ionic conductivity of the medium plays a crucial role in MFC performance by determining its internal resistance. Keeping the electrodes in very close proximity is crucial for reducing the ohmic losses; flat compartment systems have therefore been used in research, either singly, or as multiple stack electrically connected in series or in parallel. Flat systems can minimize the ohmic losses as the compartments can be placed very close, thereby reducing the travel distance of ions through the electrolyte between the electrodes. However, the drawback of this configuration when expanded to larger scale is that in order to keep the same electrode distance with a rather larger volume, the electrons would need to travel longer distances to reach the cathode, thus significantly increasing the electrode ohmic loss if the material that is used is not sufficiently conductive. Highly conductive current collectors such as stainless-steel meshes can be used alongside the carbon/ graphite electrodes, although they can significantly increase the overall cost of the MFC. The use of bipolar plates can partially solve this problem as the distance that the electrons need to travel is reduced compared to the single-cell design (Shin et al., 2006). Bipolar plates (usually made of graphite) connect the anode side of one cell to the cathode of the next cell. The electrons generated at the anode only need to cross the bipolar plate to the cathode. However, this creates a stackin-series arrangement of the cells and one of the common problems encountered in such systems is the cell reversal, that is the reversal of the cell polarity, which turns some cells in the stack into electrolytic cells (Aelterman et al., 2006b).
4.18.10 Opportunities for Bioelectrochemical Systems In spite of the great need for improvement, MFCs are undoubtedly a promising technology that offers a vast range of potential applications. Bioelectrochemical wastewater treatment is a novel and promising approach to the production of renewable energy and thus has been the main focus of investigations. Nevertheless, the key characteristic of BESs highlights the fact that they decouple the oxidative and the reductive process, offering the unique opportunity of having clean electrons (reducing power) derived from renewable resources that can potentially be used to drive a multitude of biotechnological processes. Important examples for bioremediation processes include the reductive dechlorination of chlorinated compounds (Aulenta et al., 2007), the reduction
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of soluble metals like hexavalent uranium into more insoluble forms (Gregory and Lovley, 2005), and denitrification (Clauwaert et al., 2007a, Virdis et al., 2008). Other niche applications of MFCs include 1. Electricity production in remote areas. MFCs can produce power from waste biomass, which is ubiquitous and could supply those areas with small amounts of electricity while reducing the environmental impact on local waterways. 2. Bacterial batteries. Energy stored in the form of sugars or other organic substrates can produce environmental friendly power that could be used for the small appliances. 3. Online sensors. The production of electric current in the presence of biodegradable material could be exploited for the online detection and quantification of soluble organics in waterways or wastewater treatment plant (WWTP) effluents.
4.18.10.1 Wastewater Treatment The power densities generated by MFCs are much smaller than those of chlorofluorocarbons (CFCs). Therefore, MFCs cannot compete with CFCs as power producers, but they become much more attractive if the production of electricity is combined with wastewater treatment. The COD contained in wastewaters can be thoroughly removed while producing a CO2-neutral power, which could potentially cover at least the electricity requirements of the WWTP. Complex substrates have been successfully used to generate power in MFCs, including domestic wastewater (Liu et al., 2004), anaerobic digesters’ effluent (Aelterman et al., 2006a), brewery wastewater (Feng et al., 2008), and paper-recycling wastewater (Huang and Logan, 2008). Additionally, if anodic carbon oxidation is coupled with cathodic nitrogen removal, the use of MFCs opens up new perspectives for an integrated and sustainable wastewater treatment process. Wastewater treatment with MFCs would also offer the unique feature of online monitoring of the process through current and electrode-potential measurement that can rapidly advise of system failures. A drop in the current coupled with a rise of the anodic potential would indicate, for instance, a drop in the catalytic activity of the anodic biofilm or a failure in the feeding system. The removal of organics contained in wastewater is considered as energy efficient when no energy is consumed for aeration. If this is provided by means of passive aeration, for instance, it would save around 0.7–2 kW h1 kg1 COD removed (Logan et al., 2008) for conventional aeration, and energy can indeed be harvested from the substrate, with a theoretical upper limit of 4.4 kW h1 kg1 COD. Moreover, as the energy gain for bacteria growing at the MFC anode is generally lower than that for aerobic processes (Section 4.18.6), sludge production would also be reduced. The bacterial-growth yield in MFCs is in fact expected to be between that of high energy-yielding aerobic processes (0.4– 0.6 g COD biomass g1 COD substrate according to Heijnen (1999)) and that of low energy-yielding anaerobic treatment (0.01–0.14 g COD biomass g1 COD substrate, Heijnen (1999)). Rabaey et al. (2003) reported a measured yield of 0.07–0.22 g COD biomass g1 COD substrate for a glucose-fed
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MFC. Freguia et al. (2007b) reported growth yields at anodes ranging from 0 to 0.3 g COD biomass g1 COD substrate acetate. MFCs can achieve significant increased organic removal rates compared to aerobic processes. Laboratory reactors have in fact reached current densities of the order of B10 A m2 anode surface area (Fan et al., 2007; Torres et al., 2007). Rozendal et al. (2008a) evaluated that this would correspond to a volumetric wastewater treatment capacity of B7.1 kg COD m3 reactor d1, assuming a minimal compartment thickness of 1 cm. If the same performances could be obtained on a larger scale, wastewater treatment with MFC would even outcompete traditional aerobic treatments, which are able to process B0.5–2 kg COD m3 d1 (Logan et al., 2006), with few advantages:
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•
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MFCs can produce power and treat wastewater in a single stage, whereas anaerobic processes require expensive power-generation facilities: an anaerobic digester for the production of methane and a power-production stage in a gas turbine/engine. MFCs can obtain good effluent quality while treating dilute influents at temperatures below 20 1C (Pham et al., 2006), while anaerobic digesters work best with high-strength wastewater at increased temperatures (30–37 1C) and require further polishing of the effluent. In a (bio)electrochemical system, it is possible to utilize streams containing sulfur (Dutta et al., 2008; Rabaey et al., 2006) with no need for gas treatment, whereas the methane produced during anaerobic digestion normally contains traces of H2S that need to be removed before using the gas in a turbine (due to corrosion concerns and environmental regulations) or fed to a chemical fuel cell (as sulfide poisons the Pt catalyst).
However, to achieve practical implementation at a reasonable scale, several challenges will have to be solved (Section 4.18.9). Moreover, the capital costs of MFCs have to be reduced drastically as the material costs are very high and there is a limited economy of scale benefit, primarily due to the close anode/cathode distance required. In conclusion, MFCs should be considered more as a complementary system rather than as in competition with anaerobic digestion. Anaerobic digestion can be applied to the treatment of high-strength waste streams. Industrial effluents represent perfect examples of these applications. MFCs may operate better at a smaller scale, when anaerobic digestion would suffer from the high costs of gas treatment and handling; and with a more dilute waste stream such as, for example, the effluent from an anaerobic process. Furthermore, the challenge of wastewater complexity is yet to be addressed for real-scale applications. More studies using real wastewater are required to improve the knowledge of the degradation pathways of complex substances. Laboratory MFCs fed with wastewater mainly convert readily biodegradable organics, whereas more complex materials generally pass straight through the system. This can be due to the generally short hydraulic-retention time (Rabaey et al., 2005b) and the slower conversion rate of the more recalcitrant material.
4.18.10.2 Nitrogen Removal Nitrogen removal represents a topic of particular interest for MFC application. Currently, nitrogen is removed from wastewater by means of two sequential processes, both promoted by microorganisms: nitrification and denitrification. Nitrification is an autotrophic process that converts ammonium into nitrate using oxygen and an inorganic carbon source. Denitrification is instead a heterotrophic anoxic process that utilizes nitrate as an electron acceptor during the oxidation of an organic carbon source. Due to the competition between aerobic and anoxic organisms for the available organics, supplementary carbon supply is often used (typically methanol) in addition to the carbon already present in the wastewater, to increase the efficiency of denitrification. Cathodic denitrification was recently demonstrated in MFCs (Clauwaert et al., 2007a). In Virdis et al. (2008), the MFC was integrated with an external aerated vessel for nitrification, and the system was able to simultaneously remove carbon and nitrogen. As in the MFC configuration, the oxidative and reductive biomasses are kept physically separate by the IEM, and the competition between organisms can be minimized to achieve highly efficient denitrification at lower C/N ratios than generally required by heterotrophic denitrification. This represents an important advantage of MFCs as denitrification is driven by electrons directly supplied by the anode with no need for the organics to be added directly into the denitrification stage.
4.18.10.3 Bioremediation The possibility of removing metals and chlorinated compounds by means of bioelectrochemical systems is of particular importance for applications of this technology in bioremediation. A study by Gregory and Lovley (2005) reported the microbial reduction of uranium using an electrode as an electron donor which caused the conversion of soluble U(VI) into the rather insoluble U(IV), which precipitated onto the electrodes. Cathodic reduction of perchlorate, an industrial by-product (i.e., from the production of pyrotechnic compounds and lubricant oils) found in the environment due to a historical lack of regulation in its manufacturing and discharge, was also recently described (Thrash et al., 2007). Hydrogen likely served as an electron shuttle for dissimilatory perchlorate-reducing bacteria, although an isolate from a perchlorate-reducing reactor could accept electrons from the cathode via an added redox mediator. Reductive dechlorination of chlorinated compounds, such as trichloroethene (TCE), is typically achieved through the oxidation of an organic electron donor. It was recently shown that TCEdechlorinating bacteria could directly use a cathode as an electron donor (electrode polarization: 450 mV vs. SHE) (Aulenta et al., 2009).
4.18.10.4 H2 Production Hydrogen gas can effectively be produced through MECs. These are electrolysis-type BESs that are capable of producing hydrogen at the cathode on applying a small voltage (40.2 V in practice), while oxidizing organic matter at the anode (Liu et al., 2005c; Rozendal et al., 2006b).
Microbial Fuel Cells
The architecture of MECs is nearly identical to that of MFCs, except for the fact that an MEC requires gas collection at the cathode. Cathodic hydrogen production on plain carbon electrode is very slow due to high overpotentials. Platinum has been the most commonly used catalyst (Rozendal et al., 2006b). However, it was recently discovered that bacteria could be effectively used as catalysts for hydrogen production (Rozendal et al., 2008b), thus overcoming the disadvantages connected to the use of platinum-based cathodes (Section 4.18.5). More recently, the use of nickel and stainless steel in the form of flat sheets or brushes was observed to outcompete platinum as cathodic catalyst (Call et al., 2009; Selembo et al., 2009). MECs are a promising technology for sustainable hydrogen production from wastewater. While MFCs recover energy from wastewater in the form of electricity, MECs recover energy in the form of hydrogen. Nevertheless, to function as wastewater treatment systems, MECs need to guarantee reasonable COD conversion rates. Based on current H2-production performances, COD-loading rates would need to be of the order of 6.5 kg COD m3 d1 (Logan et al., 2008), which is between the range of activated sludge systems and anaerobic digesters, thus making the MEC technology competitive when compared with traditional wastewater treatment. However, MECs need to be more cost effective than existing technologies and since electric energy is consumed during their operation, the higher costs must be compensated for by sufficient hydrogen production. It is estimated that full-scale MEC systems require 1 kW h1 m3 of H2 produced and can produce up to 10 m3 H2 m3 d1 (Rozendal et al., 2007), which is equivalent to an energy requirement of B1.5 kW h1 kg1 COD treated (Logan et al., 2008), and which is similar to the energy consumption for activated sludge treatment (Rozendal et al., 2008a). On the contrary, energy recovery through anaerobic digestion does not require significant energy inputs. However, compared to MECs, anaerobic digestion produces a gas (methane) that is less valuable than hydrogen. On the other hand, anaerobic digestion is a well-established technology, whereas microbial electrolysis requires great research efforts on both engineering and biochemical aspects.
4.18.10.5 Bioelectrochemical Production of Value-Added Chemicals Currently, it is expected that the capital costs for a full-scale BES will always remain several times higher than that of conventional wastewater treatment systems (Rozendal et al., 2008a). Therefore, bioelectrochemical wastewater treatment will become economically advantageous when the larger investments are compensated for by the larger value of the products obtainable. Electricity production using MFCs has the disadvantage of its low revenue, which puts electricity among the least valuable products (Rozendal et al., 2008a). As we have seen in Section 4.18.10.4 energy can be recovered in a BES not only as electricity but also as hydrogen. In addition, BESs can offer other interesting opportunities to improve their economical feasibility. For instance, the hydrogen produced in a BES can be used to create other products in situ. Several researchers have already reported methane production as a side product in membrane-less MECs, due to hydrogen
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scavenging (Call and Logan, 2008; Clauwaert and Verstraete, 2009). When the IEM is omitted, hydrogenotrophic methanogens can use the hydrogen produced at the cathode and combine it with the carbon dioxide produced at the anode, thus producing methane. Even though methane has a lower energy content compared to hydrogen per unit of mass, removing the IEM from the MECs would significantly lower its capital costs as well as reduce the system’s ohmic losses and pH gradients. MECs could thus be used in combination with anaerobic digestion facilities at the polishing stage by treating the residual organics present in the effluent (Clauwaert and Verstraete, 2009). In addition, direct methane production without intermediate hydrogen production was also recently observed in a biocathode dominated by Methanobacterium palustre (Cheng et al., 2009), demonstrating that BESs can be used to convert electricity into a biofuel while also capturing carbon dioxide. It is expected that, in future, BES innovations will proceed on these lines. A whole range of value-added chemicals requires the reduction of power for their production. When CO2 and O2 impurities are present together with H2, the production of biopolymers such as polyhydroxyalkanoates (PHA) by hydrogen-oxidizing bacteria can be foreseen in membrane-less or loop-based MECs (Ishizaki et al., 2001). Moreover, alcohols can also be produced from VFAs using hydrogen as an electron donor (Steinbusch et al., 2008), or a mediator (methyl viologen) (Steinbusch et al., 2009). Moreover, hydrogen peroxide production has been obtained by coupling organic oxidation at the anode with oxygen reduction at the cathode and adding a small voltage (Rozendal et al., 2009).
4.18.11 Outlook Current approaches to waste management will have to change in the future since waste will have to be considered as an alternative resource rather than an inconvenient burden to dispose of. Wastewater, in particular, represents an important resource of nutrient (primarily nitrogen and phosphorus), energy (as energy contained in chemical bonds of organic matter), and water itself. In a sustainable society, wastewater treatment will no longer be regarded as a treatment per se, its sole purpose being the removal of contaminants, often requiring a great deal of nonrenewable energy (e.g., from coal extraction), which may ultimately cause more environmental damage than the direct discharge of the untreated wastewater. In the future, we will no longer refer to wastewater treatment plants but rather to bioelectrochemical-resource-recovery plants, or biorefineries. With this new picture emerging, the raw wastewater would follow several sequential treatment stages, the first stage of which would be a pre-treatment to remove the solids, for instance, through dissolved air flotation. The solid fraction can then be sent to an anaerobic digester wherein some biogas is formed and solid liquid/separation creates a sludge that can be used for composting, while the supernatant can be sent back to the main flow. A pre-fermenter would be likely added after the pre-treatment to breakdown complex organics and produce an effluent richer in VFAs that are better metabolized in a BES anode for electron extraction. After the anodic
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passage, the effluent would be primarily rich in nitrogen as ammonium, which can be recovered through thermal volume reduction, or as struvite. Further specific treatments would depend on the final utilization of the effluent; tertiary treatments will produce water suitable for use in other processes (as cooling water, for instance), or even be able to reach drinking standards through advanced treatment processes. The electrons harvested during the anodic passage would be conveyed to the cathodic side of the BES where they can be used to drive a wide array of processes, from direct electricity production through to MFC, or perhaps to produce hydrogen through MECs, or moreover to produce other value-added chemicals such as methane, hydrogen peroxide, alcohols, biopolymers, or biofuels as seen previously.
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4.19 Water in the Pulp and Paper Industry H Jung and D Pauly, Papiertechnische Stiftung, Munich, Germany & 2011 Elsevier B.V. All rights reserved.
4.19.1 4.19.2 4.19.2.1 4.19.2.2 4.19.2.3 4.19.3 4.19.3.1 4.19.3.2 4.19.3.2.1 4.19.3.2.2 4.19.3.2.3 4.19.3.2.4 4.19.3.2.5 4.19.3.2.6 4.19.3.3 4.19.3.3.1 4.19.3.3.2 4.19.4 4.19.4.1 4.19.4.2 4.19.4.2.1 4.19.4.2.2 4.19.4.2.3 4.19.4.3 4.19.4.3.1 4.19.4.3.2 4.19.4.3.3 4.19.5 4.19.5.1 4.19.5.2 4.19.5.3 4.19.6 References
Overview of Pulp and Papermaking Water in the Pulp and Paper industry Functions of Water in Papermaking Historical Evolution of Water Systems Current Water Consumption Levels in the Pulp and Paper Industry Water Use Freshwater Process Water Circuitry Primary, secondary, and tertiary water circuits Detrimental substances General principles of circuitry Closed water circuits Assessment of freshwater use and circuitry Wastewater Characterization of wastewater from the pulp and paper industry Wastewater discharging Water Treatment Freshwater Treatment Circuit Water Treatment Objectives of circuit water treatment Mechanical circuit water treatment Advanced circuit water treatment Wastewater Treatment Preliminary mechanical treatment: Mechanical processes for removal of solids Biological treatment Advanced and tertiary treatment Potentials and Limits of Water Saving Limiting Effects of System Closure Heat Balance Economic Benefits Improving Water Efficiency in Paper Manufacturing Industries – 30 Years of Success
4.19.1 Overview of Pulp and Papermaking Paper is currently a commodity product. The worldwide consumption of paper is growing steadily and it is hard to imagine the world without paper. Papermaking is based on a principle that is roughly 2000 years old. Today, in principle, the same process steps, which were used in the past, are included. The papermaking process can be divided into four main process steps (Figure 1), which can be either integrated at one site or located at several different sites. These process steps include
• • • •
pulp production (chemical or mechanical pulp and recycled fiber pulp (RCF)), stock preparation, paper machine, and coating and finishing.
Papermaking starts with the provision of the stock components such as fibers, fillers, and chemical additives. Primary
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fibers (chemical and mechanical pulp) are obtained from wood and annual plants by chemical pulping or mechanical defibration. Secondary fibers are produced from recovered paper. All these components have to be properly prepared for optimum use in papermaking. Stock preparation is followed by the approach flow system, which links stock preparation to the paper machine. Paper or board is produced at the paper machine. In doing so, a sheet is formed from a highly diluted fiber suspension and dewatered by means of filtration, pressing, and thermal drying. Coating and calendaring improve the surface quality of the paper and board. The final steps include slitting, sheeting, and packaging of the final product for shipment.
4.19.2 Water in the Pulp and Paper industry 4.19.2.1 Functions of Water in Papermaking Water is one of the key components in pulp and papermaking. Without water, the production of paper would be unthinkable.
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Water in the Pulp and Paper Industry Raw material Wood / recovered paper / annual plants
Pulp production Chemical / thermal / mechanical
Pulp
Stock preparation Suspending / screening and cleaning / refining
Machine stock
Papermaking process Approach flow / paper machine
Paper / board Coating / finishing
Final product Figure 1 Process steps in papermaking.
Water performs numerous functions in the papermaking process. It is used as medium for suspension, dwelling, and transfer processes, and it serves to separate as well as to restore the bonds between fibers. Other uses include showers in the wire and press sections, sealing of pumps, and cooling and cleaning purposes. Furthermore, water in the form of steam is used as an energy carrier.
4.19.2.3 Current Water Consumption Levels in the Pulp and Paper Industry At the onset of industrial papermaking, paper was produced with high specific water consumption. The pulp and paper industry has improved the processes in the last few decades for economical and ecological reasons and, as a result, was able to reduce water consumption significantly. This was only possible because of increasing closure of in-mill water circuits and consistent reuse of clarified process water by former freshwater consumers. A survey conducted by the Papiertechnische Stiftung (PTS) and the German Pulp and Paper Association (VDP) showed that the average specific effluent volumes of Germany’s pulp and paper industry decreased from 46 to approximately 10 m3 per metric ton of product produced between 1974 and 2007 (Figure 2). Nevertheless, the German pulp and paper industry remains one of the six biggest consumers of industrial water (Federal Statistical Office, 2008). The consumption level in the different pulp and paper mills can vary because of both general and process-related reasons such as raw materials used, paper grades produced, and plant structure. Furthermore, local boundary conditions, such as requirements on wastewater discharge, have an impact on the consumption level. High specific effluent volumes occur particularly in specialty paper grades. These mills are often faced with structural handicaps that cause increased specific effluent volumes: small and obsolete paper machines, low production rates, frequent grade changes, and often very high quality requirements on the final product. The lowest water requirements can be found in mills that produce packaging papers, such as corrugated base paper or board. Some of these mills have already managed to close their water circuits completely, resulting in a zero effluent production.
4.19.3 Water Use 4.19.3.1 Freshwater
4.19.2.2 Historical Evolution of Water Systems According to Zippel (2001), there are three phases in the historical evolution of paper-mill water systems. Phase 1 began in the 1920s. During this phase, the basics of water circuit design were established. Freshwater saving potentials were initially introduced predominantly for economic reasons. Phase 2 began in the 1960s. During this phase, the final effluent became more important for paper mills. This was caused as a means of reducing solid losses and thereby increasing the yield on the one hand and, on the other hand, as a result of the increasing ecological awareness of the general population. Subsequently, mechanical and biological wastewater treatment plants were installed. A few mills even managed to close their water circuits completely. Phase 3 (starting in the 1970s) was marked by initial attempts undertaken to deal with the consequences of system closure. Thus, the third phase was characterized by basic investigational work on the constituents of the process water and their impact on runnability of the paper machine and paper quality.
Depending on the availability and local conditions, either surface water or groundwater is used as freshwater. Drinking water is used for certain purposes, such as trim squirts. In the German pulp and paper industry, roughly 80% of the fresh water is taken from surface waters (Jung et al., 2009). In stateof-the-art mills, there are only few freshwater consumers. Typical freshwater consumers include
• • • •
high- and low-pressure showers for felt conditioning and wire cleaning, trim squirts, sealing water for liquid-ring vacuum pumps and packing glands, and additive preparation and dilution.
In view of the limited freshwater volume available, it must be used efficiently. Hence, freshwater used for cooling purposes (oil coolers and steam condensers) should be collected and reused as fresh warm water in the paper machine. Process water should be used for all other purposes, such as stock dilution, consistency control, or cleaning.
Water in the Pulp and Paper Industry
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Specific effluent volume (m3 per metric ton of product)
50
40
30
20
10
0 1972
1976
1980
1984
1988
1992
1996
2000
2004
2008
Year Figure 2 Averaged specific effluent volume in the German pulp and paper industry (Jung et al., 2009).
Typically, approximately 40% of the entire freshwater volume is used for the high- and low-pressure showers in the wire and press sections. Depending on the paper grade and the nozzles used, different flow rates are used for showers in the wire and press sections. The entire consumption in the European paper industry for both showers in the wire section and showers in the press section averages out to approximately 1.0–2.5 m3 per metric ton of paper, depending on the degree of water circuit closure (Kappen et al., 2004). The sealing water consumption in liquid-ring vacuum pumps is highly dependent on the installed system. If a sealing water circuit is installed with a cooling tower, the freshwater can be less than 0.5 m3 per metric ton of paper. Without a sealing water circuit, the consumption typically amounts to approximately 4–5 m3 per metric ton of paper (Kappen et al., 2004). Sealing water is needed in packing glands to lubricate the sealing faces and to remove solids. According to Kappen et al. (2004), freshwater requirements amount to 0.15 m3 h1 in pumps and agitators and 0.2 m3 h1 in refiners and deflakers.
4.19.3.2 Process Water 4.19.3.2.1 Circuitry In papermaking, it is quite important to provide both adequate water quality and the required volume of water for every single consumer. Using freshwater for all purposes would consume several 100 m3 per metric ton of paper. The objectives of the water circuit system are to provide the required amount and quality of water for every consumer paying attention to economical and, at the same time also, to ecological aspects. In meeting these requirements, most of the water used in the pulp and paper industry is process water that has been recycled in different loops. Hence, the installation and proper design of water circuits are of fundamental importance for pulp and papermaking, since it contributes to enhanced product quality and reduced effluent volume.
Process water is mainly used for pulping and consistency control in the individual process steps. It is also used to a greater extent for purposes for which freshwater was formerly used, such as (low-pressure) showers, foam destruction, sealing water of liquid-ring vacuum pumps, or additive preparation. Process water is produced in the thickening and dewatering stages of the papermaking process by separating liquid phase from solid phase. In the stock preparation loops, this is done by disk filters, screw and double wire presses, and drum thickeners. Wire section, press section, and savealls provide the required process water volumes at the paper machine. If water quality achieved is still inadequate, advanced treatment technologies such as membrane or ozone treatment can be employed. The possibilities for designing water circuits greatly vary and depend on a number of parameters. One important parameter is the grade of paper being produced and the corresponding raw material being used.
4.19.3.2.2 Primary, secondary, and tertiary water circuits Based on the connection to the core process (sheet formation on the wire), it is generally possible to differentiate between three categories of water circuits: primary, secondary, and tertiary water circuits. Figure 3 illustrates the primary and secondary water circuits. The primary circuit consists of white water 1 originating from the wire section. This circuit is the largest as far as the volumetric flow rate is concerned. The circulating flow rate depends on the retention in the wire section and the consistency in the headbox. Its objective is to dilute the main stock flow after the machine chest in the approach flow system to a consistency of approximately 0.7–1.5%. The excess flow rate is part of the secondary circuit. Besides the excess flow rate of white water 1, the secondary circuit originates from the forming section and from the press section (see Figure 3). Most of this water is preferably fed to a saveall, and the recovered fibers are sent to the blend or machine chest and stock preparation, respectively. The clarified
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Water in the Pulp and Paper Industry Mixing chest
Machine chest
Cleaner
Screen
Wire section
Press section
From stock preparation (DIP, chemical pulp, mechanical pulp, fillers, etc.
Primary circuit
White water 1
Showers, etc.
To stock preparation
White water 2
Secondary circuit Saveall
Waste water
Figure 3 Schematic illustration of principal water and stock flows in a paper machine (Hutter, 2008).
water is sent to a buffer tank and from there it is supplied to the process. The possible fields of application of clarified water are manifold: pulping, consistency control, foam destruction, and showers (mainly in the wire section). Further treatment (e.g., membrane filtration) might be necessary in the case of sensitive applications such as sealing waters or high-pressure showers. As the papermaking process is typically supplied by freshwater, there is always an excess of process water. This excess water is part of discharged wastewater. A tertiary circuit is required when, at least, a part of the treated wastewater is recirculated. In zero effluent systems, all treated wastewater is recirculated. In order to eliminate detrimental substances from the papermaking process, the recirculated wastewater should undergo full biological treatment. Possible fields of application of the recirculated wastewater are manifold and depend on the water quality attained. Besides being used as pulping or cleaning water, it may also be used as sealing water or as spraying water in showers after adequate pretreatment. Attention must be drawn to the danger of scale formation as biologically treated water often has a high calcium concentration (Demel et al., 2004a, 2004b).
• • • • • • •
a reduction in additive efficiency, a reduction in optical and strength properties, negative effects on drainage and paper drying, negative impacts on sizing, odor formation, deposits, and/or foam generation.
The main sources of detrimental substances and contraries in paper-mill process water are fibrous raw materials, additives, and freshwater (Negro and Tijero, 1998). Table 1 provides an overview of the composition and origin of detrimental substances. The content of detrimental substances is typically measured using sum parameters, such as anionic trash, cationic demand, or chemical oxygen demand (COD). Inorganic dissolved substances are measured as increased conductivity (Stetter, 2006). The COD denotes the volume of oxidizable substances in a water sample. It is considered to be balanceable and is thus a suitable optimization parameter.
4.19.3.2.4 General principles of circuitry 4.19.3.2.3 Detrimental substances Due to the increasing use of recovered paper and the reduced freshwater consumption, constituents known as detrimental substances have accumulated in water circuits, leading to growing problems in the papermaking process. Detrimental substances are substances that have a negative impact on the papermaking process and on product properties. Auhorn defined them as follows: "Detrimental substances are dissolved or colloidally soluble anionic oligomers or polymers and nonionic hydrocolloids" (Auhorn, 1984). They can result in
Figure 4 schematically illustrates a simplified water and stock system in a paper mill. Both the stock preparation loop and the paper machine loop can use freshwater. Wastewater is discharged mainly from the paper machine loop. Moreover, water is exchanged between the paper machine and stock preparation loops depending on the transfer consistency of the pulp coming from stock preparation. Based on a specific effluent volume of 10 m3 per metric ton of paper and a COD input of 10 kg per metric ton of raw material, this results in a COD concentration of 1.7 g l1 in the stock preparation loop and 1.2 g l1 in the paper machine loop.
Water in the Pulp and Paper Industry
There are two general principles used in designing water circuits that are described on the basis of this simplified model mill:
• •
loop separation and countercurrent arrangement.
Water circuits can be subdivided into separate loops by installing thickening units such as screw presses or double wire presses. In many cases, these thickening units are also necessary for downstream units such as dispergers. At the same time, soluble detrimental substances will be retained in the stock preparation water system. An increase of up to 30% in the transfer consistency between stock preparation and the Table 1 contraries
Composition and origin of detrimental substances and
Chemical compounds
Origin
Sodium silicate
Peroxide bleaching, deinking, recovered paper Filler dispersing agent Filler dispersing agent Coated broke, recovered paper Freshwater Chemical and mechanical pulp
Polyphosphate Polyacrylate Starch Humic acids Lignin derivates, lignosulfonates, hemicelluloses Fatty acids Volatile fatty acids
Chloride Calcium Sulfides Exopolymer saccharides
Mechanical pulp, deinking Anaerobic processes (high hydraulic retention times, spoiled recovered paper) Chemical additives Recovered paper, fillers Anaerobic processes, sulfate High C/N ratio
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paper machine results in a significant reduction in the exchanged water volume. As the wastewater is still being discharged from the paper machine loop, there is no sink for the detrimental substances in the first loop (stock preparation). Detrimental substances build up in the stock preparation loop, resulting in a COD concentration of 4.9 g l1. It is not possible to relieve the paper machine loop (Figure 5). A countercurrent arrangement (Figure 6) completes the above-described principle of loop separation. The highly concentrated filtrate from the thickening unit is discharged to the wastewater treatment plant. The water deficit in the stock preparation loop is compensated by adding water from the paper machine loop. Hence, the most contaminated water is being discharged, while the better-quality water is being used in the more sensitive paper machine loop. The water flows in a direction opposite to the stock flow. This leads to significant relief of the paper machine loop, resulting in a COD concentration of 0.5 g l1 or 58% of the initial situation described above. Strict separation of the stock preparation water loops from the paper machine loop combined with a well-designed countercurrent arrangement is essential to meet high runnability and quality requirements because this strategy keeps detrimental substances out of the paper machine. Unlike the countercurrent dewatering arrangement in paper mills, a countercurrent washing arrangement is typically installed in chemical pulp mills. Substances that are dissolved during digestion, delignification, and bleaching are carried along into the next process steps together with the fibers. To accumulate and recirculate these substances to the digester, the washing liquor passes through a countercurrent washing arrangement within the different process steps (Figure 7). This ensures that most of the organic load and the digesting chemicals are recirculated to the digester, which guarantees the efficiency of the bleaching chemicals. Furthermore, it helps to
5 COD SP loop (g l−1)
Raw materials 10 kg COD t−1 Freshwater Stock preparation
4 3 2 1
20 l kg−1 0 5
Paper machine 10 l kg−1 Wastewater treatment plant
Paper Figure 4 Simplified schematic illustration of the water and stock systems of a paper mill.
COD PM loop (g l−1)
5%
4 3 2 1 0
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Water in the Pulp and Paper Industry 5 COD SP loop (g l−1)
Raw materials 10 kg COD t−1 Freshwater Stock preparation
4 3 2 1
20 l kg−1 0 5
Paper machine 10 l kg−1 Wastewater treatment plant
COD PM loop (g l−1)
30%
Paper
4 3 2 1 0
Figure 5 Loop separation.
5 COD SP loop (g l−1)
Raw materials 10 kg COD t−1 Freshwater Stock preparation
4 3 2 1
10 l kg−1
2 l kg−1 10 l kg−1 Wastewater treatment plant
Paper
COD PM loop (g l−1)
Paper machine
0 5
8 l kg−1
30%
4 3 2 1 0
Figure 6 Countercurrent arrangement.
relieve the pulp dewatering machine from detrimental substances as efficient as possible (Borschke, 2006).
4.19.3.2.5 Closed water circuits Complete closure of water circuits implies eliminating any sort of effluent discharge. For some mills, it is the last resort to be able to continue production at that particular location. Motivating factors include costs of discharging effluents, absence of receiving waters, or the necessary discharge rights if the mill is moved to a new location. The specific effluent volume in the case of closure is 0 m3 per metric ton of paper. Freshwater is
used only to compensate for the loss of water by evaporation and in the finished product, and for the water removed together with the rejects. This volume normally amounts to approximately 1.5 m3 per metric ton of paper. The process is thus subject to massive limitations. Only few consumers can continue to be supplied with freshwater, leading to extremely high concentrations of detrimental substances which will be bled out of the system only by transferring them into the paper. Hence, when preparing water circuits for mill closure, first, all options for optimization of the water circuits must be exhausted. A subsequent installation of internal circuit water
Water in the Pulp and Paper Industry Deknotting and screening
Washing
Washing
Oxygen Washing delignification
Pulp from digester
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Liquor Subsequent process steps
O2
Filtrate
Digester Evaporation plant
FT
FT
FT
FT
Filtrate tank
Filtrate tank
Filtrate tank
Filtrate tank
COD in the water circuit
Figure 7 Countercurrent washing in chemical pulp manufacturing (Borschke, 2006).
2. Circuit water treatment
1. Optimization of water circuits
Closure of water circuits
most delicate part of the papermaking process, which is why the white water should contain as few disturbing substances as possible. By comparing the COD levels (filtered samples) in the wastewater prior to biological treatment and in white water 1, the K1 value makes it possible to determine the utilization of freshwater (Equation (1)). K1 value significantly less than 1 indicates freshwater which is discharged to the effluent treatment plant directly without relieving the paper machine loop:
K1 ¼ Specific effluent volume
CODEffluent CODWhite water
ð1Þ 1
Figure 8 A stage-by-stage approach for water circuit closure.
treatment units makes it possible to remove dissolved substances (kidney technology). Only then subsequent closure can be met successfully (Figure 8). Possible kidney technologies include integrated biological treatment, membrane filtration, or ozone treatment (see Section 4.19.4.2).
4.19.3.2.6 Assessment of freshwater use and circuitry In order to be able to optimize a water circuit, it must first be clarified whether or not the following conditions are fulfilled:
• • •
freshwater should be used effectively and not passed directly to the wastewater treatment plant, the contaminant load at the paper machine should be as low as possible, and contaminants should be discharged wherever possible using the smallest effluent volumes.
The K-values established by Kappen (Kappen and Wilderer, 2002) are capable of quantifying the most important goals in optimizing circuit design. They work through comparisons of the COD levels at different locations in the water circuit. K1 value. Sheet formation is brought about by dewatering fiber suspensions in the paper machine followed by the subsequent formation of hydrogen bonds between fibers. It is the process stage that is decisive for mechanical and optical characteristics of the paper. Sheet formation constitutes the
K2 value. The K2 value expresses the COD concentration ratio in water loops of the stock preparation and paper machine (Equation (2)). To achieve maximum relief of the paper machines, the COD level in white water 1 ought to be substantially lower than that of the stock preparation system. K2 41 means that detrimental substances that give rise to COD are retained in the stock preparation system, that is, the sheet formation section and white water 1 are relieved:
K2 ¼
CODStock preparation CODWhite water 1
ð2Þ
K1/K2 ratio. K1/K2 indicates whether the wastewater discharged from the papermaking system is the optimum solution in terms of paper machine relief. To obtain a maximum COD relief through a minimum effluent flow, the water highest in COD loading must be discharged to the wastewater treatment plant. This maximum loading is found in the stock preparation system where detrimental substances accumulate. The quotient of COD levels in the wastewater and stock preparation system is referred to as the K1/K2 ratio:
K1 CODEffluent ¼ K2 CODStock preparation
ð3Þ
K1/K2 close to 1 indicates that the COD loadings of the effluents and stock preparation are nearly equal, that is, the
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Water in the Pulp and Paper Industry
effluents are discharging detrimental substances mainly from the section with the highest COD loading. This, in turn, ensures maximum paper machine relief by the given effluent volume and corresponds to the countercurrent arrangement. The K1/K2 ratio helps to evaluate the circuitries of a paper mill. A countercurrent arrangement is only realized for K1/K2 values close to 1. The above definitions apply exclusively to simple systems with one stock preparation system and one paper machine. In more complex systems, characterized by more than one line or loop in stock preparation or more than one paper machine, the COD loading of the wastewater, stock preparation, and paper machine loop are calculated as the weighted average of the individual COD loadings determined. Graphic representation of K1 and K2. The relationship between these K-values may be visualized in a K1–K2 performance characteristic (Figure 9). The K1–K2 performance characteristic depicts the current situation in a paper mill (operating point). The two lines to the left and right of the diagonal define the target range of best performance achievable under practical conditions. An operating point within the target range indicates an optimized water loop. The circuitry status and the optimization potential that exists can be visualized as the K1–K2 section that has experienced a local shift out of the target range. The success of optimization measures can be documented without much difficulty using these key parameters. The K1 value, the K2 value, and the K1/K2 ratio are the principal characteristics that make assessment of the efficiency of freshwater use and circuitry possible. The characteristics K1 and K2 make it possible to quantify the primary objectives of circuit optimization, that is, effective freshwater use, maximum paper machine relief, and effective elimination of anionic trash (Kappen and Wilderer, 2002).
4 Target area with loop separation 3
K2
Operating point 2
1 Target area without loop separation 0 0
1
2 K1
Figure 9 K1–K2 performance characteristic.
3
4
4.19.3.3 Wastewater 4.19.3.3.1 Characterization of wastewater from the pulp and paper industry In general, wastewater in the pulp and paper industry is produced in the form of excess process water, which is displaced by the freshwater input. The wastewater is loaded primarily with organics that enter the production process together with raw materials and additives. Effluents from the pulp and paper industry are still not completely understood in terms of their chemical composition (Hynninen, 2000). In the majority of cases, however, wastewater of paper mills is nontoxic and easily degradable biologically. Higher concentrations of dissolved organic and inorganic compounds are observed in productions, enabling a particularly intensive utilization of water (Mo¨bius, 2002). As far as organic loads are concerned, COD, biochemical oxygen demand (BOD5), and adsorbable organic halogens (AOX) are the key parameters that characterize papermaking effluents. Today, however, the total organic carbon (TOC) parameter is becoming more important – a development which is reflected in a growing number of measuring methods such as the cuvette test or online measuring systems. Effluent concentrations vary widely depending on
• • • •
raw materials, paper grades, specific freshwater consumption, and available installations.
The assessment of the biodegradability of effluents is based on parameters such as BOD5, COD, and their ratio in wastewater. For a completely degradable compound such as glucose, which resembles the dissolved material in paper-mill effluents, the BOD5/COD quotient is typically approximately 0.6, suggesting a very good biodegradability, whereas a lower quotient is indicative of poorer degradability and a higher residual COD. The BOD5/COD quotient from different paper mills varies roughly between 0.35 and 0.5. Moreover, some important inorganic effluent parameters, such as salt loads, have to be taken into consideration. Calcium and sulfate concentrations play a special role in the operation of anaerobic treatment plants. When treating effluents with a high calcium content, poorly soluble calcium carbonate may be precipitated. In plants employing carrier material, such precipitation products may cause deposit formation. In mixed reactors, precipitation products tend to accumulate in the sludge, impeding thorough mixing of effluents and sludge and finally reducing the share of active biomass. When treated effluents are recirculated back into production, additional precipitation problems may arise in the consumers due to the pH shift of the decreased buffer capacity. The growing use of calcium carbonate as a filler and coating pigment and the ever-tighter-closed water circuits increase the calcium concentrations in circuit water and the wastewater. This applies in particular to mills that convert recycled paper. In the case of high sulfate concentrations, anoxic conditions may trigger sulfate reduction and lead to sulfide formation. This may disturb the degradation processes (methanogenesis) in anaerobic treatment plants, whereas in
Water in the Pulp and Paper Industry
aerobic biological treatment such high sulfate concentrations may foster the growth of undesirable filamentous microorganisms. As an additional drawback, the hydrogen sulfide that forms may give rise to bad odors and corrosion phenomena. Sulfate concentrations up to 600 mg l1 are to be expected in the wastewater of paper mills producing mechanical paper due to the aluminum sulfate used for resin sizing. The sulfate concentrations are substantially lower for woodfree papers. Even higher concentrations can occur in the production of recycled fiber-based papers. Sulfate originates from recovered papers and becomes increasingly concentrated as a result of tightly closed water circuits, typical for paper mills converting recovered paper. Depending on the treatment process used, other parameters such as pH, conductivity, and temperature are important for operational safety. Normally, in paper industry effluents, phosphorus and nitrogen compounds serving as nutrients for microorganisms are either absent or only available in insufficient quantities (Hamm, 2006). Therefore, it must be ensured that dosages of nutrients in treatment plants provide a sufficient supply for the microbiota. However, simultaneously, the permissible limit values in final effluents have to be met.
4.19.3.3.2 Wastewater discharging In the German pulp and paper industry, most effluents undergo full biological treatment. Ninety-five percent of the production volume is produced in mills with an integrated biological wastewater treatment plant or mills that discharge their wastewater to municipal wastewater treatment plants; 4% of the annual production volume is produced in mills with a closed water circuit; and only 1% of the production volume comes from mills that discharge their effluents without biological treatment (Jung et al., 2009).
4.19.4 Water Treatment 4.19.4.1 Freshwater Treatment As mentioned in Section 4.19.3.1, the source of freshwater in the pulp and paper industry is usually surface water. Typically, freshwater does not meet the required quality parameters of the manufacturing process and therefore has to be treated. Well water seldom needs treatment. Objectives of freshwater treatment in the pulp and paper industry include
• • • •
removal of solids, removal of color and organic substances, decrease in hardness and removal of other dissolved salts, and, in some cases, the disinfection of the water.
Water quality can be improved by a range of treatment measures. Factors influencing the choice of the treatment method and equipment include required water quality, water volume to be treated, space available for freshwater treatment plant, and, to some extent, how well the plant operation and supervision can be integrated with the other operations in the mill (Hynninen, 2000).
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Predominantly mechanical or chemical–mechanical treatment technologies are used for freshwater treatment in the pulp and paper industry. According to a survey conducted by PTS and VDP, more than 90% of the surface water used as freshwater for the papermaking process is treated by filtering. An additional 75% of this water is conditioned by chemical coagulation, flocculation, and subsequent sedimentation. The volume of freshwater treated with biocides has increased significantly, whereas the use of chlorine has decreased in the past few years (Jung et al., 2009). Freshwater is softened and desalinated for boiler house use and for the production of some specialty papers (e.g., photographic base paper or cigarette paper; Stetter, 2006).
4.19.4.2 Circuit Water Treatment 4.19.4.2.1 Objectives of circuit water treatment At the beginning, the objective of circuit water treatment was primarily to recover fiber furnish from papermaking effluents. Under the economic and ecological necessity of reducing effluent volumes and loads, the circulation water treatment process took on ever-greater importance and function: circuit water treatment must provide clarified water with a predefined quality and has to remove interfering substances from the system. In doing so, circuit water treatment became responsible for removing not only insoluble and colloidal components but also dissolved substances. Therewith, circuit water treatment helps to stabilize production processes and ensures product quality. The objectives of circulation water treatment include
• • •
recovery of raw materials, production of mill water with a low solid concentration available, and reduction of contaminants in the circulation water.
Depending on the required water quality, the requirements on circuit water treatment vary from reducing solid losses in the case of relatively coarse treatment to preparing shower and sealing water in the high-pressure range in the case of precision treatment.
4.19.4.2.2 Mechanical circuit water treatment Sedimentation, flotation, and filtration methods in particular are employed in mechanical circuit water treatment. These techniques can also be used in combination with one another. Methods for screening and classification are mainly used in stock preparation. The market share of the individual types of savealls moves in the direction of a two-part system, as sedimentation is steadily declining, whereas filtration and flotation are both expanding due to innovations and technical improvements (Zippel, 2001). Hydraulic surface load, solid surface load, and purification performance are the key parameters for the layout of the treatment units. Other parameters that have to be considered are the concentration of suspended solids in the feed, presence of colloidal and dissolved substances, additive demand, available space, and overall energy consumption. Two of the most important factors are the investment and operational costs. Cost effectiveness is obtained by reducing raw material losses,
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Water in the Pulp and Paper Industry
increasing process performance, and the benefits arising from stable and efficient paper production (Weise et al., 2000). The fiber recovery unit, or synonymously the saveall, is supplied with the excess from the primary circuit (wire pit overflow) and water removed from the wire and press sections by the vacuum system. In rare cases, water from the floor channels or from the wet broke, for example, is supplied to circuit water treatment. The process connected to the saveall has to be designed so that a constant feed flow is maintained. Sedimentation. Sedimentation is generally the simplest form of a saveall, and conventional sedimentation savealls have long been known to be reliable and safe to operate. Nowadays, sedimentation plays a minor role and is commonly used only in old plants for circuit water treatment. A general disadvantage of sedimentation plants is a low density of the sediment. Hydraulic retention times are very long in some cases and can also provoke anaerobic degradation accompanied by the correspondingly disadvantageous consequences (odor, microbial contamination, etc.) that affect the entire water circuit. Long hydraulic retention times also become a problem if rapidly changing production programs are to be run on the paper machine. Flotation. Flotation denotes the use of air bubbles to float undissolved substances to the surface of a suspension. Hydrophobic or hydrophobized particles adhere to the air bubbles, rise through the suspension, and are carried along to the surface and scooped off there by a suitable skimming device. Different flotation processes vary according to how bubbles are introduced into the suspension. Dissolved air flotation (DAF) has established itself in circuit water treatment. In this process, water is supersaturated with compressed air and then supplied to the flotation chamber (Figure 10). The resulting reduction in pressure causes very fine air bubbles to form that become attached to the suspended particles. Pressure saturation current can be the entire inflow, a partial flow, or recirculated clarified water (recycling process). A general problem associated with flotation units is a sharp fluctuation in inflow loadings. Fluctuations both in the volumetric flow rate and in the solid surface loading produce poor results.
Filtration. Filtration technologies are well suited for separating solid particles from suspension with assistance of a porous filter medium. Compared with other processes, good separation properties and high-quality clarified water that can be achieved are advantageous. Disadvantages are high investment and operating costs due to the considerable amount of maintenance work. In the pulp and paper industry, disk filters (Figure 11) are by far the most common type for mechanical circuit water treatment. A disk filter comprises several disks that consist of individual segments covered with a filter medium that rotate in a vat. The filtrate consistency declines during the filtration process and the filtrates are typically collected separately as cloudy filtrate and clear filtrate. In some cases, super-clear filtrate may also be produced. As an alternative to disk filters, drum filters can be used, which usually reduces the cost factor. However, for most applications, the hydraulic capacity of drum filters is too low and only one filtrate quality is produced. This normally makes them unsuitable for saveall application. Drum filters are often used as simple but reliable thickeners, for example, in the broke-handling system (Zippel, 2001; Weise et al., 2000).
4.19.4.2.3 Advanced circuit water treatment As a result of an increasing closure of water circuits, the use of freshwater for a steadily growing number of consumers in a system must be restricted. In order to replace the freshwater at these locations, the clarified water must be of high quality. In many cases, complete elimination of solids is required, especially for showers in the high-pressure range and sealing water. Large volumes of clarified water needed in a closed or virtually closed system that at the same time places high requirements on clarified water quality for only very few consumers have promoted the use of multistage treatment processes. Methods for fine cleaning of pretreated circuit water are based on filtration. The objectives include continuously improving the water quality and serving as a police filter, if any upstream treatment method fails. A wide variety of methods are employed, including drum filters with very fine filter media
Rotary contact Spiral scoop
Clarified water pipes
Inlet distribution Rotating carriage ADT distribution Floated sludge outlet Recycle suction Raw water inlet Inspection window Settled material outlet
Clarified water outlet
Dissolved air in water inlet
Figure 10 Dissolved air flotation unit for circuit water treatment. Adapted from Krofta Waters International (2010).
Water in the Pulp and Paper Industry
Cake removal
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Cake removal
Filtrate
Suspension
outflow
inflow
Overflow
Suspension vat
Filter cake
Filter cake
Figure 11 Disk filter saveall system. Adapted from Wilichowski M (2009) Folien zur Vorlesung Mechanische Verfahrenstechnik I þ II. http:// www.mb.hs-wismar.de (accessed March 2010).
Process water
Freshwater
Paper and board production at 55−60 °C 100% recycled paper
Quality E
Quality D
Quality A
Quality C
Quality B
Biogas Acid
Buffer tank / sedimentation
pH
UASB/IC thermophil
Sedimentation
Aeration
Police filter
Solids
Solids Ozonation (opt.)
Heating
NF/ RO
Biogas Retentate Effluent
Figure 12 Kidney technology concept (Pauly, 2001).
(microfiltration), cartridge and backwash filters, sand filters, and membrane technology. The treatment units for elimination of dissolved substances accomplish important tasks in narrowing water circuits. They relieve the water circuits of detrimental substances and therewith avoid production restrictions due to limited freshwater capacities. The so-called kidney technologies, such as integrated biological treatment, softening, membrane technology, and ozonization, are promising approaches for obtaining effluent-free paper production by way of circuit closure. Combinations of treatment technologies make it possible to provide optimized solutions for different objectives. The decisions as to which concept and which treatment technology is to be used depend on specific boundary conditions. Figure 12 is a schematic view of a potential kidney concept for water circuit closure in a paper mill converting 100% recovered papers. Full circuit closure is not necessarily the solution of choice. Nevertheless, in many cases, advanced integrated treatment steps yield both economical and
ecological advantages (Pauly, 2002). Starting in 2008, the European Aquafit4Use research project focused on high waterreuse rates. The project also highlights maximum reduction in energy and chemicals, leading to more efficient use of limited resources by developing tailor-made treatment technologies and concepts (Pauly, 2008). Biokidney. Progressive system closure in a paper mill leads to increased concentrations of dissolved and colloidal compounds that in turn can give rise to increased microbial activity, slime formation, foaming, pitch disposition, corrosion, altered wet-end chemistry, and odor problems. Biological treatment of the circuit water can reduce or eliminate the buildup of troublesome compounds. Integrated biological treatment process is called the biokidney. Biological processes are state of the art in paper-mill wastewater treatment plants and are suitable for the elimination of biologically degradable substances, the reduction of sulfates, and the preliminary purification for nanofiltration and reverse osmosis.
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Water in the Pulp and Paper Industry
Flow
Conc.
Exhaust gas p
Inflow UF
Aeration
T pH
Sedimentation
T
Fi
Retentate overflow
Storage vessel
LC
Storage tank
Permeate vessel
Red.
Pellets
Exhaust air
UASB outlet
Separator
Permeate
Compressed air
Module
T Fi
p
Exhaust air
Insulated heated room
Outflow
Figure 13 Kidney concept – thermophilic anaerobic treatment and softening step – implemented at a Belgian paper mill in the frame of EU-project Paper Kidney; further downstream options: membrane treatment and ozonation (Pauly, 2002).
Biological treatment of process water has been carried out within a very wide range of operating conditions. Both anaerobic and aerobic process designs have been shown to be successful. Thermophilic treatment of process waters (Figure 13; Pauly, 2002) has a distinct advantage of eliminating the need for process water cooling and reheating for water recycling. A biokidney improves runnability due to a better overall quality of the process water caused by a decrease in soluble organic matter. Using biological treatment for COD reduction has allowed some packaging paper mills to operate with zero effluent systems (see, e.g., Stra¨tz, 2008; Herberz and Bahn, 2006). Practical trials conducted with aerobic and anaerobic laboratory-scale biokidneys also demonstrated that biokidneys have the potential to remove odorous compounds. (Jung et al., 2007). Membrane technology. This technology has experienced difficulties in trying to make headway into the papermaking sector. There are a variety of reasons for these difficulties, including scaling and fouling caused by high concentrations of salts and other detrimental substances, not fully developed concepts, and high investment costs due to high volumetric flow rates. A constant increase in the interest expressed by the European paper industry in membrane technology confirms that this method is due to evolve into a key technology for ¨ ller, continued water savings in the future (Simstich and O 2007). Membrane technology typically improves the quality of process waters substantially, since it removes suspended solids, microorganisms, and colloidal COD. Even salts can be separated out using reverse osmosis. Ultrafiltrated water is free of suspended solids and colloids. Bacteria, latex, and other micro-stickies are removed, for example. As anionic trash is cut to approximately half of its original level, the quality of the filtrate makes trouble-free recirculation back into the process possible (Sutela et al., 2006).
Fields of application for membrane technology in the pulp and paper industry are numerous and may offer many advantages, depending on treated partial flow such as
• • • • •
reduction in volumes of freshwater and effluent, increased product quality due to reduced system loading, recovery of raw materials from the effluents (e.g., coating color pigments), enhanced possibilities for the recirculation of biologically treated wastewater, and compliance with statutory limits on effluent concentrations.
In paper machine water circuits, ultrafiltration systems are mainly employed to provide high-quality water for use in high-demand consumers such as high-pressure showers or chemical dilutions. This leads to reduced freshwater consumption and a better machine runnability and paper quality. Ozone treatment. This treatment can be used as an internal treatment technology for decoloration of process waters. The brown color of paper-mill waters is mainly caused by derivatives similar to lignin and humic acid that are characterized by C¼C double bonds. These are the preferred sites of attack by ¨ ller, 2007). ozone, which then destroys them (Bierbaum and O As ozone is one of the strongest oxidants known, it is also possible to eliminate COD by oxidizing water components. ¨ ller and Offermanns (2002) have shown that recirculation of O ozonized waters into papermaking process is possible on the mill scale without any adverse effects on the process or product quality. Numerous other ozone applications – such as COD and AOX reduction, decoloration, microbial and odor control, and improvement of biosludge characteristics – in mill water systems (freshwaters, circuit waters, or effluents) are conceivable as well.
Water in the Pulp and Paper Industry 4.19.4.3 Wastewater Treatment 4.19.4.3.1 Preliminary mechanical treatment: Mechanical processes for removal of solids Effluents from pulp and paper mills contain solids and dissolved matter. The principal methods used to remove solids from pulp and paper mills effluents include screening, settling/clarification, and flotation. The choice of method depends on the characteristics of the solid matter to be removed and the requirements placed on the purity of the treated water. The separation of solids from the effluents is accomplished with the help of screens, grid chambers, and settling tanks. Screens are units which operate according to the sieving/filtration process. The function of the screens is to remove coarse, bulky, and fibrous components from the effluents. If necessary, fractionated particle separation can be achieved by graduating the gap width (bar screen, fine screen, inlet screen, and ultrafine screen). For reasons of operating reliability of wastewater treatment plants, it is also necessary to separate the grit transported with the effluents and other mineral materials from the degradable organic material. Grit separation from effluents can prevent operational troubles such as grit sedimentation, increased wear, and clogging. The grit separating systems currently in use are subdivided into longitudinal grit traps, circular grit traps, and vortex grit traps, depending on their design and process layout. Sedimentation technology is the simplest and most economical method of separating solid substances from the liquid phase. High efficiency is achieved in subsequent effluent treatment processes when the solid substances suspended in the effluents settle in a sedimentation tank as completely as possible, and settled sludge is removed from the sedimentation tank. Sedimentation tanks must be appropriately designed and operated. Alternative sedimentation equipment, with sets of lamella-shaped passages, is employed in the paper industry, especially for effluents with high fiber concentrations. Mechanical effluent treatment alone, however, is not sufficient to keep lakes and rivers clean, since it is incapable of removing colloidal, suspended, and dissolved substances.
4.19.4.3.2 Biological treatment Biological wastewater treatment is designed to degrade pollutants dissolved in effluents by the action of microorganisms. The microorganisms utilize these substances to live and reproduce. Pollutants are used as nutrients. A prerequisite for such degradation activity, however, is that the pollutants are soluble in water and nontoxic. Degradation process can take place either in the presence of oxygen (aerobic treatment) or in the absence of oxygen (anaerobic treatment). Both these naturally occurring principles of effluent treatment give rise to fundamental differences in the technical and economic processes involved (Table 2). The paper industry uses a variety of effluent treatment systems. The preferred process combination for each individual case depends on the grade-specific quality of the effluent that is to be treated. Experience shows that multistage processes based on an aerobic–aerobic or anaerobic–aerobic processing principle enable significantly more reliable
Table 2 treatment
679
Main characteristics of anaerobic and aerobic wastewater
Anaerobic treatment l
COD41000 mg l Low amount of excess sludge Energy generation by use of biogas Low energy demand Low required space Sensitive against high sulfate and calcium concentrations No fully biological degradation
Aerobic treatment High High High Fully
amount of excess sludge energy demand required space biological degradation
operation of the plant. The same effect can be achieved through a cascade system, which allows a graduation of the loading conditions. Among the German pulp and paper mills with onsite wastewater treatment plants, 60% have only aerobic treatment (operated as one- or two-stage processes) for their effluents, whereas 40% have an additional anaerobic stage (Jung et al., 2009). Anaerobic treatment. Anaerobic processes are employed for treatment of more highly polluted effluents such as effluents from recovered paper converting mills (Hamm, 2006). Anaerobic microorganisms conduct their metabolism only in the absence of oxygen. Anaerobic processes are characterized by a small amount of excess sludge produced and low energy requirements. As biogas is produced during the degradation process, anaerobic processes produce an excess of energy. Biogas is a mixture of its principal components, methane and carbon dioxide, with traces of hydrogen sulfide, nitrogen, and oxygen. Biogas is energetically utilized mainly in internal combustion engines or boilers. In its function as a regenerative energy carrier, biogas replaces fossil fuels in the generation of process steam, heat, and electricity. The composition and quality of biogas depend on both effluent properties and process conditions such as temperature, retention time, and volume load. Before discharge into surface waters, anaerobically treated effluents have to undergo aerobic posttreatment, because – according to the current state of the art – fully biological degradation of paper-mill effluents is not feasible (Mo¨bius, 2002). When introducing anaerobic technology into the pulp and paper industry, operational problems and their possible consequences, shown in Table 3, must be taken into account: Among different types of anaerobic reactors, ICs reactors (internal circulation) have achieved a share of more than onethird of the operating reactors and are currently the most frequently used reactors in the German pulp and paper industry. The rest of the market is shared by Biobeds and UASB reactors (UASB, Upflow Anaerobic Sludge Blanket) as well as reactors operating according to the contact sludge principle. Aerobic treatment. Aerobic microorganisms require oxygen to support their metabolic activity. In effluent treatment, oxygen is supplied to the effluent in the form of air by special aeration equipment. Bacteria use dissolved oxygen to convert organic components into carbon dioxide and biomass. In addition, aerobic microorganisms convert ammonified organic nitrogen compounds and oxidize ammonium and
680
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nitrite to form nitrate (nitrification). The key factors for the success of an aerobic process are an adequate amount of nutrients in relation to the amount of biomass, a certain temperature and pH regime, and the absence of toxic substances (Hynninen, 2000). Aerobic processes are characterized by high volumes of excess sludge and higher energy demands compared to anaerobic processes. Furthermore, these reactors typically have large space requirements. Aerobic treatment allows fully biological degradation of paper-mill effluents. The BOD5 efficiency achievable with welloperated activated sludge processes is typically within the range of 90–98% (Hamm, 2006). The drawbacks of aerobic treatment technology include the relatively high operating costs due to the aeration of the effluent. On the other hand, aerobically operated plants exhibit higher plant stability and are less sensitive to fluctuations in effluent and plant parameters. Among different types of aerobic treatment technologies, activated sludge processes are currently the most frequently used treatment technologies in the German pulp and paper industry and have achieved a share of three-quarters of the operating reactors. Both moving-bed bioreactors (MBBRs) and biofilters represent another 10% of the reactors used (Jung et al., 2009). Secondary clarification. Secondary clarification is intended to separate the biomass (activated sludge) formed in biological reactors and is therefore a key element in all processes employed in the final stage of a treatment plant. The quality of the separation process is just as crucial for the final effluent quality as is for the biological treatment itself. As far as activated sludge process is concerned, secondary clarification determines the bioreactor performance. Separation and thickening of the recirculated sludge are crucial for
Table 3 Operational problems and possible consequences on anaerobic treatment in the pulp and paper industry
sludge volumes in biological treatment and for the potential sludge loading as well. Correct dimensioning of secondary clarification is therefore of great importance for overall plant performance.
4.19.4.3.3 Advanced and tertiary treatment Tertiary and/or advanced wastewater treatment is used to remove specific wastewater constituents that cannot be removed by secondary treatment. Different treatment processes are necessary to remove nitrogen, phosphorus, additional suspended solids, refractory organics, or dissolved solids. Sometimes it is referred to as tertiary treatment because advanced treatment usually follows high-rate secondary treatment. However, advanced treatment processes are sometimes combined with primary or secondary treatment (e.g., chemical addition to primary clarifiers or aeration basins to remove phosphorus) or used in place of secondary treatment (e.g., overland flow treatment of primary effluent). The reasons for advanced effluent treatment include
• • •
reduction in costs (discharge fee), compliance with limit values, and increase in production.
Advanced wastewater treatment in the pulp and paper industry is mainly focused on additional biological membrane reactors, membrane filtration techniques such as micro-, ultra-, or nanofiltration, and ozone treatment. Due to the relatively limited full-scale experience, relatively high costs, and greater complexity of water treatment, there have been only few fullscale applications of tertiary treatment of mill effluents up to now. The method that is ultimately chosen depends on the treatment aim and economic efficiency of the method in a given application. Table 4 shows the treatment aims that can be achieved by the different methods.
Operational problem
Possible consequences
4.19.5 Potentials and Limits of Water Saving
High concentrations of suspended solids in the feed flow High sulfate concentrations
Displacement of biomass Loss of pellets Displacement of methane
4.19.5.1 Limiting Effects of System Closure
bacteria
Inhibiting or toxic effects of
When reducing specific effluent volume within the framework of water circuit optimization, typical limits occur that usually require considerable investment to ensure that they will not be
sulfide High calcium concentrations Additives used in production (especially biocides and detergents)
Performance losses Precipitation of CaCO3 Displacement of biomass Inhibiting/toxic influences Poorer degradation
Insufficient supply of nitrogen and phosphorus Temperature variations Fluctuating organics loads (e.g., shock loads)
performance Decomposition/washout of pellets Unstable operation Performance losses Loss of pellets Unstable operation Performance losses Excessive production of organic acids Methanation disturbed
Table 4
Treatment aims of different advanced treatment methods
Treatment method
Aim of treatment
Biofiltration
Reduction in COD and BOD concentration
Ozone treatment Membrane treatment
Filtration processes Denitrification and phosphate precipitation
Removal of suspended solids Elimination of residual COD Decoloration Elimination of residual COD Elimination of suspended solids Demineralization Decoloration Removal of suspended solids Nitrogen and phosphate elimination
Water in the Pulp and Paper Industry
exceeded (Figure 14). A limit in this sense is the freshwater volume that is taken into the system as process freshwater and is used for cooling prior to its final use (2). The second limit is water volume that accumulates together with the rejects and is discharged together with the effluents (3). The third limit is the maximum COD value that the respective product can tolerate in the white water (4). In a selected circuit, this value also corresponds to a minimum effluent volume for the respective system. The above-mentioned limits differ in every individual system. The factors that influence these limits include the existing plant technology, raw materials used, and paper grades produced. A limit encountered in narrowing water circuits that is similar to the cooling water requirements discussed above are the rejects that accumulate when discharged with the paper mill effluent. The proportion of effluents contained in the rejects compared to the total effluent volume may amount to 40–50% of the total effluent volume, especially in paper mills with an integrated deinking plant. If the effluent volume of such a plant is to be reduced drastically, water volume added to the effluents together with the rejects constitutes a lower limit. If a further reduction in the specific effluent volume is intended, then the rejects must be dewatered and part of the filtrate returned to the water circuit. Low specific effluent volumes result in growing system loads in process waters in terms of dissolved and colloidal material (Figure 15) that cause severe quality deterioration (slime spots, odor, color shifts, etc.) and a drop in productivity (machine failures due to scaling and corrosion, slime formation, web breaks, etc.). This situation is aggravated by the use of heavily loaded waste paper.
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If specific effluent volume is to be reduced successfully, the impact of such measures on the papermaking process must also be taken into consideration. Only if we succeed in reconciling the goal of preventing effluent production with the goal of reliable production and satisfactory product quality, can the narrowing and ultimate closure of water circuits come about successfully.
4.19.5.2 Heat Balance Narrowing and closure of water circuits lead to increased temperatures in the stock and water systems of paper mills, taking a constant energy input into account. Nowadays, temperatures of 40–50 1C are achievable in the paper machine loop without additional steam heating (Zippel, 2001). Loop separation and the countercurrent arrangement enable the paper mills to reduce the transfer of detrimental substances coming from highly loaded loops (e.g., stock preparation) into the subsequent process steps, thus relieving paper machine loop. Regarding heat balance of the stock and water system, this is disadvantageous as the highly loaded loops are typically also the hottest (e.g., thermomechanical pulp plant). Heat with quite a high-temperature level is transferred to the effluent. However, at the paper machine, a higher-temperature level would be desirable to improve mechanical dewatering and in turn decrease the energy consumption for thermal drying. Besides the above-mentioned effects, there are other positive and negative impacts of higher process temperatures:
Specific effluent volume l kg−1 Cooling Cooling water water
1
In receiving waters
Fresh water (process water)
2
Waste water
3 4
Cooling water (process water)
Rejects (waste water)
5 COD 5' 5
Rising COD
4
1
To the wastewater treatment plant
} Fresh water
Production
Evaporation
Waste water
Figure 14 Limits in reducing the specific effluent volume. (1) Current situation; (2) cooling water limitation; (3) reject limitation; (4) maximum white water loading; (5) closed water circuit.
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COD concentration in WW1 (g l−1)
30.000 25.000 20.000 15.000 10.000 5.000 0 0
2
4
6
8
10
12
Specific effluent volume (m3 per metric ton of paper) Figure 15 Chemical oxygen demand (COD) concentration in white water 1 of European paper mills producing corrugated base paper as a function of the specific effluent volume.
• • • • •
•
The solubility and activity of most functional chemicals increase with increasing water temperature. The consumption of certain additives such as wet strength agents may increase due to increased temperatures in the water circuit. Slime formation can be restricted by increasing the process temperature above a certain limit. There might be greater formation of anaerobic metabolic products such as hydrogen sulfide. High water temperatures in the papermaking process reduce the energy consumption for pulping and increase the cleaning efficiency of showers. On the other hand, high water temperatures have negative impacts on the energy efficiency of liquid-ring vacuum pumps and the hall climate. Finally, high process temperatures lead to high effluent temperatures. Without any countermeasures, this can cause problems in aerobic effluent treatment plants (poor oxygen solubility) and with the statutory temperature limits.
Integration of waste heat streams is one possibility for paper mills to reduce their energy consumption, but presents them with the conflicting challenges of ensuring both maximum waste heat utilization and safe compliance with statutory limits on effluent temperature. Heat integration measures help optimizing heat balance of paper mills and are a cost-effective way to reduce the specific energy demand of paper mills, thus achieving a productivity increase. Apparently, conflicting objectives, such as increased process and decreased effluent temperatures, may be achieved by appropriately selected measures for heat-balance optimization. Based on available heat sources and sinks and considering other boundary conditions, there are several potential scenarios for heat integration and utilization of waste heat by means of water–water heat exchange or air–water heat exchange. Pinch analysis and process simulation are useful tools for an evaluation of the individual scenarios and an optimization of heat balances. Studies have shown that the
replacement of steam used for process- or freshwater heating yields particularly profitable energy savings (Jung, 2008).
4.19.5.3 Economic Benefits There are many reasons to reduce the specific effluent volume. One important reason is the reduction in water-related costs. In the German pulp and paper industry, the costs of discharging effluents into receiving waters are high and average h0.40 m3 for direct dischargers. Discharging and treating effluents for indirect dischargers, however, are considerably more expensive. The latter involves average costs amounting to h1.12 m3. Reducing effluent volume is very attractive, especially for indirect dischargers. Additional costs arise due to a user fee for freshwater outtake and the operational costs for freshwater treatment (Jung et al., 2009). Despite the above-mentioned problems encountered in narrowing the water circuits, potentials for a reduction of the effluent volume have been discovered in many paper mills studied by PTS in the past few years (Figure 16). Besides water-related costs, another possibility is to reduce energy-related costs by reducing energy consumption due to increased process temperature. As a rule of thumb, every 101 increase in process temperature equals approximately 1% increase in dryness after mechanical dewatering in the wire and press sections. This allows energy consumption in the drying section to be reduced by up to 4%.
4.19.6 Improving Water Efficiency in Paper Manufacturing Industries – 30 Years of Success Water is one of the key components in papermaking. Using more than 1 billion m3 of water per year, the paper industry in Europe had been challenged to reduce the impact on regionally available water resources as one of the most important industrial water consumers. Legislation, stringent discharge
Water in the Pulp and Paper Industry
683
Specific effluent volume (m3 per metric ton of product)
25 Production rate proportional weighted mean 20
15
10
5
8 mills
6 mills
4 mills
4 mills
From recovered paper
Wood containing
Wood free
Specialty paper
0
Figure 16 Optimization potentials of paper-specific effluent volumes.
standards, as well as process and product demands force industry to ensure higher water quality corresponding to increasing costs. For the water-consuming industry, water is no longer regarded as a consumable or utility but as a highly valuable asset. Attention to water scarcity and pollution results in new legislative directives, forcing industries to reduce water use and pollution, and motivating them to implement innovations and carefully observe the impact of measures. The Water Framework Directive (WFD) is one of the main drivers for sustainable water use in Europe, which forced the member states to pay more attention to sustainable and efficient water use. Competent decision making at the top management and well-trained and motivated staff delivered substantial progress in reducing the water consumption in the pulp and paper industry: high competence in closing water circuits, substantially supported by process modeling and automation, and kidney technologies as internal process water treatment, lead to a significant decrease of the average specific effluent volume in the past 30 years. The European collaborative research project, AquaFit4Use, started in 2008, focuses on optimization of existing water circuits and development of new treatment concepts to support the European sustainability policy, such as reducing the use of scarce freshwater, improving the water quality (micropollutants, salts, etc.), and sharing corresponding experiences with other sectors.
References Auhorn W (1984) Das Sto¨rstoff-Problem bei der Verringerung der spezifischen Abwassermenge. Wochenblatt fu¨r Papierfabrikation 2: 37--48. Bierbaum S and O¨ller H-J (2007) Anlagenkonzepte zur Ozonbehandlung von Papierfabriksabwa¨ssern. Allgemeine Papier Rundschau 3: 38--40. Borschke D (2006) Zellstoff- und Papierfabrikation – Prozesswassersysteme im Vergleich. Wochenblatt fu¨r Papierfabrikation 17: 971--981.
Demel I, Dietz W, Bobek B, and Hamm U (2004a) Criteria for the recirculation of biologically treated water to the production. ipw – Das Papier 1: 37--40. Demel I, Dietz W, Bobek B, and Hamm U (2004b) Criteria for the recirculation of biologically treated water to the production (II). ipw – Das Papier 2: 33--35. Federal Statistical Office (2008) Statistical Yearbook 2008. Wiesbaden, Germany. http:// www.destatis.de (accessed March 2010). Hamm U (2006) Environmental aspects. In: Holik H (ed.) Handbook of Paper and Board, pp. 208--218. Weinheim: Wiley-VCH. Herberz J and Bahn W (2006) 10 Jahre Betriebserfahrungen mit einer integrierten biologischen Reinigung im geschlossenen Wasserkreislauf. In: Jung H and Simstich B (eds.) Proceedings Wasserkreisla¨ufe in der Papiererzeugung Verfahrenstechnik und Mikrobiologie, pp. 8/1–8/10. Munich, Germany, 05–06 December. Munich: PTS. Hutter A (2008) Wasserkreisla¨ufe und Wasserqualita¨t in der Papiererzeugung. In: Jung H and Simstich B (eds.) Proceedings Wasserkreisla¨ufe in der Papiererzeugung, pp. 1/1–1/20. Munich, Germany, 02–03 December. Munich: PTS. Hynninen P (ed.) (2000) Papermaking Science and Technology Book 19 Environmental Control. Helsinki, Finland: Fapet Oy. Jung H (2008) Optimisation of the heat balance of papermills. PTS News 1: 30--33. Jung H, Hentschke C, Pongratz J, and Go¨tz B (2009) Wasser- und Abwassersituation in der deutschen Papier- und Zellstoffindustrie – Ergebnisse der Wasserumfrage 2007. Wochenblatt fu¨r Papierfabrikation 6–7: 280–283. Jung H, Pauly D, Beimfohr C, et al. (2007) Odour control – eliminating odour problems in the paper industry. PTS-News 2: 25--29. Kappen J, Hutter A, Bobek B, and Hamm U (2004) Qualitative and quantitative requirements on the water supply of internal consumers. INFOR-Project No. 52R, Munich/Darmstadt. Kappen J and Wilderer PA (2002) Key parameter methodology for increased water recovery in the pulp and paper industry. In: Lens P, Hulshoff Pol L, Wilderer P, and Asano T (eds.) Water Recycling and Resource Recovery in Industries: Analysis, Technologies and Implementation, pp. 229--251. London: IWA Publishing. Mo¨bius CH (2002) Waste Water of the Pulp and Paper Industry, 3 rd edn., Revision December 2008. Augsburg, Germany. http://www.cm-consult.de (accessed March 2010). Negro C and Tijero J (1998) Water in the pulp and paper industry. In: Blanco MA, Negro C, and Tijero J (eds.) Paper Recycling: An Introduction to Problems and their Solutions, pp. 17--46. Luxembourg: European Communities. O¨ller H-J and Offermanns U (2002) Successful start-up of the world’s 1st ozone-based effluent re-circulation system in a paper mill. In: Graham NJD (ed.) Proceedings of the International Conference Advances in Ozone Science and Engineering: Environmental Processes and Technological Applications, pp. 365–372. Hong Kong, People’s Republic of China, 15–16 April. Hong Kong: The Hong Kong Polytechnic University and The International Ozone Association. Pauly D (2001) Kidney-technology opens up new opportunities of integrated white water treatment in recycling mills. In: Gopalaratnam N and Panda A (eds.)
4.20 Water in the Textile Industry J Volmajer Valh, A Majcen Le Marechal, S Vajnhandl, T Jericˇ, and E Sˇimon, University of Maribor, Maribor, Slovenia & 2011 Elsevier B.V. All rights reserved.
4.20.1 4.20.1.1 4.20.1.2 4.20.1.2.1 4.20.1.2.2 4.20.2 4.20.2.1 4.20.2.2 4.20.2.2.1 4.20.2.2.2 4.20.3 4.20.3.1 4.20.3.1.1 4.20.3.1.2 4.20.3.1.3 4.20.3.2 4.20.3.2.1 4.20.3.2.2 4.20.3.2.3 4.20.4 References
Textile Industry Textile and Clothing Industry in Europe Processes in Textile Industry Fibers Finishing processes Characteristic of Textile Water and Wastewater Supply Water Textile Wastewater Textile wastewater from different process steps General characteristics of textile wastewater Treatment and Reuse of Textile Wastewater Wastewater Treatment Technologies Physical methods Chemical processes Biological treatment processes Reuse Pollution-prevention techniques Chemicals and water reuse and recycle: Start-of-pipe approach Process-water reuse and recycle: End-of-pipe approach Conclusions
4.20.1 Textile Industry The textile industry is one of the longest and most complicated industrial chains in the manufacturing industry. It is a fragmented and heterogeneous sector dominated by smalland medium-sized enterprises (SMEs), with a demand mainly driven by three main end uses: clothing, home furnishing, and industrial use. The textile industry is composed of a wide number of subsectors, covering the entire production cycle from the production of raw materials (man-made fibers) to semiprocessed (yarn, and woven and knitted fabrics with their finishing processes), and final products (carpets, home textiles, clothing, and industrial-use textiles) (EURATEX, 2000). The textile industry is a very diverse and heterogeneous industry, with its products being used by virtually everybody – private households and businesses alike. Downstream parts of the textile industry – such as the clothing industry – consume the output of more upstream parts (such as fabrics of all types and colors). The textile industry is also intertwined with the agricultural sector when it needs inputs in the form of natural fibers (such as cotton or wool), and with the chemical industry when it comes to the wide range of man-made fibers (such as nylon or polyester). Hardly any other industrial sector could do without the so-called technical textiles, which include products which are as diverse as filters, optical fibers, packing textiles, ribbons and tapes, air bags, insulation, and roofing materials (Stengg, 2001).
685 685 686 686 686 687 687 689 689 692 695 695 696 697 699 701 702 702 702 703 703
The textile industry is a significant contributor to many national economies, encompassing both small- and large-scale operations worldwide. In terms of its output or production and employment, the textile industry is one of the largest industries in the world. The textile manufacturing process is characterized by high consumption of different resources: water, fuel, and a variety of chemicals in a lengthy process that generates a significant amount of waste. The main environmental problems associated with the textile industry are typically those associated with water pollution caused by the discharge of untreated effluents. Other environmental issues of equal importance are air emission, notably volatile organic compounds (VOCs), excessive noise or odor, as well as workspace safety (UNEP, 1994).
4.20.1.1 Textile and Clothing Industry in Europe The textile and clothing sector is an important part of the European manufacturing industry, giving employment to more than 2 million people. Its importance for social and economic cohesion is increased by the fact that it is dominated by a large number of SMEs, which are often concentrated in particular regions, thus contributing greatly to their wealth and cultural heritage (Stengg, 2001). Being one of the oldest sectors in the history of industrial development, the textile and clothing industry is often referred to as a ‘traditional industry’, as a sector belonging to the socalled ‘old economy’. These notions divert attention from the fact that the European textile and clothing industry has
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Water in the Textile Industry
undergone significant restructuring and modernization efforts during the past 10–15 years, making redundant about onethird of the total work force, increasing productivity throughout the production chain, and reorienting production toward innovative, high-quality products. Like many other sectors, the textile and clothing industry has been greatly affected by the phenomenon of globalization. Europe and the United States are not only important producers of textile and clothing products, but also the most attractive outlets for the so-called exporting countries, many of which are situated in South-East Asia. It should be noted that many developing countries and, indeed, even least developed countries have become very competitive in textiles and clothing, as they combine low-wage costs with high-quality textile equipment and know-how imported from more industrialized countries (Stengg, 2001). The textile and clothing industry is one of the world’s most global industries, and constitutes an important source of income and employment for many European Union (EU) countries. It is important to be aware of how the European textile and clothing industry operates, as well as its many complex structures and processes. The textile industry is a multifaceted area requiring a deep understanding of design, management, and technology. It plays a crucial role in creating innovative and attractive products of multiple uses for various users. It accounts for 5.7% of the production value of world manufacturing output, 8.3% of the value of manufactured goods traded in the world, and over 14% of world employment (Perivoliotis, 2002). Research and innovation have been important tools for the European textile and clothing industry to assert its leading position in global markets. The importance of research and innovation for continued industrial competitiveness is on the increase. The importance of the textile (and clothing) industry in the European economy is shown in Table 1 (EURATEX, 2002). The figures in Table 1 cover only a part of the total number of manufacturing companies in 2000 (i.e., they cover only companies with more than 20 employees). This portion of the industry represents
• • •
• • •
textile finishing, industrial and other textiles (including carpets and wool scouring), and home textiles.
4.20.1.2 Processes in Textile Industry The textile chain begins with the production or harvest of raw fiber. The basic steps in this chain are schematically represented in Figure 1 (US EPA/625/R-96/004, 1996).
4.20.1.2.1 Fibers Two general categories of fibers are used in the textile industry: natural and man-made (comments made by UK to the First Draft of the BREF Textiles, UK, 2001). Man-made fibers encompass both purely synthetic materials of petrochemical origin and regenerative cellulosic materials manufactured from wood fibers. A more detailed classification of fibers is presented in Table 2.
4.20.1.2.2 Finishing processes Pretreatment. Pretreatment processes should ensure (UBA, 1994)
• • •
the removal of foreign materials from the fibers in order to improve their uniformity, hydrophilic characteristics, and affinity for dyestuffs, and finishing treatments; the improvement of the ability to absorb dyes uniformly; and the relaxation of tensions in synthetic fibers.
Pretreatment processes and techniques depend
3.4% of EU manufacturing, 3.8% of the added valued, and 6.9% of industrial employment.
•
The textile chain is composed of a wide range of industrial subsectors, using the entire range of fibers. European industry is still engaged in all production stages, ranging from raw materials (in particular, the production of man-made fibers), to semiprocessed products (in particular, spinning, weaving, knitting, Table 1
and finishing activities), to the final products (e.g., home textiles, carpets, technical textiles, and garments) (Stengg, 2001). The complexity of the sector is also reflected in the difficulty of finding a clear-cut classification system for the different activities involved. As for the scope of this chapter, it is confined to those activities in the textile industry that involve wet processes. This refers primarily to those activities falling within the following new Classification of Economic Activities in the European Community (NACE):
• •
on the kind of fiber to be treated (natural or synthetic fibers), on the form of the fiber (flock, yarn, woven, or knitted fabrics), and on the amount of material to be treated.
Pretreatment operations are often carried out in the same type of equipment used for dyeing (in batch processing, in
Share of the EU-15 textile–clothing industry in the manufacturing industry (companies with 20 employees or more)
2000
Turnover (EUR, billion)
Added value at factor costs (EUR, billion)
Employment (million)
Turnover (%)
Added value (%)
Employment (%)
Textile Clothing
100.5 61.5
31.2 18.2
0.89 0.73
2.1 1.3
2.4 1.4
3.8 3.1
Total textile and clothing
162.0
49.4
1.62
3.4
3.8
6.9
4756.8
1308.0
23.62
100.0
100.0
100.0
Total manufacturing
Water in the Textile Industry
Polymers
Fibers manufacturing
Man-made fibers
Natural fibers
Fibers preparation
Finishing processes Pretreatment Dyeing
Loose fibers /stock
Yarn manufacturing – Spinning
Printing
Yarn
Finishing Fabric production Coating and laminating Carpet back coating
– – – –
Weaving Knitting Tufting Needle felt
particular, the material is most often pretreated in the same machine in which it is subsequently dyed). Dyeing. It is a method for coloring a textile material in which a dye is applied to the substrate in a uniform manner to obtain an even shade with a performance and fastness appropriate to its final use (Bailey et al., 2000; EURATEX, 2000). From a molecular point of view, four different steps are involved: 1. The dye, previously dissolved or dispersed in the dye liquor, diffuses from the liquor to the substrate. 2. The dye accumulates on the surface of the textile material. 3. The dye diffuses/migrates into the interior of the fiber until this is uniformly dyed. 4. The dye must be anchored (fixation) to suitable places within the substrate. Textiles can be colored at any of several stages of the manufacturing process and therefore the following coloring processes are possible:
• • •
Washing Fabric Drying Manufacture of end products
687
• • •
flock or stock dyeing; top dyeing, wherein fibers are shaped in lightly twisted roving before dyeing; tow dyeing, which consists in dyeing the mono-filament material (called tow) produced during the manufacture of synthetic fibers; yarn dyeing; piece (e.g., woven, knitted, and tufted cloths) dyeing; and ready-made goods (finished garments, carpet rugs, bathroom sets, etc.).
Clothing, knitwear, carpet, etc.
Figure 1 Schematic presentation of textile production.
Table 2
Classification of fibers
Natural fibers Animal origin Raw wool Silk fiber Hair Vegetable origin Raw cotton fiber Flax Jute Chemical fibers Natural polymers fibers Viscose, cupro, lyocell Cellulose acetate Triacetate Synthetic polymer fibers Inorganic polymer Glass for fiber glass Metal for metal fiber Organic polymer Polyester Polyamide Polyacrylonitrile Polypropylene Elastane
Dyeing can be carried out in a batch or in continuous/semicontinuous mode. The choice between the two processes depends on the type of makeup, the chosen class of dye, the equipment available, and the cost involved. Both continuous and discontinuous dyeing involve the following steps:
• • • •
preparation of the dye, dyeing, fixation, and washing and drying.
Printing. This is a process for applying color to a substrate. Print color is applied only to defined areas to obtain the desired pattern. This involves different techniques and different machinery with respect to dyeing, but the physical and chemical processes that take place between the dye and the fiber are analogous to dyeing. A typical printing process involves the following steps:
• • • •
Color-paste preparation. When printing textiles, the dye or pigment is not in an aqueous liquor; instead, it is usually finely dispersed in a printing paste, in high concentration. Printing. The dye or pigment paste is applied to the substrate using different techniques: Fixation. Immediately after printing, the fabric is dried and then the prints are fixed mainly with steam or hot air. After-treatment. This final operation consists in washing and drying the fabric (it is not necessary when printing with pigments or with other particular techniques such as transfer printing).
688
Water in the Textile Industry
Finishing (functional finishing). The term finishing covers all those treatments that serve to impart to the textile the desired end-use properties. These can include properties relating to visual effect, handling, and special characteristics such as waterproofing and nonflammability. Finishing may involve mechanical/physical and chemical treatments. Washing. Washing with water is normally carried out in hot water (40–1001C) in the presence of wetting agent and detergent. The detergent emulsifies the mineral oils and disperses the undissolved pigments. Washing always involves a final rinsing step to remove the emulsified impurities. Dry cleaning is sometimes necessary, especially for delicate fabrics. In this case, the impurities are carried away by the solvent, which is usually tetrachloroethylene (perchloroethylene). In the same step, softening treatments may also be carried out. In this case, water and surfactant-based chemicals are added to the solvent. Drying. It is necessary to eliminate or reduce the water content of the fibers, yarns, and fabrics following wet processes. Drying, in particular, by water evaporation, is a highenergy-consuming step.
Synthetic
Cotton
Wool
Fiber preparation
Scouring
Spinning
Carbonizing
W W Texturing
Warping Knitting
W Yarn dyeing W
Knitting Sizing
Heat setting
Weaving
Carbonizing
Singeing W Desizing W
W Scouring / washing
Wool felting
W Bleaching Singeing W
W Dyeing
Mercerizing
W Printing
4.20.2 Characteristic of Textile Water and Wastewater
W Finishing
4.20.2.1 Supply Water Cutting / sewing
The textile industry is very water intensive. Water is used for cleaning the raw material and for many flushing steps during the whole production (Water Treatment Solutions, 2010). In Figure 2, a general flowchart for processes in textile manufacturing is shown, and the processes that need the input water (marked with rounded W) (Bisschops and Spanjers, 2003). Processes using water are desizing, scouring or kiering, bleaching, mercerizing, dyeing, washing, neutralization, and salt bath. Most of them are presented in Tables 3–5. Textile operations vary greatly in water consumption. Wool and felted fabrics processes are more water intensive than other processing subcategories such as wovens, knits, stock, and carpet. Water use can vary widely between similar operations as well (US EPA/625/R-96/004, 1996). The highest water use generally refers to natural fibers. Synthetic fibers require lower water volumes per unit of product, mainly due to the lower cleaning and scouring needs (Matioli et al., 2002). Water consumption varies greatly among unit processes. Certain dyeing processes and print after washing are among the more intensive unit processes. Within the dyeing category, certain unit processes are particularly low in water consumption (e.g., pad batch) (US EPA/625/R-96/004, 1996). An abundant supply of clean water is necessary in order to run a dyeing and finishing plant. Dye houses are usually located in areas where the natural water supply is sufficiently pure and plentiful. Rivers, lakes, and wells represent the major sources of freshwater available for use in wet processing (Tomasino, 1992). Almost all dyes, especially chemicals, and finishing additives are applied to textile substrates from water baths. In addition, most fabric-preparation steps, including desizing, scouring, bleaching, and mercerizing, use aqueous systems.
End product
Figure 2 General flowchart for processes in textile manufacturing.
Table 3 Average water supply for different textile wet processes (Correia et al., 1994) Material
Process
Water usage (l kg1)
Cotton
Desizing Scouring or kiering Bleaching Mercerizing Dyeing
3–9 26–43 3–124 232–308 8–300
Wool
Scouring Dyeing Washing Neutralization Bleaching
46–100 16–22 334–835 104–131 3–22
Nylon
Scouring Dyeing
50–67 17–33
Acrylic
Scouring Dyeing Final scour
50–67 17–33 67–83
Polyester
Scouring Dyeing Final scour
25–42 17–33 17–33
Viscose
Scouring and dyeing Salt bath
17–33 4–13
Acetate
Scouring and dyeing
33–50
Water in the Textile Industry Table 4
Water usage (l kg1) for different materials and processes (Correia et al., 1994)
Material
Process Desizing
Wool Cotton Synthetic Nonspecified
Scouring
Bleaching
Dyeing
Printing
4–77.5 2.5–43 17–67
40–150 38–143 38–143
280–520
30–50
12.5–35
20–300
Table 5 Average, minimum, and maximum water supply for different textile operations (US EPA/625/R-96/004, 1996) Subcategory
689
Table 6
Liquor ratio for various dyeing processes
Process
l kg1
Dyeing winches Hank machines Jet dyeing Package dyeing Pad batch ULLR dyeing
20–30 30 7–10 5–8 5 5
Water usage (l kg1) Minimum Average Maximum
Wool scouring 4.2 Wool finishing 110.9 Low water use processing 0.8 Woven fabric finishing Simple processing 12.5 Complex processing 10.8 Complex processing plus desizing 5.0 Knit fabric finishing Simple processing 8.3 Complex processing 20.0 Hosiery processing 5.6 Carpet finishing 8.3 Stock and yarn finishing 3.3 Nonwoven finishing 2.5 Felted fabric finishing 33.4
11.7 283.6 9.2
77.6 657.2 140.1
78.4 86.7 113.4
275.2 276.9 507.9
135.9 83.4 69.2 46.7 100.1 40.0 212.7
392.8 377.8 289.4 162.6 557.1 82.6 930.7
The amount of water used varies widely in the industry, depending on the specific processes operated at the mill, the equipment used, and the prevailing management philosophy concerning water use. Different types of processing machinery use different amounts of water, particularly in relation to the bath ratio in dyeing processes (the ratio of the mass of water in an exhaust dyebath to the mass of fabric). Washing fabric processes greater quantities of water than dyeing. Water consumption of a batch-processing machine depends on its bath ratio and also on mechanical factors, such as agitation, mixing, bath and fabric turnover rate (called contact), turbulence, and other mechanical considerations, as well as physical flow characteristics involved in washing operations. All these factors affect washing efficiency (US EPA/625/R-96/004, 1996). The influence of the equipment and process selected is presented in Table 6 (EPA Victoria, 1998). From Table 6 we can see that hank machines and dyeing winches are the biggest water consumers (20–30 l kg1). Pad batch and ultralow liquor ratio dyeing processes need only 5 l kg1. The quantity of water used for a particular process also depends on equipment modernization and development. As an example, batch dyeing machines for knitwear have gone from 30 l kg1 to only 6 l kg1 of treated material over the last four decades (Wenzel and Knudsen, 2005). In general, heating of dyebaths constitutes the major portion of energy consumed in dyeing. Therefore,
low-bath-ratio dyeing equipment not only conserves water but also saves energy, in addition to reducing steam use and air pollution from boilers. Low-bath-ratio dyeing machines conserve chemicals as well as water and also achieve higher fixation efficiency. However, the washing efficiency of some types of low-bath-ratio dyeing machines, such as jigs, is inherently poor; therefore, a correlation between bath ratio and total water use is not always exact (US EPA/625/R-96/004, 1996). Water quality for all processes should be of such quality as to avoid any process and final-product-quality problems. Mostly, fresh softened water is used for all processes, although sometimes water of lower quality can be used as well. Three types of water quality are suggested for use in textile industry (Lockerbie and Skelly, 2003; Vandevivere et al., 1998): 1. High-quality water. It can be used for all processes, such as dyebaths, print pastes, finishing baths, and final rinse bath (Table 7). Consumption of such water is 10–20% of the total water consumption. Four different sources are presented: fresh softened water, recycled effluent (proposed), mains drinking-water prescribed concentrations or values (PCVs), and Confederation of British Wool Textiles (CBWT) water specification. 2. Moderate-quality water. It is used for washing-off stages after scouring, bleaching, dyeing/printing, and finishing (Table 8). About 50–70% of total water consumption consists of such water needs. Final rinse bath in the washing processes should be always high-quality water to ensure that the material is free from traces of contamination. 3. Low-quality water. It can be used for washing-down equipment, screen washing in print works, and general washdown of print paste containers and floors (Table 9). Quantity presents only 10–20% of total water consumption, but it is wasteful to use high-quality water for such operations.
690
Water in the Textile Industry
Table 7
Water quality suitable for all processes
Colora (mg l1 Pt scale) COD (mg l1 O2) pH Total hardness (mg l1) Chloride (mg l1) Sulfate (mg l1 SO4) Fe (mg l1) Cu (mg l1) Cr (mg l1) Al (mg l1) Mn (mg l1) Zn (mg l1)
Fresh softened water
Recycled effluent
Main water PVCs
CBWT specification
None visible
None visible 20–50 6.5–7.5 90b 500
20 5.5–9.5
None visible 6.0–8.0
250 (Ca), 50 (Mg) 400 250 0.2 3 0.05 0.2 0.050 5
60–80b
6.5–7.5 50b 300 0.05 0.05
0.1 0.005 0.01 0.02
0.1 0.1
0.05 0.1
a
Suggested specification for water with no visible color absorbance in 10 mm cell: 450 nm, 0.020.04; 500 nm 0.020.05; 550 nm, 0.010.03; 600 nm, 0.010.02. Measured as ppm CaC03. COD, chemical oxygen demand; PVC, polyvinylchloride; CBWT, Confederation of British Wool Textiles.
b
Table 8 Suggested water quality suitable for washing-off processesa
Table 9 only
Parameter
Parameter
b
Color COD (mg l1) pH Total hardness (ppm CaCO3) Chloride (mg l1) Fe (mg l1) Cu (mg l1) Cr (mg l1)
Maximum recommended level None visible 200 7.0–8.0 100 500–2000 0.1 0.05 0.1
a
Final rinse bath to use high-quality water. Suggested specification for water with no visible color absorbance in 10 mm cell: 450 nm, 0.020.04; 500 nm, 0.020.05; 550 nm, 0.010.03; 600 nm, 0.010.02.
a
Suggested water quality suitable for equipment washdown
Color COD (mg l1) pH Total hardness (ppm CaCO3) Chloride (mg l1) Fe (mg l1) Cu (mg l1) Cr (mg l1)
Maximum recommended level None visible 500–2000 6.5–8.0 100 3000–4000 0.1 0.05 0.1
a
b
Suggested specification for water with no visible color absorbance in 10 mm cell: 450 nm, 0.020.04; 500 nm, 0.020.05; 550 nm, 0.010.03; 600 nm, 0.010.02.
4.20.2.2 Textile Wastewater
In Europe, 108 million tons of wastewater is produced on a yearly basis and 36 million tons of chemicals and auxiliaries have to be removed from the textile wastewater. Textile wastewater typically contains a complex mixture of organic and inorganic chemicals, due to the wide variety of the process steps.
The textile industry is one of the most polluting industries. Many different processes are used and almost all of them generate wastewater. Wastewater from textile sector is composed of cleaning water, process water, noncontact cooling water, and storm water. The amount and the composition of wastewater vary and depend on different factors, including the nature of the processed fabric, applied dye, or special finishing; the type of the process; the equipment used; and the prevailing management philosophy regarding water use. Changes in machines, used chemicals, or any characteristic of the processes also change the nature of the generated wastewater. Scouring, dyeing, printing, finishing, and washing generate the majority of the textile wastewater. Large-volume wastes include wash water from preparation and continuous dyeing, alkaline wastewater from preparation, and batch dye wastewater containing large amounts of dye, salts, acids, or alkalis, and also other toxic additives in smaller amounts. Primary sources of biological oxygen demand (BOD) include waste chemicals or batch dumps, starch-sizing agents, knitting oils, and degradable surfactants.
4.20.2.2.1 Textile wastewater from different process steps The following processes in the textile industry produce wastewater containing different pollutants: 1. Desizing. It is the process for removing the size chemicals from the textile. Wastewater from the desizing process varies according to the used sizes and recipes and contains pollutants such as different additives, surfactants, enzymes, acids or alkalis, as well as the size themselves. The generated wastewater can be the largest contributor to the overall BOD and the total suspended solids (TSSs). When the natural sizes, based on starch or proteins, are used for sizing, the wastewater after desizing is characterized by high BOD and BOD/COD ratio. If sizing is carried out using synthetic materials, such as polyvinyl alcohol or carboxymethyl cellulose, the BOD reduction can be up to
Water in the Textile Industry
90%. Possible pollutants in wastewater after desizing process are shown in Table 10 (Correia et al., 1994). 2. Scouring. It is the process for removing different impurities from both natural and synthetic materials. The intensity of the scouring process depends on the type of material. Oils, fats, waxes, minerals, and plant matter can be present in natural fibers, whereas synthetic fibers can contain spin finishing and knitting oils. These impurities can be removed either with water or with organic solvents. Water scouring is usually preferred over solvent scouring, because water is nonflammable, nontoxic, plentiful, and cheaper. For cotton scouring, hot alkaline solutions, containing detergents or soaps, are used. Sourcing effluents can also contain herbicides, insecticides, defoliants, and desiccants, which are used in the growing of cotton, as well as fungicides such as pentachlorophenols used to prevent mildew during storage and transportation of cotton. Raw-wool scouring is the most polluting process in the textile industry. The pollution load results from impurities present in raw wool such as wax, suint, urine, feces, vegetable
691
matter, mineral dirt, and, on the other hand, the soap detergent and alkali used during the scouring and washing processes. Wool grease is the major problem in treating wool, because of its nonbiodegradability. It is a mixture of cholesterol esters, long-chain fatty acids, free fatty acid, free alcohol, and hydrocarbons. Synthetic souring requires less scouring than cotton or wool. Inorganic and organic substances which could be present in wastewater after scouring, for different fibers, are shown in Table 11 (Correia et al., 1994). 3. Bleaching. It is commonly used to remove natural coloring of cotton and other fibers. In this step, the most common agents are hydrogen peroxide, sodium hypochlorite, sodium chlorite, and sulfur dioxide gas. Auxiliary chemicals such as sulfuric acid, hydrochloric acid, sodium hydroxide (caustic soda), sodium hydrogen sulfite (sodium bisulfite), surfactants, and chelating agents are also used and released into the wastewater. Bleaching wastewater usually has high solid content with low-to-moderate BOD levels. Inorganic and organic substances which could be present in wastewater after bleaching, for different fibers, are shown in Table 12 (Correia et al., 1994). 4. Mercerizing. It improves strength, luster, and dye affinity of cotton fabrics. Cotton fabrics are treated with solutions of sodium hydroxide (caustic soda) followed by neutralization and several rinses. Wastewater generated by mercerizing has low BOD and total solid levels but high pH (Table 13) (Correia et al., 1994). 5. Dyeing. The dyeing operations of textiles may take part in the process chain at different stages of production (fibers, yarn, or piece dyeing). Stock dyeing is used to dye fibers. Top dyeing is used to dye combed wool silver. Yarn dyeing and piece dyeing are used after the yarn has been constructed into the fabric.
Table 10
Possible pollutants in desizing effluents
Fibers
Inorganic substances
Organic substances
Cotton Linen Viscose
Naþ, Ca2þ, NH4 þ , SO4 2 , CI
Silk Acetates Synthetics
Naþ, NHþ 4, CO3 2 , PO4 3
Carboxymethyl cellulose, enzymes, fats, hemicellulosses, modified starches, nonionic surfactants, oils, starch, waxes Carboxymethyl cellulose, enzymes, fats, gelatine, oils, polymeric sizes, polyvinyl alcohol, starch, waxes
Table 11
Possible pollutants and characteristics of effluents from scouring
Fibers
pH
BOD (mg l1)
TSS (mg l1)
Inorganic substances
Organic substances
Cotton
10–13
50–2900
7600–17 400
Naþ, CO3 2 , PO4 3
Anionic surfactants, cotton waxes, fats, glycerol, hemicelluloses, nonionic surfactants, peptic matter, sizes, soaps, starch
Viscose
8.5
2832
3334
Na þ , CO3 2 , PO4 3
Acetates
9.3
2000
1778
Anionic detergents, fats, nonionic detergents, oils, sizes, soaps, waxes
Naþ, CO3 2 , PO4 3
Anionic surfactants, antistatic agents, fats, nonionic surfactants, oils, petroleum spirit, sizes, soaps, waxes
Naþ, NH4 þ , CO3 2 , PO4 3
Anionic detergents, glycol, mineral oils, nonionic detergents, soaps
Naþ, NH4 þ , Kþ, Ca2 þ , CO3 2 , PO4 3
Acetate, anionic surfactants, formate, nitrogenous matter, soaps, suint, wool grease, wool wax
Synthetics
Wool (yarn and fabric)
Wool (loose fiber)
9–14
3000–40 000
1129–64 448
692
Water in the Textile Industry
Textiles are dyed using a wide range of dyestuffs, techniques, and equipment. Each dyeing process requires different amounts of dye per unit of fabric to be dyed. In the textile industry, synthetic dyes, derived from coal tar and petroleum-based intermediates, are used. Dyes can be present as powders, granules, pastes, and liquid dispersions, with concentrations of active ingredients ranging typically from 20% to 80%. Dyeing can be performed by using continuous or batch processes. Auxiliary chemicals and controlled dye-bath conditions accelerate and optimize the migration of the dye molecules from the solutions to the fiber. The dye is fixed on the fiber thermally and/or chemically. Table 12
Possible pollutants in bleaching effluents
Fibers
Inorganic substances
Organic substances
Cotton Linen Viscose Jute
Naþ, NH4 þ CIO, CI, O2 2 , F, SiO3 2
Formate
Synthetics Acetates
SiO3 2 , PO4 3 , F
Wool
Naþ, O2 2
Table 13
The water consumption in dyeing processes is very high (up to 300 l kg1). Water is used not only in the dyeing process itself, but also for rinsing operations of the dyed material. Dyes and different auxiliaries such as organic acid, fixing agents, defoamers, oxidizing/reducing agents, and diluents are typical pollutants generated in the dyeing step. Quite a large amount of the unfixed dye leaves the dyeing unit. Metals and almost all of the salts and dyes present in the overall textile wastewater originate from dyeing operations. The possible pollutants and characteristics of effluents from dyeing processes for different fibers are listed in Table 14 (Correia et al., 1994). 6. Printing. For fabric printing, many different colorants and patterns, including a variety of techniques and machines, are used. The most common printing techniques used are rotary screen, and other methods such as direct, discharge, resist, flat screen, and roller printing often used commercially. Pigments are used for about 75–85% of all printing operations. Pigments do not require washing steps and generate little waste. Compared to the dyes, pigments are typically insoluble and have high affinity for the fibers. An important component in textile printing is the print paste, which consists of water, thickeners, dyes, urea, and various other chemicals such as surfactants and organic solvents. The printing method determines the wastewater characteristics. Printing wastewaters are small in volume
Oxalate
Possible pollutants and characteristics of effluents from mercerizing
Fibers
pH
BOD (mg l1)
TSS (mg l1)
Inorganic substances
Organic substances
Cotton Linen
5.5–9.5
45–65
600–1900
Naþ, NH4 þ , CO3 2 , SO4 2
Alcohol sulfates, anionic surfactants, cyclohexanol
Table 14 Fibers Cotton Linen
Possible pollutants and characteristics of effluents from dyeing pH
BOD (mg l1)
TSS (mg l1)
Polyester
3þ
2þ
Organic substances
11–1800
500–14 100
Na , Cr , Cu , Sb3þ, Kþ, NH4 þ , CI CO3 2 , CO4 2 , F, NO2 , O2 2 , S2, S2 O3 2 , SO3 2 , SO4 2
Naphtol, acetate, amides of naphtoic acid, anionic dispersing agents, anionic surfactants, cationic fixing agents, chloro amines, formaldehyde, formate, nitro amines, nonionic surfactants, residual dyes, soaps, soluble oils, sulfated oils, tannic acid, tartrate, urea
4.8–8
380–2200
3855–8315
Naþ, Cr3þ, Cu2þ, Sb3þ, Kþ, NH4 þ , Al3þ CI, CO3 2 , S2 O4 , SO3 2 , SO4 2 Naþ CI, CO3 2
Acetate, dispersing agents, formate, lactate, residual dyes, sulfated oils, tartrate
Naþ, NH4 þ , Cu2þ, SO4 2
Acetate, aromatic amines, formate, leveling agents, phenolic compounds, residual dyes, retardants, surfactants, thiourea dioxide
Naþ, NH4 þ , Cl, S4 O6 2 , CIO, SO3 2 , NO3
Acetate, anionic surfactants, antistatic agents, dispersing agents, dye carriers, EDTA, ethylene oxide condensates, formate, mineral oils, nonionic surfactants, residual dyes, soaps, solvents
Polyamide Acrylic
þ
5–10
Viscose
Wool
Inorganic substances
1.5–3.7
175–2000
480–27 000
833–1968
Acetate, formate, polyamide oligeines, residual dyes, sulfonated oils
Water in the Textile Industry
and contain urea, dyes or pigments, organic solvents, and metals. The concentration of the pollutants in printing wastewater is higher than that in dyeing wastewater. 7. Finishing. This can refer to the chemical or mechanical treatments performed on fiber, yarn, or fabric to improve appearance, texture, or performance. Mechanical finishes can involve brushing, ironing, or other physical treatments used to increase luster and feel of textiles, such as heat setting, napping, softening, optical finishing, shearing, and compacting. The application of chemical finishes to textile can impart a variety of properties ranging from decreasing static cling to increasing flame resistance. Chemical treatments are optical finishes, adsorbent and soil-release finishes, softeners and abrasion-resistant finishes, and physical stabilization and crease-resistant finishes. Wastewaters from the finishing units are extremely variable in composition and can contain resins, waxes, softeners, acetate, stearate, as well as toxic organic compounds (pentachlorophenols and ethylchlorophosphates).
4.20.2.2.2 General characteristics of textile wastewater Textile wastewater is characterized mainly by measuring BOD, chemical oxygen demand (COD), suspended solids, and dissolved solids. Typical characteristics of textile industry wastewater are presented in Table 15. Wastewaters from the textile industry are usually polluted with recalcitrant or hazardous organics, such as dyes, surfactants, metals, salts, and persistent organic pollutants (POPs) as well. They are discussed in the following: 1. Dyes. Most of the wastewater produced during the textile material processing is colored. The main sources of color in the textile effluents arise from dyes and pigments in the dyeing and printing operations. It is known that the presence of very small quantities of dyes in water (less than 1 ppm) is highly visible due to their brilliance. There are more than 10 000 commercially available dyes with a production of over 7 105 tons yr1 (Zollinger, 1987). The exact data on the quantity of dyes discharged into the environment are also not available. It is assumed that 2% of the dyes produced is discharged directly in aqueous effluent, and B10% is subsequently lost during the textile coloration process (Easton, 1995). Dyes cause a lot of problems in the environment. They can remain in the environment for an extended period of time, because of high thermal and photo stability (the half-life of hydrolyzed Reactive Blue 19 is about 46 years at pH 7 and 25 1C Table 15
Characteristics of textile wastewater
Parameters
Values
pH BOD (mg l1) COD (mg l1) TSS (mg l1) TDS Chloride (mg l1) Total Kjeldahl nitrogen (mg l1) Color (Pt–Co)
1.9–13 50–40 000 150–12 000 15–64 000 2900–3100 1000–1600 70–80 50–2500
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(Hao et al., 2000)). Depending on dye concentration and exposure time, dye can have acute and/or chronic effects on exposed organism. The greatest environmental concern with dissolved dyes is their absorption and reflection of sunlight entering the water (rivers, lakes, etc.). Light absorption diminishes photosynthetic activity of algae and seriously influences the food chain. Many dyes and their breakdown products are carcinogenic, mutagenic, and/or toxic to life. Mathur et al. (2005) studied the influence of textile dyes (known only by their trade name) on the health of textile-dyeing workers and the environment. The dyes were used in their crude form (without previous purification), because they wanted to test the potential danger that dyes represent in actual use. The results clearly indicated that most of the used dyes are highly mutagenic. Brown and DeVito (1993) studied how it is possible to predict the toxicity of new azo dyes. The systematic backtracking of the flows of wastewater from textile-finishing companies led to the identification of textile dyes as a cause for strongly mutagenic effects. Several textile dyes used in the textile-finishing companies in the European Union were examined for mutagenicity. According to the obtained results, the dyes which were considered to present a potential toxicity have been withdrawn from the market and have been replaced with less harmful and biodegradable substances (Ja¨ger et al., 2004; Schneider et al., 2004). Degradation of dye Direct Blue 14 led to the carcinogenic aromatic amine o-tolidine (Platzek et al., 1999). Dyes can cause allergies such as contact dermatitis (Pratt and Taraska, 2000) and respiratory diseases(Estlander, 1988; Wilkinson and McGechaen, 1996; Zuskin et al., 1998), allergic reaction in eyes, skin irritation, and irritation to mucous membrane and the upper respiratory tract. As it is known, reactive dyes form covalent bonds with cellulose, woolen, and polyacrylate fibers. It is assumed that in the same manner, reactive dyes can bond with –NH2 and –SH groups of proteins in living organisms. Many investigations have been made on respiratory diseases in workers dealing with reactive dyes. Certain reactive dyes have caused respiratory sensitization of workers occupationally exposed to them (Majcen Le Marechal et al., 1996). Organic dyes contain substituted aromatic and heteroaromatic groups. The color of dyes results from conjugated chains or rings that can absorb different regions of wavelength. The chromophores of organic dyes are usually composed of double carbon–carbon bonds, double nitrogen– nitrogen bonds, double carbon–nitrogen bonds, and aromatic and heterocyclic rings containing oxygen, nitrogen, or sulfur. Azo dyes, which contain one or more azo bonds, are the most widely used synthetic dyes and are present in 60–70% of all textile dyestuffs produced (Carliell et al., 1995). Azo dyes can be used on natural fibers (cotton, silk, and wool) and synthetic fibers (polyesters, polyacrylic, rayon, etc.). Azo dyes are mostly used for yellow, orange, and red colors. Biodegradation of more than 100 azo dyes have been tested and it was found that only a very few were degraded aerobically. The degree of stability of azo dyes under aerobic conditions depends on structure of the molecule. Dye C.I. Acid Orange 7 is one of the rare dyes which is aerobically biodegradable. Under anoxic conditions, azo dyes
694
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are cleaved to aromatic amines, which are not further metabolized under anaerobic conditions but are readily biodegraded in an anaerobic environment (Figure 3) (Vandevivere et al., 1998). Anthraquinone dyes constitute the second most important class of textile dyes. They have a wide range of colors in almost the whole visible spectrum, but they are most commonly used for violet, blue, and green colors. With regard to method and domain of usage, dyes are classified into acid, reactive, direct, basic, disperse, metal complex, vat, mordant, and sulfur dyes. Most commonly in use today are reactive and direct dyes for cotton and viscose-rayon dyeing and disperse dyes for polyester dyeing. Reactive dyes are termed chemically as colored compounds with a functional group capable of forming a covalent bond with a suitable substrate. Reactive dyes represent 20– 30% of the total dyes in the market. Reactive dyes are characterized by low fixation rate, and around 30% of the applied reactive dyes are wasted because of dye hydrolysis in the alkaline conditions of the dyebath. As a result, dyehouse effluents typically contain 0.6–0.8 g dye dm3 (Stenken-Richter and Kermer, 1992). Generally, dyes can be classified with regard to (1) their chemical structure, (2) the method and domain of usage, and (3) chromogen (Table 16).
HO
N HO3S
Ar — NH2
N
Figure 3 Chemical structure and degradation under anoxic condition of the azo dye with C.I. Acid Orange 7.
Table 16
In the textile-dyeing process, dyes are always used in combination with other chemicals (acids, alkali, salts, fixing agents, carriers, dispersing agents, and surfactants) which are partly or almost completely discharged in the wastewater together with the numerous additives and impurities present in the commercial dye products. Public perception of water quality is greatly influenced by the color. Therefore, the removal of color from wastewater is often more important than the removal of the soluble colorless organic substances. 2. Metals. Many textile mills have metals in their effluent, but their concentration decreased in the last decade, mainly because of the reduction of the metal contents in the dye. Metals include copper, cadmium, chromium, nickel, zinc, and lead. Metals enter the textile effluents in many ways: incoming supply water, metal parts (such as pumps, pipes, and valves), oxidizing and reducing agents, electrolyte, acid and alkali, dyes and pigments, certain finishes, herbicides, and pesticides. However, the main source of heavy metals is the dyeing process. Dyes may contain metals such as zinc, cobalt, and chromium. In some dyes, metals can form an integral part of the dye molecule; metals are functional, but in most dyes metals are just impurities generated during the dye manufacture. Mercury or other metals may be used as catalysts in the synthesis of dyes and may be present as by-products. Concentrations of metals in the dyeing effluents can be in the range 1–10 mg l1. For example, after dyeing of wool with basic dyes, the concentration of cadmium in wastewater is 7.5 mg l1. The concentration of chromium in dyeing effluents after dyeing cotton with direct dyes is 12.05 mg l1. Dyeing viscose with direct dyes revealed measurements of 2.7 mg of chromium l1, 8.52 mg of copper l11, and 1.95 mg of lead l1 in the wastewater. (EURATEX, 2000).
Classification of the dyes Classification
With regard to chemical structure (C.I.) With regard to method and domain of usage (C.I.) With regard to chromogen n-p* With regard to the nature of donor– acceptor couple With regard to the nature of polyenes Acyclic and cyclic
Cyanine
Subclass
Characteristic
Azo, anthraquinone, triphenylmethane, indigo, etc. Direct, acid, basic, reactive, reductive, sulfuric, chromic, metal-complex, disperse, pigment, etc. Absorptive, fluorescent and dyes with energy transfer, etc. 1-Aminoanthraquinone, p-nitroaniline, etc.
The classification of a dye by chemical structure into a specific group is determined by the chromophore Dyes used in the same technological process of dyeing and with similar fastness are classified into the same group This classification is based on the type of excitation of electrons, which takes place during light adsorption These chromogens contain a donor of electrons (unbound electron couple), which directly bonds to the system of conjugated p electrons
Polyolefins, annulenes, carotenoids, rhodopsin, etc.
Polyene chromogen contains sp2 (or sp) hybridized atoms. The molecules enclose single and double bonds that form open chains, circles, or a combination of both Cyanine chromogens have a system of conjugated p electrons, in which the number of electrons matches the number of p-orbitals
Cyanines, amino-substituted di- and tri-arylmethane, oxonols, hydroxyarylmethanes, etc.
Water in the Textile Industry
3. Salts. The presence of salts in textile wastewater has been identified as a potential problem by several authors. Salts in textile processes are used as raw materials or produced as by-products of neutralization, or in other reactions. Salt is used mostly to assist the exhaustion of ionic dyes, particularly anionic dyes, such as direct and reactive dyes on cotton. Typical cotton batch-dyeing operations use salts in the range 20–80% weight of dyed material. The concentration of salts in such wastewater is 2000–3000 ppm (Matioli et al., 2002). Sodium chloride (common salt) and sodium sulfate (Glaubers salt) constitute the majority of total salt use. Other salts used as raw materials or formed during the textile operations include magnesium chloride (Epson salt) and potassium chloride, and others in low concentrations. 4. Persistent organics or hazardous organics. The persistent molecules present in textile wastewater belong to very diverse chemical classes, each used in relatively small amounts. The persistent organics include surfactants or their byproducts, dyeing auxiliaries such as polyacrylates, phosphates, sequestering agents (ethylenediaminetetraacetic acid (EDTA)), deflocculating agents (lignin or naphtahalenesulfonates), antistatic agents for synthetic fibers, carriers in disperse dyeing of polyester, fixing agents in direct dyeing of cotton, preservatives (substituted phenol), and a large number of finishing auxiliaries used for fireproofing, mothproofing, and water proofing. The most toxic among POPs are the commonly named dioxins and dioxin-like compounds. Dioxin (Figure 4) is the term for a group of chemical compounds with 75 polychlorinated dibenzo-p-dioxins (PCDDs) and 135 polychlorinated dibenzofurans (PCDFs). The textile industry is a potential source of PCDD/Fs. They can arise from the various processes involved in the industry (Krizˇanec and Majcen Le Marechal, 2006): Pesticide pentachlorophenol (PCP) is used as a biocide for cotton and other materials. Pesticides, such as pentachlorophenol, are known to be contaminated with PCDD/Fs. Dyestuffs are contaminated by PCDD/Fs. Textile processes may utilize chlorinated chemicals contaminated by PCDD/Fs. Washing processes in alkaline media are part of the textile finishing processes. Large volumes of effluent water are released into the environment. The main source of dioxins in the textile industry are dioxazine and antraquinone dyes and pigments, produced 9 Clx
O
1
8 7 6
O
9 2
ClY
3 4
Clx
1
8 7 6
O
2
ClY
3 4
X + Y = 1– 8 (75 congeners)
X + Y = 1– 8 (135 congeners)
Polychlorinated dibenzo-p-dioxins
Polychlorinated dibenzofurans
Figure 4 Molecular structure of the polychlorinated dibenzo-p-dioxins and dibenzofurans.
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from chloranil as intermediate product, and chloranil itself used as a catalyst in the production of dyes and pigments. Various dyes and pigments were analyzed for the presence of PCDD/Fs. Considerable levels of PCDD/Fs were determined in some dioxazine dyes and pigments, phatalocyanine dyes, and in printing inks. Concentrations of PCDD/Fs in Direct Blue 106 dye, Direct Blue 108 dye, and Violet 23 pigment were in the mg kg1 range with octachlorodibenzodioxin (OCDD) and octachlorodibenzofuran (OCDF) as dominant homologs. The concentration of OCDD in Direct Blue 106 was 41.9 mg kg1 and the concentration of OCDF was 12.4 mg kg1 (Williams et al., 1992). Hutzinger and Filder (US Environmental Protection Agency, 2000) found mg kg1 range levels of PCDD/Fs for higher chlorinated congress in sample of Ni-phthalocyanine dye. Results of the analyses of PCDD/Fs were reported for four printing inks obtained from a supplier in Germany. In the two inks used for rotogravure printing and two used for offset printing, the content of PCDD/Fs ranged from 17.7 to 87.2 ng TEQ kg1 (TEQ, toxicity equivalent; Santl et al., 1994). A high concentration of mixed polychlorinated and polybrominated dibenzo-p-dioxins and polychlorinated and polybrominated dibenzofurans (PBCDD/Fs) was detected after flame-retardant finishing-textile processes. A flame-retardant finish on upholstery material on the basis of PVC, Sb2O3, and hexabromocyclododecane results in the final product concentrations up to 19 mg kg1 of PBCDD/Fs. PCP and other chlorophenols can be the source of PCDD/Fs in wastewaters. A generation of dioxins was reported from the direct photolysis of pentachlorophenolcontaining water. Waddell et al. (1995) investigated the formation of dioxins by the ultraviolet (UV) photolysis of pentachlorophenol with or without addition of H2O2. Their study showed high levels of PCDD, especially OCDD. The presence of halogenated organic compounds (adsorbable organic halides (AOX)) in textile wastewater may derive from hypochlorite bleaching operations or from spent liquors following shrink-proofing finishing treatment by chlorine. The effluents after bleaching with hypochlorite may contain up to 100 mg dm3 AOX including considerable amounts of chloroform. Some reactive dyes also contain AOX. In the effluent from textiledyeing operation, an average of 0.75 mg dm3 was measured (Grutner et al., 1994). 5. Toxicity of wastewater. The toxicity of textile wastewater varies considerably among different processes in textile industry. Wastewater of some processes have high aquatic toxicity, while others show little or no toxicity. It is impossible to identify all toxic compounds used in textile production, because of the huge variety of chemicals used and the lack of data about their toxicities. Textile wastewater can contain thousands of different compounds, and identifying and testing all of them are practically impossible and too expensive. In general, the overall toxicity is determined by the toxicity test of the whole effluent stream on aquatic organisms, which is a cost-effective method. Table 17 summarizes the results for about
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Table 17 textile mills
Results from aquatic toxicity testing of effluent from 75
Toxicity (%)
Number of tests
o9 10–19 20–29 30–39 40–49 50–59 60–69 70–79 80–89 90–100 4100 (no toxicity)
7 6 8 2 4 9 3 8 2 3 38
Table 18 Agent Salt Surfactants Metals Organics Biocides Toxic anions
Typical causes of aquatic toxicity Chemical example NaCI, Na2SO4 Ethoxylated phenols Copper, zinc Chlorinated solvents Pentachlorophenol Sulfide
Source Dyeing Multiple processes Dyes Scour, machine cleaning Wool fibers contaminant Sulfur dyeing
75 companies (Horning, 1977); toxicity in the table is LC50 in percent and the higher number represents the lower toxicity. The source of aquatic toxicity can be dyes, salt, surfactants, ionic metals, toxic organic chemicals, biocides, and toxic anions. Examples of compounds in each of these classes and their source are shown in Table 18 (EPA/625/R96/004, 1996).
4.20.3 Treatment and Reuse of Textile Wastewater 4.20.3.1 Wastewater Treatment Technologies Textile wastewater may be treated by physical, chemical, or biological methods. For decoloration and degradation of textile wastewaters, many treatment technologies have been developed, but every existing technology presents limitations – advantages and disadvantages. Textile wastewater is very complex, so the use of a universal wastewater treatment seems to be impossible. The wastewater-treatment technologies used will depend on the wastewater characteristic (type, dye concentration and auxiliaries, and pH). It is apparent that a single wastewater-treatment system is unable to overcome all problems by itself to provide an efficient treatment of effluents and be cost effective at the same time. In this section, an overview of treatment technologies used in textile effluents is presented. Dyes containing wastewater can be treated by chemical or physical methods of dye removal, which refer to the process called decoloration, and by means of biodegradation, which tells us more about the fate of dyes in the environment.
Physical methods include different precipitation methods (coagulation, flocculation, and sedimentation), adsorption (on a wide variety of inorganic and organic supports), filtration, reverse osmosis, ultrafiltration, and nanofiltration. Biological treatments differ according to the presence or absence of oxygen and are termed aerobic and anaerobic treatment, respectively. Since biological treatments simulate degradation processes that occur in the environment, they are also called biodegradation. Chemical treatment methods are those in which the removal or conversion of dyes and other contaminants is brought about by the addition of chemicals or by chemical reactions (reduction, oxidation, compleximetric methods, ion exchange, and neutralization). The treatment of colored wastewaters is therefore restricted not only to the reduction of ecological parameters (COD, BOD, total organic carbon (TOC), AOX, temperature, and pH), but also to reduction of dye concentrations in wastewaters.
4.20.3.1.1 Physical methods The physical methods of treating textile wastewater are as follows: Adsorption. It is the process of collecting soluble substances that are in solution on a suitable interface. Adsorption methods for decoloration are based on the high affinity of many dyes for adsorbent materials. Some physical and chemical factors have an influence on dye removal by adsorption. These factors are dye-adsorbent interactions, adsorbent surface area, particle size, temperature, pH, and contact time. The main criteria for selection of an adsorbent should be based on characteristics such as high affinity and capacity for target compounds and the possibility of adsorbent regeneration (Santos et al., 2007). Adsorption on sludge is the main abiotic mechanism of removing dyes from wastewater. The most important factors influencing the adsorption test are sludge quality, water hardness, duration of the test, and test-substance concentration. Pagga and Taeger (1994) have described the static and dynamic removal studies involving water-soluble dyes (acid and reactive) and poorly soluble dyes (disperse). Activated carbon is the most commonly used method of dye removal by adsorption. It is very effective in adsorbing cationic, mordant, and acid dyes, and to a slightly lesser extent, disperse, direct, vat, pigment, and reactive dyes (Nassar and El-Geundi, 1991; Raghavacharya, 1997). Its performance depends on the type of carbon used and the characteristic of the wastewater. It is, like many other dye-removal treatments, well suited for one particular waste system and ineffective for another. Activated carbon is relatively expensive and has to be regenerated offsite with losses of about 10% in the thermal regeneration process (Robinson et al., 2001). Biomass referring to the dead plant and animal matter is also a suitable adsorbent for wastewater treatment. The adsorption of organic material onto various types of waste biomass such as sawdust (Poots et al., 1976a), peat (Poots et al., 1976b), chitin (McKay et al., 1982), bagasse pith (Al-Duri et al., 1990), carbonized wool waste (Malmary et al., 1985), wood chips (Nigam et al., 2000), maize cob (El Geundi, 1991), banana pith (Namasivayam et al., 1993), rice husk,
Water in the Textile Industry
hair, cotton waste, and bark (McKay et al., 1987) has been studied. The capacities of these materials have been examined through their adsorption of synthetic dyes. Two mechanisms are presented on the decoloration occurring in the biomass – adsorption and ion exchange. Both of them are influenced by dye–sorbent interaction, sorbent surface area, particle size, temperature, pH, and contact time. Biomass of different origins has been used for decoloration of acid, direct, and reactive dyes. Of all the described adsorbents, only a few have characteristics necessary for commercial use. Considering the price and binding capacity, quarternized lignocellulose-based adsorbents are the most appropriate for treating wastewatercontaining acid dyes. After the adsorption processes, the adsorbent needs to be regenerated, which adds to the cost of the process, and is sometimes a very time-consuming procedure. Decoloration with alternative materials such as zeolites, polymeric resins, ion exchangers, and granulated ferric hydroxide has also been studied in order to decrease adsorbent losses during regeneration. Filtration methods. Ultrafiltration, nanofiltration, and reverse osmosis can be used in the textile industry. These methods can be used not only for both filtering and recycling pigment-rich streams, but also for mercerizing and bleaching wastewaters. The specific temperature and chemical composition of the wastewater determine the type and porosity of the filter to be applied. The main drawbacks of membrane technology are high investment costs, potential membrane fouling, and the production of a concentrated dyebath which needs to be treated (Mishra and Tripathy, 1993; Xu and Lebrun, 1999). Coagulation and flocculation processes. These are widely used in several wastewater treatments in Germany and France. Coagulant agents such as aluminum sulfate, ferrous and ferric sulfate, ferric chloride, calcium chloride, copper sulfate, as well as several copolymers such as pentaethylene, hexamine, and ethylediene dichloride are used to form flocks with the dye, which are then separated by filtration or sedimentation. Coagulation–flocculation methods were successfully applied for decoloration of sulfur and disperse dyes, whereas acid, direct, reactive, and vat dyes presented very low coagulation–flocculation capacity. Polyelectrolyte can also be dosed during the flocculation phase to improve the flock settleability (Lee, 2000; Anjaneyulu et al., 2005). The main advantage of these processes is decoloration of the waste stream due to the removal of dye molecules from the dyebath effluents, and not due to a partial decomposition of dyes, which can lead to an even more potentially harmful and toxic aromatic compound. The major disadvantage of coagulation–flocculation processes is the production of sludge.
4.20.3.1.2 Chemical processes Some of the chemical processes are described in the following: Oxidation. The simplicity of its application makes oxidation the most commonly used chemical decoloration process. With conventional oxidation treatments, it is difficult to oxidize dyes (mainly for removing color) and toxic organic compounds in textile effluents. The development of so-called advanced oxidation processes (AOPs) has overcome the chemical limitations of conventional chemical oxidation
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techniques. The goal of AOPs is to generate free hydroxyl radicals (OHd) which may represent a rate increase of one to several orders of magnitude compared with normal oxidants in the absence of catalysts. Hydroxyl radicals oxidize the dyes and toxic organic compounds. In AOPs, oxidizing agents such as ozone and hydrogen peroxide are used with catalysts (Fe, Mn, and TiO2), either in the presence or in the absence of an irradiation source. Table 19 shows the oxidation potential of common species. Fenton’s reagent. Hydroxyl radicals are activated by Fe2þ (ferrous ions) in an acid solution (pH ¼ 3–4) (Table 20) from hydrogen peroxide. In this process, it is important to find the optimal concentration of hydrogen peroxide because excess of H2O2 acts as a scavenger of radicals, disturbs the COD measurements, and is toxic for microorganisms. This method is suitable for the oxidation of wastewaters, which inhibit biological treatment or are poisonous. Fenton’s reagent offers a cost-effective source of hydroxyl radicals and it is easy to operate and maintain. The advantages of this system are COD, color, and toxicity reduction and the disadvantage is sludge generation, through flocculation; impurities are transferred from the wastewater to the sludge, which contains the concentrated impurities and is still ecologically questionable. Conventionally, it has been incinerated to produce power, but such a disposal, according to some, is far from being environment friendly. To avoid this problem, Gnann et al. (1993) suggest the regeneration of Fe2þ from iron sludge at pHo1, with the so-called Fenton sludge recycling system (FSRS), in which Fe(III)-sludge deposition is eliminated. Fenton’s reagent as a decoloration agent has been studied by many authors and it is suitable for different dye classes: acid, reactive, direct, metal-complex, disperse, and vat dyes, as well as pigments. Low decoloration rates were observed when C.I. Vat red (50%) and C.I. Disperse Blue (0.5%) were treated (Slokar and Majcen Le Marechal, 1997). Studies on the decoloration and mineralization of commercial reactive dyes using solar Fenton and photo-Fenton reaction indicated good color removal. The use of solar light was proved to be clearly
Table 19
Oxidation potential of common oxidizing agents
Oxidizing agents
Oxidation potential (V)
Fluorine (F2) Hydroxyl radical (OHd) Atomic oxygen Ozone (O3) Hydrogen peroxide (H2O2) Potassium permanganate (KMnO4) Hypochlorous acid (HCIO) Chlorine (Cl2) Bromine (Br2) Molecular oxygen (O2)
3.06 2.80 2.42 2.07 1.78 1.67 1.49 1.36 1.09 1.23
Table 20 Degradation of hydrogen peroxide into hydroxyl radicals activated by Fe2þ Fe2þ þ H2O2-Fe3þ þ OHd þ HO
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beneficial for the removal of color, aromatic compounds, and TOC (Garcı´a-Montan˜o et al., 2006; Torrades et al., 2004). Ozone. Once dissolved in water, ozone reacts with a great number of organic compounds in two different ways, namely direct molecular and indirect free radical-type reactions. The direct reactions are often highly solute selective, slow, and are dominant in acidic solutions. They are suitable for opening aromatic rings by means of ozone cycloaddition. The indirect hydroxyl radical reactions are nonselective, fast, proceed more rapidly with increasing pH, and constitute a significant portion of ozonation at basic pH. Indirect attacks are suitable for mineralization of TOC (Zhao et al., 2004). Although the original purpose of oxidation with ozone is disinfection of potable water, it can also be used for removing many toxic chemicals from wastewater to facilitate the decomposition of detergents, chlorinated hydrocarbons, phenols, pesticides, and aromatic hydrocarbons (Science Applications International Crop., 1987). The advantages of ozonation include
• • • • • •
decoloration and degradation occur in a single step, danger to humans is minimal, no sludge remains, all residual ozone can be decomposed easily into oxygen and water, little space is required, and ozonation is easily performed (Oguz et al., 2005).
The disadvantage is its very short half-life in water – ozone decomposes in about 20 min. The time can be significantly shortened if compounds such as dyes are present (Rice et al., 1986). Ozone stability is affected by the presence of salts, pH, and temperature. If alkaline salts are present, the solubility of ozone is reduced, while neutral salts may increase its solubility (Mallevialle, 1982). Under alkaline conditions, ozone decomposes more rapidly than under acidic conditions. With increasing temperature, ozone solubility decreases (Perkins et al., 1980). Studies of decoloration presented by several authors revealed that ozone decolorizes all dyes, except nonsoluble disperse and vat dyes which react slowly and take longer time (Namboodri et al., 1994; Marmagne and Coste, 1996; Liakou et al., 1997). Color removal strongly depends on dye concentration. Ozonation alone has low TOC and COD removal. Species such as oxalic, glyoxalic, and acetic acids cannot be completely mineralized by ozone alone at least at neutral or acidic pH (Hoigne and Bader, 1983). To enhance the efficiency of ozonation, a combination of various advanced oxidation processes has been developed, such as ozon/ UV, ozon/H2O2, and catalytic ozonation. Ozone–UV. Combination of ozone with UV results in a net enhancement of organic-matter degradation due to direct and indirect production of hydroxyl radicals upon ozone decomposition and H2O2 formation (Table 21). UV radiation decomposes ozone in water and generates highly reactive hydroxyl radicals. Hydroxyl radicals oxidize organics more rapidly than ozone itself. The efficiency of ozone/UV treatment depends on operating temperature (at higher temperature the ozone solubility is lower), pH (degradation favors neutral or slightly alkaline medium), and ozone-flow rate. For comparison of both ozonation and
Table 21 Direct and indirect production of hydroxyl radicals in O3– UV process Direct O3 þ hn-O2 þ O O þ H2O-OHd þ OHd O þ H2O-H2O2 H2O2 þ hn-OHd þ OHd Indirect O3 þ H2O þ hn-O2 þ H2O2 H2O2 þ hn-OHd þ OHd
Table 22
Reactions between O3 and H2O2
Initiation HO2 þ O3 -HO2 þ O3 H þ þ O3 # HO3 -OH þ O2 H2 O2 þ O3 -H2 O þ 2O2
kr ¼ 2.2 106 l mol1 s1 kr ¼ 1.1 105 l mol1 s1 kro102 l mol1 s1
Promotion OH þ O3 -O2 þ HO2 OH þ H2 O2 -H2 O þ HO2 OH þ HO2 -H2 O þ O2
kr ¼ 1.1 108 l mol1 s1 kr ¼ 2.7 107 l mol1 s1 kr ¼ 7.5 109 l mol1 s1
ozone/UV process, the degradation of eight commercial azo dyes in water (Shu and Huang, 1995a) and a model dyehouse wastewater (Perkowski and Kos, 2003) has been studied. In both studies, the ozone/UV process did not significantly enhance the degradation rates; the dye competed with ozone for UV absorbance. However, ozone/UV treatment, in terms of COD removal, is more effective compared to that by ozone (Bes-Pia et al., 2003). Ozone/H2O2. Addition of hydrogen peroxide to ozone enhances the production of hydroxyl radicals. The aqueous reactions between ozone and hydrogen peroxide are rather complex. The mechanisms and the kinetics of the production of hydroxyl radicals from ozone and hydrogen peroxide are known. The reactions and reaction rate constants are shown in Table 22. In the initiation sequence, reactive OHd radicals are generated. During the promotion reactions, the hydroxyl radicals are converted into the peroxy radical. At acidic pH, H2O2 reacts only very slowly with ozone, whereas at pH values greater than 5, a strong acceleration of ozone decomposition by hydrogen peroxide has been observed. The ozone decomposition rate increases with increasing pH. Decoloration with O3/H2O2 process is applicable for direct, metal-complex, or blue disperse dyes. There are some problems with the decoloration of acid and red disperse dyes, though, as well as with mixtures of direct, metal-complex, disperse, and reactive-dye decoloration. The efficiency of the decoloration with O3/H2O2 for a few of the dyes is presented in Table 23. H2O2/UV. In H2O2/UV processes, hydroxyl radicals are formed when water-containing H2O2 is exposed to UV wavelengths of 200–280 nm. The most commonly used UV source is low-pressure mercury vapor lamps with a 254-nm peak emission.
Water in the Textile Industry Table 23
Decoloration of dyes with O3/H2O2
Table 25
Textile dye
Decoloration (%)
Time (min)
Red 219 Blue 186 Direct Yellow 44 Direct Yellow 50 Red 23 Red 26 Direct Red 5B Direct Blue 1 Direct Blue 25 Direct Blue 71 Disperse Yellow 3 Disperse Yellow 64 Red 13 Red 60 Red 279 Blue 60 Palanil Blue 3RT Sulfo/disperse dye Reactive Yellow 37 Reactive Yellow 125 Reactive Yellow 125 Remazol Yellow RNL Reactive Red 35 Reactive Red 195 Blue 27 Blue 221 Green 13 Reactive dyes Vat dyes Azoic dyes
100 85 100 100 100 100 99 100 100 90 95 100 100 100 99 100 90 98 93 98 100 93 99 100 94 100 98 100 80 87
5 1 0.5 0.5 0.5 0.5 45 0.5 0.5 7 1 4.5 0.7 1 98 0.7 31 30 4 2.5 7 4 4.5 6 0.9 9 4 1 30 30
Adapted from Slokar YM and Majcen Le Marechal A (1997) Methods of decoloration of textile wastewaters. Dyes and Pigments 37(4): 335–356.
Table 24
The main reactions that occur during the H2O2/UV process
H2O2 þ hn-OHd þ OHd RH þ OHd-H2O þ Rd-further-oxidation
Problems such as sludge formation and regeneration, and increased pollution of wastewater caused by ozone, can be avoided by oxidation with hydrogen peroxide activated with UV light. The only chemical used in the treatment is H2O2, which, due to its final decomposition into oxygen, is not problematic. The most direct method for generation of hydroxyl radicals is through the cleavage of H2O2. Photolysis of H2O2 yields hydroxyl radicals by direct process with a yield of two radicals formed per photon absorbed at 254 nm. Hydroxyl radicals can oxidize organic compounds (RH)-producing organic radicals (Rd), which are highly reactive and can be further oxidized (Table 24) (Tuhkanen, 2004). The maximum absorbance of H2O2 is needed to generate sufficient hydroxyl radicals because of low absorption coefficient. However, high concentration of H2O2 scavenges the radicals, making the process less effective, while low concentration of hydrogen peroxide does not generate enough hydroxyl radicals to be consumed by the dye and this leads to
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Reactions of H2O2 as a radical scavenger
OH þ H2 O2 -HO2 þ H2 O HO2 þ H2 O2 -HO2 þ H2 O þ O2 HO2 þ HO2 -H2 O2 þ O2
a slow rate of oxidation. Therefore, an optimum hydrogenperoxide dose needs to be verified experimentally (Table 25). The rate of dye removal is influenced by the intensity of UV radiation, pH, dye structure, and dyebath composition. In general, decoloration is most effective at neutral pH medium, at higher UV radiation intensity (1600 W rather than 800 W), with an optimal H2O2 concentration, which is different for different dye classes, and with a dyebath that does not contain oxidizing agents having an oxidizing potential higher than that of peroxide. According to Shu and Huang (1995b) acid dyes are the easiest to decompose, and with an increasing number of azo groups, the decoloration effectiveness decreases. Yellow and green reactive dyes need longer decoloration times, while other reactive dyes as well as direct, metal-complex, and disperse dyes are decolorized quickly. In the group of blue dyes examined, only blue vat dyes were not decolorized. For pigments, H2O2/UV treatment is not suitable, because they form a film-like coating on the UV lamp, which is difficult to remove. Several authors (Georgiou et al., 2002; Neamtu et al., 2002; Galindo and Kalt 1999; Colonna et al., 1999) reported complete decoloration of reactive and azo dyes in 30–90 min. The results indicated that H2O2/UV processes could be successfully used for the decoloration of acid, direct, basic, and reactive dyes but it proved to be inadequate for vat and disperse dyes (Yang et al., 1998). A comparative study between ozone and H2O2/UV was carried out on simulated reactive dyebath effluent containing a mixture of monochlorotriazinetype reactive dyes and various auxiliary chemicals. The H2O2/ UV process presented the decoloration rates close to those rates obtained with ozone but at a lower cost (Alaton et al., 2002). H2O2/UV systems may be set up in a batch or in a continuous column unit (Namboodri and Walsh, 1996). Decoloration of some dyes with H2O2/UV is presented in Table 26. Ultrasound. Sonolysis is a relatively innovative advanced oxidation process and was found to be a suitable method for the destruction of textile dyes. The ultrasonic irradiation of liquids generates cavitation (typically in the range 20–1000 kHz). Cavitation is a phenomenon of micro-bubble formation. Micro-bubbles grow during the compression/rarefaction cycles until they reach a critical size, and implode generating heat and highly reactive radical species. Inside the cavitation bubbles, the temperature and pressure rise to the order of 5000 K and 100 MPa, respectively. Under such conditions, water molecules degrade releasing hydroxyl radicals (OHd) and hydrogen radicals (Hd) as mentioned in Table 27. These radical species can either recombine or react with other gaseous molecules within the cavity, or in the surrounding liquid, after their migration. Pyrolitic and radical reactions inside, or near, the bubble and radical
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reactions in the solution are two major pathways for sonochemical degradation. Hydrophilic and nonvolatile compounds mainly degrade through hydroxyl-radical-mediated reactions in the solution, while hydrophobic and volatile
Table 26
Decoloration of dyes with H2O2/UV
Textile dye
Decoloration (%)
Time (min)
Reactive Yellow 37a Reactive Yellow 125a Remazol Yellow RNLa Reactive Red 35a Reactive Red 195b Reactive Black 5c Acid Yellow 17d Orange 10d Blue 21b Blue 27a Green 13a Vat Bluea Red 1d Red 14d Red 18d Blue 186e Black 1d Direct Yellow 4d Direct Blue 71a Palanil Blue 3RTa
85 96 85 100 100 100 98.2 100 95 100 93 15.5 99.9 100 99.1 80 89.9 83.2 98.5 96
8 8 8 8 60 4 40 60 150 5 8 10 30 60 40 10 60 60 3 10
a
Hoigne and Bader (1983). Liakou et al. (1997). c Ince and Go¨rnec (1997). d Bes-Pia et al. (2003). e Pittroff and Greorg (1992). b
Table 27
Radical formation and depletion during water sonolysis d
H2O-))) OH þ Hd OHd þ Hd-H2O 2OHd-H2O þ Od 2OHd-H2O2
Table 28
species degrade thermally inside or in the vicinity of the bubble. Reactive azo dyes are nonvolatile, water-soluble compounds and their passage into the gas cavity is unlikely. Hence, oxidative radical reactions in the bulk solution are expected to be the major route for their destruction. According to several studies, it is difficult to obtain the total mineralization (degradation to carbon dioxide, short-chain organic acid, oxalate, formate, and inorganic ions such as sulfate and nitrate) of the complex textile dyes with ultrasound alone. For this reason, the combination of ultrasound with other advanced oxidation processes is a more convenient approach in the remediation of such pollutants. Sonochemical degradation of textile dyes has become quite an interesting research area confirmed by several reports over the last few years (Vajnhandl and Majcen Le Marechal, 2005). In Table 28, a comparison of individual AOP is given.
4.20.3.1.3 Biological treatment processes Biological degradation or breakdown by living organisms is the most important removal process of organics, which are transferred from industry processes into solid and aquatic ecosystems. The application of microorganisms for the biodegradation of synthetic dyes is an attractive method and offers considerable advantages. The process is relatively inexpensive, the running costs are low, and the end products of complete mineralization are not toxic. An extensive review of large numbers of different species of microorganisms tested for decoloration and mineralization of different dyes has been published by Forgacs et al. (2004). The efficiency of biological-treatment systems is greatly influenced by the operational parameters. To produce the maximum rate of dye reduction, the level of aeration, temperature, pH, and redox potential of the system must be optimized. The concentration of the electron donor and the redox mediator must be balanced with the amount of biomass in the system and the quantity of the dye present in the wastewater. The compounds present (sulfur compounds and salts) in the wastewater may have an inhibitory effect on the
Technical comparison of oxidative decoloration
Oxidation process
Advantages
Disadvantages
Fenton
Effective decoloration of both soluble and insoluble dyes. Simple equipment and easy implementation. Reduction of COD (except with reactive dyes). No alternation in volume. Simple equipment and implementation. Reduction of COD (except with reactive dyes). Applied in gaseous state. No alteration of volume. No sludge production. Effective for azo dye removal. No sludge formation. No salt formation. Short reaction times. Very short reaction times for reactive dyes. No sludge formation. No salt formation. Short reaction times. Reduction of COD. Simplicity in use. Very effective in integrated system.
Sludge formation. Long reaction times. Salt formation. Hazardous waste. Prohibitively expensive.
FSR (Fenton sludge recycling system) Ozone Ozone/H2O2
H2O2/UV Ultrasound
Salt formation. Formation of gasses (H2, O2 during electrolysis). Short half-life (20 min). Not suitable for disperse dyes. Releases of aromatic amines. Not applicable for all types. Toxicity, hazard, problematic handling. No COD reduction. Additional load of water with ozone. Not applicable to all types of dyes. Requires separation of suspended solid particles. Relatively new method and awaiting full scale application.
Water in the Textile Industry
dye-reduction process. For these reasons, it is important to study the effect of these factors on decoloration before the biological system can be used to treat industrial wastewater (Pearce et al., 2003). Biodegradation processes may be anaerobic, aerobic, or involve a combination of both. Anaerobic biodegradation. Under anaerobic conditions, a low redox potential (o 50 mV) can be achieved, which is necessary for the effective decoloration of dyes. Color removal under anaerobic conditions is also referred to as dye reduction. Many bacteria under anaerobic conditions reduce the highly electrophilic azo bond in the dye molecules and produce colorless aromatic amines. The anaerobic decoloration of azo dyes was first investigated using intestinal anaerobic bacteria (Allan and Roxon, 1974; Brown, 1981; Chung et al., 1992). Later, it was found that azo dyes can also be decolorized with various other anaerobical cultures (Brown and Laboureur, 1983; Beydilli et al., 1998; Donlon et al., 1997). The efficacy of various anaerobic-treatment applications for the degradation of a wide variety of synthetic dyes has been demonstrated in several experiments. The exact mechanism of azo dye reduction is not clearly understood yet. There may be different mechanisms involved, such as enzymatic (Haug et al., 1991; Rafii et al., 1990), nonenzymatic (Gingell and Walker, 1971), mediated (Kudlich et al., 1997), intracellular (Mechsner and Wuhrmann, 1982; Wuhrmann et al., 1980), extracellular (Carliell et al., 1995), and various combinations of these mechanisms. A complete anaerobic mineralization of the azo dye azodisalicylate was observed under methanogenic conditions (Razo-Flores et al., 1997). The reduction of azo dye under anaerobic conditions strongly depends on the presence and disponibility of the cosubstrate. It acts as an electron donor for the azo dye reduction. The decoloration of reactive water-soluble azo dyes was achieved under anaerobic conditions using glucose as a co-substrate (Carliell et al., 1996). Anaerobic decoloration of reactive dyebath effluents with tapioca as a co-substrate also enhances color-removal efficiency (Chinwetkitvanich et al., 2000). The other suitable co-substrates were hydrolyzed starch, yeast extract, and a mixture of acetate, butyrate, and propionate. Much effort has been devoted to the study of the influence of various modern technologies on the decomposition rate of the dyes and the effect of the presence of the other compounds in the media. It has been recently established that the development of high rate systems, in which the hydraulic-retention times are decoupled from the solid-retention times, facilitates the removal of dyes from textile-processing wastewater (Rice et al., 1986). The effect of nitrate and sulfate salts used in textile dyeing on the microbial decoloration of a reactive azo dye has been studied. The results indicated that nitrate delays the onset of decoloration while sulfate did not influence the biodegradation process (Carliell et al., 1998). The reduction of azo dyes proceeds better under anaerobic thermophilic conditions than under mesophilic conditions, although the thermophilic process seems to be less stable compared to the mesophilic process (Willetts et al., 2000).
701
Carliell et al. (1994) studied the biodegradation of reactive dyes and they decolorized 80% of a range of tested dyes. From a detailed study of a selected dye, it was proposed that this occurred via a reduction mechanism. The results were supported by tentative chemical identification of the dyedegradation products. Hu (1994) isolated Pseudomonas luteola bacteria; after a 6-month adaptation in colored wastewater, he obtained microorganisms capable of reductive cleavage of the azo group in the dye. Decoloration with these microorganisms was complete within 4 days. Van der Zee et al. (2001) studied the decoloration of 20 selected azo dyes by granular sludge from an upward-flow anaerobic sludge-bed reactor and for all the azo dyes tested, complete reduction was achieved. Aromatic amines, due to azo dye reduction, are not commonly degraded under anaerobic conditions. Many aromatic amines were tested, but only a few were degraded. Some aromatic amines, substituted with hydroxyl or carboxyl group were degraded under methanogenic and sulfate-reducing conditions (Kalyuzhnyi et al., 2000; Kuhn and Suflita, 1989; Razo-Flores et al., 1999). Aerobic biodegradation. It is a process that often takes place in the environment, for example, in natural ecosystems such as soil or surface waters, and it is often associated with technical systems such as wastewater-treatment plants. Although for long, it was considered that azo dyes cannot readily metabolize under aerobic conditions, some specific aerobic bacterial cultures were found to be able to reduce the azo linkage via an enzymatic reaction. The aerobic conversions of sulfonated azo dyes were studied by Heiss et al. (1992) and Shaul et al. (1991), and sometimes even a complete mineralization of sulfonated azo dyes was found. In some studies, aerobic color removal of certain azo dyes was achieved, but all these stains required an additional energy and carbon source for growth. Since the supply of this additional substrate could have easily led to the formation of anaerobic microniches, the occurrence of anaerobic azo dye reduction certainly cannot be excluded (Govindaswami et al., 1993; Horitsu et al., 1977; Wong and Yuen, 1996; Zissi et al., 1997). The aerobic biodegradation of different aromatic amines (aniline (Lyons et al., 1984), carboxylated aromatic amines (Stolz et al., 1992), chlorinated aromatic amines (Loidl et al., 1990), benzidines (Baird et al., 1977), and sulfonated aromatic amines) has been extensively studied and many of these compounds were found to be degraded. Sulfonated aromatic amines are difficult to degrade. Combination of anaerobic/aerobic biodegradation. Although the anaerobic reduction of azo dyes is generally more satisfactory than aerobic degradation, carcinogenic aromatic amines, as products of anaerobic degradation, have to be degraded by an aerobic process. Diverse technologies for the successive anaerobic/aerobic treatment of textile wastewater have been developed. Anaerobic/aerobic conditions can be implemented by spatial separation of the two sludges using a sequential anaerobic/aerobic reactor system (Zitomer and Speece, 1993). These conditions can also be imposed on a single reactor in the so-called integrated anaerobic/aerobic reactor system (Field et al., 1995).
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4.20.3.2 Reuse New ecolabels for textile products and tighter restrictions on wastewater discharges are forcing textile wet processors to reuse process water and chemicals (Vandevivere et al., 1998). It is quite difficult to define a general quality standard for textilewater reuse because of the different requirements of each fiber (silk, cotton, polyester, etc.) of the textile process (e.g., scouring, desizing, dyeing, and washing) and because of the different quality required for the final fabric (Water Treatment Solutions, 2010). The actions aimed at the reduction of pollution and reuse of wastewater may normally be ranked in the following order according to their cost effectiveness: 1. prevention of pollution generation, 2. treatment of polluted streams close to the source of pollution (start-of-pipe approach), and 3. treatment of the final effluents (end-of-pipe approach). Pollution-prevention actions can normally succeed in all companies, while the start-of-pipe approach is mostly indicated in medium or big enterprises and the simplicity of the end-of-pipe approach makes it suitable in small as well as medium enterprises (Matioli et al., 2002).
4.20.3.2.1 Pollution-prevention techniques Pollution-prevention techniques have proved to be an effective means to improve process efficiency and to increase company profits, and at the same time, they minimize environmental impact. During the implementation of each of these techniques, the specific conditions must be carefully considered and every option and change must be examined, to understand how it could affect air, land, and water-pollutant releases (Matioli et al., 2002). Some of the pollution-prevention techniques that can be adopted are as follows: 1. Quality control for raw materials. Textile companies can reduce waste emissions by working with suppliers to find out less-polluting raw materials. Pre-screening raw materials is a useful practice to determine interactions among processes, substrates, and other chemicals with the aim to reduce waste production (Matioli et al., 2002). 2. Chemical substitution. Textile manufacturing is a chemically intensive process, and therefore a primary focus for pollution prevention should be on textile process chemicals. Opportunities for chemical substitution vary substantially among mills because of differences in: (US EPA/625/R-96/004, 1996) • environmental conditions, • process conditions, • product, and • raw materials. Possible actions are replacement of chemicals as desizing agents, dyes and auxiliaries with less-polluting ones, and replacement of chemical treatment in some processes with mechanical or other nonchemical treatment (Matioli et al., 2002). 3. Process modification. Optimization of the processes can be obtained by modifying some operations. Examples of possible modifications are (Matioli et al., 2002)
•
substitution of dyeing machines using low liquor ratio (equipment able to substantially reduce bath ratio and allow considerable savings of energy, water, dyes, and chemicals), • optimization of process conditions (temperature and time), and • combining operations to save energy and water (combining scouring and bleaching). 4. Equipment modification. An effective way to reduce waste is also by modifying, retrofitting, or replacing equipment and introducing automation (Matioli et al., 2002). 5. Good operating practices. A suitable way to prevent pollution without changing industrial processes is introduction of pollution-prevention procedures, including pollution-prevention objectives in research, new facility design, and ad hoc worker-training programs (Matioli et al., 2002).
4.20.3.2.2 Chemicals and water reuse and recycle: Start-of-pipe approach Recycling (reusing water and chemicals in the same process that produced the effluent) can save water, chemicals, and energy as well. An example is the reuse of exhausted hot dyebaths to dye further batches of material. In order to reuse the dyebath, it is necessary to determine the exact quantities of residual chemicals remaining in the dyebath. As a following step, to respect the characteristics demanded by the next dyeing cycle, the dyebath must be reconstituted by adding water, auxiliary chemicals, and dyestuffs (Matioli et al., 2002; EPA/310-R-97-00, 1997). Several examples of water reuse without treatment are based on the recovery of the water used in rinsing operation. Implementation of countercurrent washing (reusing the last contaminated water from the final wash for the next-to-last wash and so on) can significantly reduce the overall water consumption and is already applied in continuous textile operations. A systematic analysis of the water networks is required every time the overall use of water needs to be optimized and new options of water treatment and reuse have to be evaluated. Tools such as pinch analysis provide a formal procedure to determine near-optimal designs of energy and mass-transfer networks (Matioli et al., 2002; Majozi et al., 1998). When applied to water-use optimization, pinch analysis allows the identification of reuse, regeneration, and treatment opportunities. This approach normally generates start-of-pipe solutions implementing specific-process effluent treatment. The process-integrated wastewater treatment required by start-of-pipe solutions is based upon the possibility of efficient, reliable, cost-effective, and easy-to-operate treatment of single wastewater streams. These results can be obtained by proper applications of membrane technology (Matioli et al., 2002).
4.20.3.2.3 Process-water reuse and recycle: End-of-pipe approach In some cases, the classical end-of-pipe approach for reuse and recycling of industrial final effluents can also be efficient and cost effective. It fits very well in some typical European areas that can be defined as ‘textile districts’. A textile district refers to an area where many textile factories, mainly small and
Water in the Textile Industry
medium enterprises, are widespread and utilize the same water and wastewater facilities (Matioli et al., 2002). The end-of-pipe treatment was the first approach examined for cleaning up the total effluent flow in order to meet the standards for reuse. End-of-pipe treatment involves multistage-process combinations typically composed of biological and physicochemical techniques. Recently, the interest in membrane processes applied to textile-wastewater reuse is increasing, thanks to technological innovations that render them as reliable and feasible alternatives to other systems (Schoeberl et al., 2004). Membrane systems can successfully remove the large amount of suspended solids in wastewater (Chen et al., 2005). Centralized treatment plant for mixed industrial and municipal wastewater uses an aerobic biological stage. Some compounds are completely degraded, while others (dyes, surfactants, and their metabolites) are either absorbed on the sludge or discharged into the final effluent. Textile wastes contain poorly degradable organics (at least in aerobic conditions). Many contain toxicants, which are also often poorly biodegradable. Traditional aerobic biological process presents serious technical limitations for the purification of textile wastewaters (Matioli et al., 2002). The EU founded the Research and Technological Development (RTD) project Integrated Waste Recycling and Emission Abatement in the Textile Industry (EU, 1999) proposed several combined process modules to improve the actual wastewater-treatment plants, aimed at the reuse of final effluents: 1. A module for chemical precipitation of heavy metals and adsorption of dyes on anaerobic sludge (consisted in a pretreatment option) (Terras et al., 1999; O’Neill et al., 1999). 2. Enhancement of the biological treatment to a sensor-protected aerobic stage to remove biodegradable organics and to oxidize reduced nitrogen compounds while monitoring potential toxicity (Terras et al., 1999; Massone et al., 1998; Guwy et al., 1998). 3. The optimization of the final polishing involving various tertiary treatment lines to bring the water up to the standard required for use by the industries (Bergna et al., 1999; Bianchi et al., 1999; Rozzi et al., 1997, 2000). Posttreatment for mixed textile and domestic effluents has been successfully tested on the following unit process: ozonation, clariflocculation, multimedia filtration, granular activated carbon adsorption, ceramic crossflow and hollow-fiber microfiltration, nanofiltration, and low-pressure reverse osmosis. All the processes were investigated at medium and large pilot scale (Matioli et al., 2002). New advanced respirometric methodologies based on respirometry and titration may be used as wastewater-characterization techniques. They are particularly suited to evaluate the possible effects of a given wastewater on the final wastewater-treatment plant, due to their organic biodegradable and refractory load and inhibitory potential. The use of these characterization methods makes it possible to prevent treatment problems due to toxic discharges (Rozzi et al., 1999). The fee lever based on the treatability of the discharges can also be used to design the influent wastewater in a given
703
treatment plant, discouraging the discharge of refractory and inhibitory compounds. It can also lead to the introduction of cleaner technologies when an industry, billed with high fees for the presence of inhibitory compounds in its wastewater, is pushed toward the application of pollution-prevention techniques. The concept of waste design should not be limited to an offline procedure of characterization and of request to industries of qualitative or quantitative changes to their discharges. This concept should be extended to an online management system, based on a network of sensors, actuators, and facilities that can allow the plant manager to detect in the sewer (or before to discharge to it) the presence of excess hydraulic loading, organic or nutrient loading, or toxicants, and put in operation measures that can allow to maintain an optimal treatment result (Matioli et al., 2002; Bortone et al., 1997).
4.20.4 Conclusions Textile processing is one of the largest and oldest industries worldwide and it is responsible for substantial resource consumption and pollution. The wet processing, that is, pretreatment, dyeing, printing, and finishing, is especially polluting and resource consuming in terms of water, energy, and chemicals and like in most industries, freshwater is used in all processes with almost no exceptions. Textile industry has significant impact on the aquatic system, both by consuming a lot of water, freshwater sources, and also by discharging effluents into the environment. Water savings, reclamation, and reuse in industry are topics of increasing economic interest due to increasing water scarcity and costs. For this reason, research and development activities within this topic are increasing, methods and tools for analyzing water savings and reuse possibilities are being developed, and solutions are being implemented. The problem of water scarcity and the need for a rational water management has raised an interest in the use of recycled and reclaimed water as well as further water-loop closure. The typical textile SME today does not implement water reuse, while fresh high-quality water is used in all the production processes. Furthermore, the process effluents are mixed and discharged after onsite or centralized treatment in conventional wastewater-treatment plants. Despite the fact that during the last decades, new knowledge and technologies related to process-water production, wastewater treatment, and water-loop closure have been developed and implemented, current available technologies for textile wastewater treatment are often limited in efficiency and cost, and are not environmentally selective enough. On the other hand, it is not always clear which treatment lines are best suited to achieve the desired water quality at the lowest cost. Besides, textile companies are mainly SMEs and the small scale could represent a problem, because the water streams might have very different compositions. Over the last 30 years, drought and water scarcities have cost the European economy an estimated h100 bn. The most severe impacts of climate change that the world is facing are related to water. Climate change is intensifying the hydrological cycle. Risks from flood, drought, and coastal
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inundation, melting of glaciers and changes in the flow regimes of rivers are growing. Despite an understanding of the dangers to the economy, social stability, and the environment, not enough attention was given until recently to reduce the impacts of climate change on water and to increase adaptation efforts. In light of this, the vision of technological platforms (Water Supply and Sanitation Technology Platform (WSSTP), textile platform) and industrial associations (European Water Partnership (EWP) and European Apparel and Textile Organization (EURATEX)) is trying to follow some new ideas and approaches that would bring water to the forefront of a comprehensive strategy and promote adaptation measures across all water-related sectors.
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4.21 Water Availability and Its Use in Agriculture D Molden, International Water Management Institute, Battaramulla, Sri Lanka M Vithanage, Institute of Fundamental Studies, Kandy, Sri Lanka C de Fraiture, International Water Management Institute, Accra, Ghana JM Faures, Food and Agriculture Organization of the United Nations (FAO), Rome, Italy L Gordon, Stockholm University, Stockholm, Sweden F Molle, Institut de Recherche pour le De´veloppement and International Water Management Institute, Colombo, Sri Lanka D Peden, International Livestock Research Institute (ILRI), Addis Ababa, Ethiopia & 2011 Elsevier B.V. All rights reserved.
4.21.1 Water Availability and Its Use in Agriculture 4.21.1.1 Sources of Water for Agriculture, Their Distribution, Use, and Possible Climate Change Effects 4.21.1.1.1 Green water 4.21.1.1.2 Agricultural water use in river basins 4.21.1.1.3 Open, closing, and closed river basins 4.21.1.1.4 Groundwater 4.21.1.1.5 Wetlands 4.21.1.1.6 Water consumption 4.21.1.1.7 Water use 4.21.1.1.8 Climate change, agriculture, and water 4.21.1.1.9 Drivers of water use 4.21.1.2 Physical and Economic Water Scarcity 4.21.1.3 Future Demands for Water 4.21.1.4 Future Scenarios for Rainfed and Irrigated Agriculture 4.21.2 Productive Use of Agricultural Water 4.21.2.1 Water Productivity in Agriculture 4.21.2.2 Rainfed Agriculture Productivity 4.21.2.3 Irrigated Agriculture and Productivity 4.21.2.4 Livestock 4.21.2.5 Aquaculture and Fisheries 4.21.3 Environmental and Health Implications of Agricultural Water Use 4.21.3.1 Impact on Rivers, Wetlands, and Biodiversity 4.21.3.1.1 Aquatic ecosystems 4.21.3.1.2 Terrestrial ecosystems 4.21.3.2 Health Impacts 4.21.3.3 Environmental and Health Mitigation 4.21.4 Water Governance 4.21.4.1 Definition 4.21.4.2 Types of Governance for River Basin Management 4.21.4.3 Basin Governance Challenges Acknowledgments References
4.21.1 Water Availability and Its Use in Agriculture With growing populations, shifting geographies, and changing dietary patterns, agriculture and food production face formidable challenges in the near future. Understanding issues of water availability and its use is fundamental for assessing and responding to these challenges. The following section examines the topics of water availability and use as they relate to agricultural production. While 3% of Earth’s total water volume is fresh (most of it is found in the form of ice in polar regions), only 1% is easily accessible for human use and is found in the physical forms of lakes, rivers, and shallow
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aquifers (UN, 1997). Water for food and agricultural production is the largest use of this finite resource.
4.21.1.1 Sources of Water for Agriculture, Their Distribution, Use, and Possible Climate Change Effects Conventionally, agricultural water resources have been thought of in terms of surface water and groundwater. This approach, however, can be limiting. Besides considering surface- and groundwater, accounting for rainfall and soil moisture, as they factor into hydrological and agricultural systems,
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allows for a systematic understanding of water. The concepts of blue water and green water are useful in thinking about water availability in relation to a broader range of agricultural practices and a variety of users (CA, 2007). Blue water refers to water found in rivers, lakes, reservoirs, and aquifers. In addition to its use in agriculture, blue water is the measured and managed freshwater resource needed to meet domestic, commercial, and hydroelectric power demands while also functioning to sustain ecosystems (UN, 2006). Of total renewable blue water resources, 9% is used annually. Cities and industries extract 1 200 km3 of blue water per year but return more than 90% of it. This return is often of degraded quality and much of the flow returns to the sea, where it supports coastal ecosystems (Figure 1). Green water refers to
soil moisture available to plants generated by infiltrating rainfall. Green water is the main source of water for rainfed agriculture, whereas blue water is the main source for irrigated agriculture. Rainfed agriculture strictly depends on green water only, whereas irrigated agriculture uses blue water to supplement soil moisture. By adding blue water to crops, farmers can maintain soil moisture in dry periods and allow their crops to fulfill yield potentials. Through the process of evapotranspiration both green and blue water are ‘‘consumed’’ by vegetation and not returned to the system like in the case of other sections. The implications of green and blue water use are quite different. Increased evapotranspiration of blue water reduces stream flow and groundwater levels. Agricultural
Global water use Rainfall (thousands of cubic kilometers per year) 110 100%
Green water Bioenergy Forest products Grazing lands Biodiversity Landscape 56%
Blue water Rivers Wetlands Lakes Groundwater
Soil moisture from rain Crops Livestock Rainfed agriculture 4.5%
Water storage Aquatic biodiversity Fisheries
Crops Livestock Aquaculture Irrigated agriculture 0.6% 1.4%
Open water evaporation 1.3%
Green water
Cities and industries 0.1%
Blue water
Ocean 36%
Landscape Dam and reservoir
Landscape Irrigated agriculture
Wetlands
Rainfed agriculture
Cities
Figure 1 Global water uses. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007). From Oki and Kanae (2006) Global hydrological cycles and world water resources. Science 313(5790): 1068–1072; UNESCO–UN World Water Assessment Programme (2006) Water: A Shared Responsibility, The United Nations World Water Development Report 2. New York: UNESCO and Berghahn Books.
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4.21.1.1.2 Agricultural water use in river basins
evapotranspiration is necessary for food production, and generally as food production increases, so does evapotranspiration. Increased evapotranspiration from green water sources is usually due to expansion of agricultural land area, a terrestrial impact, but has less impact on blue water flows. Still, any change in land use can affect river flows. In South Africa, recognition of the effects of stream flow-reducing activities has led to initiatives to control commercial forestry and to remove invasive tree species in order to reduce evapotranspiration and increase river flow (Hope, 2006).
The remaining 20% of crop evapotranspiration is from blue water drawn from surface- and groundwater sources. Blue water resources are, systematically, part of hydrological regions called river basins. River basins bounded by the area that catches water and directs it to common outlets. Basins and serve as important units of analysis because they connect various water in the basins uses. A change in use in one area often influences other uses of water. Efforts to control rivers go back many thousands of years, similarly, the practice of using these physical areas as regional units of organization for planning, developing, and managing water. More recently, in the latter half of the twentieth century, major dams were constructed which resulted in the multipurpose development and management of river basins. Hydroelectric power, flood control, water storage, and navigation became linked politically, economically, and ecologically in these river systems. Meanwhile, investment in irrigation accelerated rapidly in the 1960s and the 1970s, with irrigated area expansion in developing countries at 2.2% a year reaching 155 million hectares in 1982. During the same period, total global irrigated lands rose from 168 to 215 million hectares (Carruthers et al., 1997).
4.21.1.1.1 Green water Globally, about 80% of agricultural evapotranspiration comes directly from green water (Figure 2). This implies that the majority of the world’s agricultural production comes predominantly from rainfed lands despite major increases in large-scale irrigation infrastructure over the past half century. Some 55% of the world’s gross value of crop production is grown under rainfed agriculture on 72% of harvested land (Table 1). There are, however, large geographical differences in the percentages of rainfed and irrigated agricultural lands. For instance, over 95% of sub-Saharan Africa’s cultivated lands are strictly rainfed agriculture. Similarly, Latin America’s cultivated lands are 90% rainfed agriculture. In several countries of the Near East and North Africa, more than 40% of cultivated areas is irrigated. Meawhile investment in irrigation accelerated, in the 1960s and 1970s, from about 150 million hectares to a present total of over 270 million hectares (Faures et al., 2007). Hence, irrigated agriculture is relatively important in Asia and North Africa, while rainfed agriculture dominates in sub-Saharan Africa and Americas.
4.21.1.1.3 Open, closing, and closed river basins When a river basin can supply water to meet withdrawal demands and maintain its ecological functions, it is considerd an open basin (Seckler, 2006). A river basin is closing or closed when the volume of water use approaches or exceeds the volume of discharge. Often this is a problem of overcommitment, where water resources have been allocated beyond availability. As infrastructure develops around rivers,
More than half of production from rainfed areas
More than half of production from irrigated areas
More than 75% of production from rainfed areas
More than 75% of production from irrigated areas Global total: 7130 km3 (80% from green water, 20% from blue water) 780
220
650
235
1670 Blue water
Green water
905
1080
1480 110
Figure 2 Food crop evapotranspiration from rain and irrigation. Production refers to gross value of production. The pie charts show total crop water evapotranspiration in km3 by region. From International Water Management Institute analysis done for the comprehensive assessment for water management in agriculture using the Watersim model. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
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Water Availability and Its Use in Agriculture
Table 1
Global water and land statistics 3
Water (km ) Use
Land (millions of hectares) Statistics
Use
Total precipitation over continents 11 000 Vapor flow back to the Runoff to the oceans 40 000 atmosphere 70 000 Evapotranspiration Biomass consumed by grazing livestock Rainfed crops Irrigated crops Irrigation Rainfall Municipal use Industrial use Reservoirs
Total terrestrial land 13 000
Withdrawals
840 4910 2664 1570 650 53 88 208
Statistics
Grazing lands
3430
Rainfed harvested lands Irrigated cultivated lands
860 Harvested 340a
381 785
a
Of which 277 are equipped. From For water withdrawal statistics and equipped irrigation area, FAO (2006a); for evapotranspiration, International Water Management Institute analysis using the Watersim model; for harvested irrigated crop area, Chapter 3 on scenarios; for biomass consumed by grazing livestock, Stockholm Environment Institute calculations for the Comprehensive Assessment of Water Management in Agriculture; for municipal, industrial, and reservoir use, Shiklomanov (2000); for land statistics, FAO (2006b). Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
streams can become increasingly diverted, controlled, and utilized. In closed or closing basin scenarios, water users (especially those downstream) will not have full access to withdraw from the resource. Many rivers around the globe are closing or closed. Closing basins are sensitive to seasonal and inter annual variations of rainfall. Meanwhile, the ecological functions of a river will suffer, it is important for rivers to have adequate flow for fish and wildlife habitat, flushing sediments, diluting pollutants, preventing salinity intrusion, and sustaining estuarine and costal ecosystems. Increasing supply through inter-basin transfers is a common response to reopen closed basins, and desalinization to increase supply is a much discussed option (Falkenmark and Molden, 2008). Many closing basins are typically under stress for 1–6 months a year. China’s Yellow River dried up for the first time in 1972. In 1997, the dry-up lasted 226 days and reached 700 km upstream (Ren and Walker, 1998). The Colorado in the United States, the Indus flowing through India and Pakistan, the Murray-Darling in Australia, and most rivers in the Middle East and Central Asia are also severely overcommitted. Even basins in monsoon regions, such as Chao Phraya River in Thailand and the Cauvery River in India, experience months of closure, when salinity creeps inland as outflows of freshwater do not flush into the sea.
4.21.1.1.4 Groundwater The Earth’s fresh groundwater resources are estimated at approximately 10 000–12 000 km3, more than 200 times the volume of global annual rainfall. Only a tiny proportion, approximately 12 000 km3, of the total volume of groundwater reserves is recharged each year, compared to the large volume in stock (Doll and Fiedler, 2008). It is
estimated that, on average, 2091 m3 per capita are withdrawn from groundwater stores, with agriculture withdrawing the majority. About 2 billion people worldwide use groundwater, making it the single most utilized natural resource on the planet. The estimated annual use of groundwater is between 600 and 700 km3 (Struckmeier et al., 2005) and it keeps increasing. In the United States, for example, groundwater use in irrigation water has increased from 23% in 1950 to 42% in 2000 (Winter et al., 1998). This trend is reflected around the globe and particularly in Asia. There are many reasons why irrigation is a major user of groundwater. For farmers, the water is available when it is needed, is of reasonable quality, and very often can be abstracted without gaining permission or consulting with other users, a situation which is often much simpler than obtaining irrigation water from a canal system. Although agriculture is the largest user of groundwater, domestic dependence on groundwater use is increasing. Groundwater has historically supplied domestic water requirements in numerous urban and rural human settlements around the world. According to one estimate, more than half of the world’s population relies on groundwater for its drinking water supply (Coughanowr, 1994). In Spain, from 1960 to 2000, groundwater use increased from 2 to 6 km3 yr1 (Martinez-Cortina and Hernandez-Mora, 2003). In the Indian subcontinent, groundwater use soared from around 10–20 km3 yr1 before 1950 to 240–260 km3 yr1 by the year 2000 (Shah et al., 2003). In the United States, the volume of groundwater used as irrigation water increased from 23% in 1950 to 42% in 2000 (Winter et al., 1998). Chinese history records occasional cases of farmers lifting water from shallow wells by barrels to irrigate vegetables; however, North China had very little irrigation
Water Availability and Its Use in Agriculture
until 1950, and its tubewell irrigation revolution took off only after 1970. In total, then, the silent revolution in groundwater irrigation is essentially a story of the past 50 years (Llamas and Custodio, 2003). These can be considered as global pockets of intensive groundwater irrigation areas (Shah et al., 2007). Now there are pockets of intensive groundwater use, usually in food-producing areas of the world, such as the North China Plains, western and southern India, and parts of Mexico and the Ogallala aquifer of the USA.
4.21.1.1.5 Wetlands Wetlands act as sources of water for the majority of the global population. Agriculture’s impacts and dependencies upon wetlands are becoming increasingly significant. Wetlands are the key areas for managing extreme water flows after heavy rainfall and for providing water during droughts. Two recent global estimates have reported on the distribution of wetlands. Compiling national inventories, Finlayson et al. (1999) estimate global wetland area at 1280 million hectares. A more recent study by Lehner and Do¨ll (2004) used multiple geospatial data sets to estimate global wetland area at 917 million hectares. Accurate information on the distribution and extent of wetland ecosystems, both regionally and globally, is clearly an area requiring further work. Nevertheless, taking these data as the best-available estimates, a minimum of 131 million hectares of wetlands occur in Africa and 286 million hectares in Asia. The millennium ecosystem assessment (MEA, 2005a) identified agriculture as the major cause of wetland degradation and loss because it is the major economic activity in and around many wetlands, where crops such as rice, maize, and various vegetables and fruits are cultivated (Dries, 1989; Soerjani, 1992; Omari, 1993). However, agricultural development has considerably decreased the ecosystem services of wetlands (FAO, 2008). More recently, the comprehensive assessment of water management in agriculture (Falkenmark et al., 2007) concluded that pressures on wetlands would probably increase, with the prospect of serious loss of wetlands and ecosystem degradation. The needs of agriculture for flat, fertile land with a ready supply of water frequently make wetlands a valuable agricultural resource. In many arid and semi-arid regions of seasonal rainfall, where even major rivers can run dry for parts of the year, wetlands function to retain moisture. For this reason, they also make attractive resources for agriculture. Where people have to cope with both seasonal and interannual shortages of water, wetlands continue to be a vital resource for cultivators and pastoralists. In recent decades, agricultural use of wetlands has increased significantly in many developing countries, particularly in Africa, where they are perceived by some as the new frontier for agriculture (Wood and Dixon, 2009). This increase is driven partly by population growth, partly by the degradation of overexploited upland fields, and partly by market opportunities and the need to earn cash income (Wood and van Halsema, 2008). For poor rural households short of food, wetlands can offer good soils as well as water for irrigation, fisheries, and edible plants. In this way, wetlands can provide a safety net for poor households.
711
Some rural households increasingly use wetlands to supply local markets with irrigated vegetables and other products, which generate income. Seasonal wetlands also provide an important resource for livestock grazing. Sometimes these act as grazing land, but in some cases, they are used for hay production. This is prominent in many of African savannahs where the climate is semi-arid, rainfall is seasonal, and wetland grazing is widespread (FAO, 2008). For these households, wetlands represent a development opportunity that can lead them out of poverty.
4.21.1.1.6 Water consumption In an agricultural context, water consumption refers to water rendered immediately unusable by way of evaporation and transpiration from crops, soil, and open water bodies. Of total agricultural water consumption, the sources are estimated at 78% green water and 22% blue water. However, blue water withdrawal rates are greater than blue water consumption rates because not all water used for irrigation evaporates. The 22% of blue water consumed equals 1570 km3, whereas 2630 km3 are withdrawn for irrigation annually. In total, 60% of the water withdrawn for agriculture is consumed, while 40% returns to surface water or groundwater. The ratio of consumption to withdrawal is commonly referred to as the consumptive fraction or depleted fraction (Molden, 1996). Consumptive fractions tend to be low in water-abundant areas (where intensive water management is not cost effective), whereas they tend to be higher in water-scarce areas (where plants use shallow groundwater and farmers reuse drainage water). In the arid Middle East and North Africa, for example, the consumptive fraction is 77% with peak values close to 100%. In water-abundant areas, the consumptive fraction can be as low as 35%. Generally, it is not feasible or desirable to have a consumptive fraction higher than 70% at the basin scale, due to substantial infrastructure and environmental costs (Molden et al., 2000). However, in all scenarios the demand for freshwater increases to meet future food demands. Water consumption increases substantially in irrigated and rainfed areas).
4.21.1.1.7 Water use Annual global water withdrawals are estimated at 3830 km3, 70% of which is used for agriculture (i.e., 2664 km3) (FAO, 2006a). The net evapotranspiration from irrigation is 1570 km3 yr1, while the majority of total evapotranspiration is directly from rainfall. About 1000 km3 or 25–30% of the 3830 km3 of total water withdrawals originate from groundwater, and is mostly used for drinking and irrigation purposes. In the past century, industrial and municipal water demands, including those for energy generation, have grown in relative proportion to agricultural water demands. As competition between these sectors intensifies, agriculture can expect to receive decreasing shares of developed freshwater resources. Again, geographical differences are important to note. Approximately 70% of the world’s irrigated land is in Asia. Of Asia’s total cultivated land, however, only 34% is irrigated. Furthermore, China and India alone account for more than half of the irrigated land in Asia. By contrast, however, there is very little irrigation in sub-Saharan Africa.
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Water Availability and Its Use in Agriculture
There has been relatively small investment in irrigation in Africa compared to massive investments in Asia, where irrigation fueled the green revolution. Demand for water for industrial and municipal uses, including for energy generation, is growing relative to demand for agriculture. As competition for water from these other sectors intensifies, agriculture can expect to receive a decreasing share of developed freshwater resources.
4.21.1.1.8 Climate change, agriculture, and water Predicted climate change scenarios show many potential risks to agriculture and agricultural water use:
• • • • •
increased precipitation intensity and variability could therefore the occurrence of flood and drought conditions as well as runoff patterns and the risk of crop failure; changing sea levels, water logging and after cause, could cause seawater intrusion and groundwater salinization in delta zones and other coastal areas; increasing temperatures will likely reduce crop productivity and increase water requirements in low latitude regions, thereby directly decreasing water-use efficiency; irrigation demands will likely increase; and water scarcity will increase in areas with growing populations and decreased precipitation.
At the watershed scale, changes in evaporation, precipitation, and water-storage cycles will alter the seasonal, annual, and interannual water availability for both terrestrial and aquatic agro-ecosystems (FAO, 2003). Several studies have also linked increased temperatures and evaporation, and decreased rainfall with greater needs for irrigation (Barnett et al., 2005; Bates et al., 2008; IPCC, 2001). Therefore, under these conditions, issues surrounding water demand and availability will increasingly affect agricultural activities, food security, forestry, and fisheries (Bates et al., 2008). In addition to these longterm climate issues, the severity of specific climate events will also influence agriculture around the globe. For example, more than 90% of simulations predict increased droughts in the subtropics by the end of the twenty-first century, while increased extremes in precipitation are projected in the major agricultural production areas of southern and eastern Asia, eastern Australia, and northern Europe (Bates et al., 2008).
4.21.1.1.9 Drivers of water use Population growth and changing diets are the two prominent drivers of increased food demand and, as it follows, increased water use (the following discussion is after Fraiture et al., 2007). From 6.1 billion people living on the planet in 2000, global population is projected to grow to 7.2 billion in 2015, 8.1 billion in 2030, and 8.9 billion in 2050 (UN, 2003). This growth curve is projected to level off after mid-century, except in sub-Saharan Africa where populations are projected to continue to grow. Furthermore, in regions where incomes increase, diets often change. In these scenarios, while the production of staple cereals goes up, greater numbers of people will also shift away from eating cereals as their primary food source and begin consuming greater quantities of livestock products, such as fish, and high-value crops. The world food supply increased from about 2400 kcal per person per day in
1970 to 2800 kcal per person per day in 2000, a 16.6% increase. However, geographical differences must be used in context here. In developed countries during the same period, food supply increased from 3050 to 3450 kcal per person per day, while in sub-Saharan Africa supply only increased from 2100 kcal per person per day to about 2200 kcal per person per day. The growth in per capita food consumption has been accompanied by significant changes in the commodities people choose to consume. Meat consumption has increased in all regions except sub-Saharan Africa, and industrial countries are by far the largest meat consumers, at 103 kg per person per year, a trend that is projected to continue for the next 50 years. The same patterns apply to dairy products as well. In total, wealthier populations consume more food per person and eat richer, more varied diets, while producing these foods means using more water. Increased urbanization and urban migration also drive food production and agricultural water demands. In the 1960s, two-thirds of the world’s population lived in rural areas, and 60% of the economically active population worked in agriculture. Today these ratios have changed. About half of the people alive today live in rural areas. Furthermore, a little more than 40% of the economically active population depends directly on agriculture as a means of well-being. In absolute terms, rural populations will begin declining in the next few years, and by 2050, two-thirds of the world’s people will live in cities and mega-cities. But, again, global averages do not express significant regional variations. In many poor countries in South Asia and sub-Saharan Africa, the rural population will continue to grow until about 2030, while the number of people depending on agriculture in these places will continue to rise (CA, 2007). Rapid rural-to-urban migration in developing countries also influences farming practices and water demand. In this process, more men are migrating to urban centers leaving women, older people, and children behind in rural areas. Consequently, in developing countries, women’s presence in agricultural economies is growing, rising from 39% in 1961 to 44% in 2004, whereas in developed countries these numbers are falling, dropping from 44% to 35% over the same period (FAO, 2006b). As this happens, issues of gender will be increasingly important to water management, productivity, equity, and governance. As cities expand in size, their demands on and claims to water resources increase, often at a loss to rural agricultural areas. In many cities today, poor or nonexistent urban planning and enforcement of land-use regulations compound water management problems. Urbanization also physically affects hydrological environments. Buildings, roads, and parking lots, among other human structures, create impermeable surfaces while sewage systems redirect large volumes of water. As a result, surface areas available for water infiltration are decreased, and increased runoff can become a significant problem. In river basins affected by urban footprints, peak discharge occurs quicker and reaches higher volumes. This can result in greater stream channel erosion, possible channel destruction, and habitat degradation, while it can also damage human life and property. Furthermore, these surfaces reduce groundwater recharge and can decrease long-term groundwater inflow to streams. Urban centers are cites of water
Water Availability and Its Use in Agriculture
pollution. The increased presence of sediments, nutrients, microbes, toxic metals, and organics is a major externality of urbanization. All of these factors can have significant effects on human health, downstream environments, and agricultural systems. Hydroelectric power generation is another significant driver influencing water availability and agricultural production. Dams worldwide produce 715 000 MW or 19% of the world’s electricity. The process of hydroelectric power generation requires water storage and stream flow regulation, both of which can influence water availability for agriculture and other users. Thus, significant volume of the world’s blue water resources is held in river basins where multipurpose water management is linked to energy production. Many factors relate to the nature of hydroelectric production condition the way these water resources are managed. Dams have the ability to store and release water at specific times and this means electricity can be generated on demand. In the same way, dams also have the ability to regulate water for irrigation, navigation, and recreation. As dams are not sources of CO2 emissions, provided vegetation is cleared before following up hydroelectric power may be an attractive energy option in the future. However, other issues, including habitat degradation and cost efficiency, are at stake and need to be considered in analyzing trade-offs. At a different scale, smallscale hydro or micro-hydro power has been increasingly popular, especially in remote areas where other power sources are not feasible. Small-scale hydropower systems are installed in small rivers or streams with little or no marked environmental effects. In poor areas, many remote communities have no electricity. Micro-hydro power, with a capacity of 100 kW or less, allows communities to generate electricity. Changing consumption patterns, more people moving into urban areas, and increasing demand for low-carbon energy mean that agricultural water use will see greater outside competitive pressures. Small holders and individual water users, with little political power in water governance processes, will face greater risks in these conditions. The following section examines concepts of water scarcity as a means of understanding how these risks affect people differently. Water scarcity is often viewed as a physical issue, where aggregated demand for water by all potential uses is lays than the available supply (FAO, in preparation). Such approach, however does not capture the knowledge.
4.21.1.2 Physical and Economic Water Scarcity Another tool for examining water availability is the concept of water scarcity (Seckler et al., 2000; Molden et al., 2007). Rather than analyzing water availability from a hydrological approach (using river basins as units of analysis), water scarcity focuses on social and political regions (using populations as units of analysis). Evaluating water scarcity begins at a micro-level, one can the water security of individuals. Individuals are water secure when they have consistent access to safe and affordable water to satisfy their needs for drinking, washing, food productions, and other livelihood endeavors; they are water insecure when these needs cannot be met. A region is water scarce when a large number of people
713
are water insecure (Rijsberman, 2006). In adapting such approaches, economic, financial, social political and institutional dimensions of the problem of access to water become as relevant as the physical availability of water. These multiple dimensions of the problem have been captured in the dual concept of physical and economic water scarcity. Access to water is difficult for millions of people for social, political, and economic reasons, in addition to physical resource constraints. About 2.8 billion people live in areas facing water scarcity, and more than 1.2 billion of them – onefifth of the world’s population – live in areas of physical water scarcity (Molden et al., 2007 – trends chapter). Another 1.6 billion people live in basins that face economic water scarcity, where human and institutional capacity or financial resources are likely to be insufficient to develop adequate water resources even though adequate water is available to meet human needs (Figure 3). Within these regions, poor people suffer disproportionately from the implications of scarcity. Lack of finance, lack of human capacity, poor management, and a lack of good governance all contribute to water scarcity. Physical water scarcity occurs when available water resources are insufficient to meet all demands, including minimum environmental flow requirements (Figure 3). Arid regions are most often associated with physical water scarcity; however, an alarming new trend of artificial physical water scarcity is affecting even regions where water is abundant. This problem is due to the over-allocation and over-development of water resources, leaving no scope for making water available to meet new demands except through interbasin transfers. In these scenarios, there is not enough water to meet both human demands and environmental flow needs. The implications of physical water scarcity include severe environmental degradation, such as river desiccation and pollution, declining groundwater tables, water allocation disputes, and failure to meet the needs of individuals and groups. Some 1.2 billion people live in river basins where human water use has surpassed sustainable limits. Meanwhile, another 500 million people live in river basins that are fast approaching physical water scarcity. While physical scarcity introduces complex problems, investments in good management can mitigate many of the issues. Economic water scarcity occurs when the investments needed to keep up with growing water demand are constrained by limited financial, human, or institutional capacities. Much of the scarcity felt by people is due to problems with institutions and politics, favoring one group over another, not listening to the voices of women and disadvantaged groups, for instance. Problems of economic water scarcity include: inadequate infrastructure development, where people have trouble getting enough water for agriculture and domestic purposes; high vulnerability to seasonal water fluctuations, including floods and long- and short-term droughts; and inequitable distribution of water even though infrastructure exists. Much of sub-Saharan Africa experiences economic water scarcity, and there are many areas across the globe where water resources are inequitably distributed. Further water development could ease problems of poverty and inequality.
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Water Availability and Its Use in Agriculture
Little or no water scarcity
Approaching physical water scarcity
Physical water scarcity
Economic water scarcity
Not estimated
Figure 3 Areas of physical and economic water scarcity. Definitions and indicators: (1) Little or no water scarcity. Abundant water resources relative to use, with less than 25% of water from rivers withdrawn for human purposes. (2) Physical water scarcity (water resources development is approaching or has exceeded sustainable limits). More than 75% of river flows are withdrawn for agriculture, industry, and domestic purposes (accounting for recycling return flows). This definition – relating water availability to water demand – implies that dry areas are not necessarily water scarce. (3) Approaching physical water scarcity. More than 60% of river flows are withdrawn. These basins will experience physical water scarcity in the near future. (4) Economic water scarcity (human, institutional, and financial capital limit access to water even though water in nature is available locally to meet human demands). Water resources are abundant relative to water use, with less than 25% of water from rivers withdrawn for human purposes, but malnutrition exists. From International Water Management Institute analysis done for the comprehensive assessment for water management in agriculture using the Watersim model. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
Table 2
Withdrawals by nonagricultural sector will increase by a factor of 2.2 by 2050
Region
Sub-Saharan Africa East Asia South Asia Central Asia and Eastern Europe Latin America Middle East and North America OECD countries World
Agriculture
Domestic
Manufacturing
Thermo cooling
Total nonagricultural
2000
2000 2050
2000
2050
2000
2050
2000
2050
68 518 1095 244 175 173 233 2630
7 48 15 40 31 14 121 278
2 21 4 68 12 3 135 245
8 159 29 236 42 10 131 617
1 32 15 48 10 7 262 376
18 75 55 52 134 22 307 664
10 101 34 156 53 24 518 902
60 419 175 377 254 82 590 1963
35 185 90 88 78 51 152 681
Annual increase (%) 2000–50
3.7 2.9 3.3 1.8 3.2 2.5 0.3 1.6
Note: Units are in km3 unless otherwise indicated. From Comprehensive Assessment for Water Management in Agriculture (2007). Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
4.21.1.3 Future Demands for Water If improvements in land and water productivity or major shifts in production patterns do not occur in the near future, crop water consumption would increase 70–90% by 2050 depending upon actual population growth, changing income levels, and water requirements for livestock and fisheries. In this scenario, crop water consumption would go from a current rate of 7130 km3 yr1 to somewhere in the range of 12 050–13 500 km3 yr1. This estimated range accounts for crop water depletion for food and feed production,
plus losses through evaporation from soil and open water sites. Nevertheless, even with improvements in water productivity, agriculture will continue to consume a large portion of the world’s developed water supply. This topic is discussed further in the following section. Industrial and domestic demand for water will continue to increase with urbanization. Withdrawals for nonagricultural sectors are expected to more than double by 2050, and, as it follows, there will be increasing competition for water between sectors (Table 2). In most countries, urban water demands receive customary or legal priority over water for
Water Availability and Its Use in Agriculture
715
10
8 Projection (high)
mt ha–1
6
OECD countries FAO
IWMI Projection (low)
4 World IWMI
2 FAO
0
sub-Saharan Africa
1961
1970
1980
1990
2000
2010
2020
2030
2040
2050
Figure 4 Global water withdrawals increase. Points marked FAO (Food and Agriculture Organization) are based on projections in Bruinsma (2003); those marked IWMI (International Water Management Institute) are based on projections in Seckler and others (2000). From FAOSTAT (2006), for 1960–2003; International Water Management Institute analysis done for the comprehensive assessment for water management in agriculture using the Watersim model, for 2000–50. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
agriculture (Molle and Berkoff, 2006). Greater competition for water will leave less for agriculture, particularly near large cities in water-short areas. The regions of the Middle East, North Africa, Central Asia, India, Pakistan, Mexico, and northern China, among other areas, will see greater competition for water as urban centers continue to develop there. Estimates also show that while the proportion of water diverted for nonagricultural sectors increases, agriculture remains the largest water user among the productive sectors. Although major trade-offs will occur between all water-using sectors, they will be particularly pronounced between agriculture and the environment as the two largest water-demanding sectors (Figure 4) (Rijsberman and Molden, 2001). Unlike agricultural water consumption, only a small part of the water diverted for domestic and industrial purposes is consumed. In urban areas, 75–85% of water diverted flows back to rivers, lakes, and groundwater as return flow. In many urban areas, particularly in water-scarce developing countries, wastewater is used for high-value vegetable production, a livelihood activity for millions of city dwellers (Gupta and Gangopadhyay, 2006; Hussain et al., 2001, 2002; Raschid-Sally et al., 2005). The use of urban wastewater for irrigation will increase as water becomes scarcer in urbanizing areas. If by 2050 half of return flows from cities are reused, 200 km3 of wastewater could be used for irrigation. This would represent only 6–8% of future agricultural withdrawals, but the economic values generated could be substantial. Much of the wastewater would likely be used to produce highly valued vegetables, helping sustain the livelihoods of millions of small farmers (Hussain et al., 2001, 2002). As reuse of city wastewater for agriculture poses environmental and health risks, these can be minimized with proper management. Water demand for managing ecological functions has also created greater resource competition, as reflected in changing
policies for water allocation and pricing. In many countries, rising incomes are correlated with increasing demands for restoring and maintaining environmental services. The demand for environmental amenities adds pressure on scarce water resources. A first-cut estimate by Smakhtin et al. (2004) indicates that 20–45% of long-term annual flows must be preserved to maintain essential ecosystem services. UNESCO (2006) suggests that 100 km3 need to be added to estimates of future water demands to account for current overexploitation of groundwater and 30 km3 must be added to account for the mining of nonrenewable groundwater, or fossil groundwater.
4.21.1.4 Future Scenarios for Rainfed and Irrigated Agriculture As outlined above, there will be greater demands on agriculture and water by the year 2050. Considering agricultural productivity, in the context of different irrigation scenarios, is an effective way to understand how these demands can be met. The comprehensive assessment of water management in agriculture (CA, 2007) provided scenarios to allow us to explore various futures. Most importantly, if irrigation development were to remain static from now until 2050, the agricultural potential of rainfed agriculture would be sufficient for meeting the projected additional food requirements in 2050 such as cultural values. In an optimistic yield growth scenario, in which cereal yields grow by 72%, the demand for agricultural commodities is met by increasing rainfed-harvested area by 7% (this work follow Fraiture et al., 2007). The contribution of rainfed agriculture to the total gross value of food supply would increase from 52% in 2000 to 60% in 2050 (CA, 2007). In the optimistic yield scenario, sub-Saharan Africa, Asia, and Latin America can be largely self-sufficient in producing major food
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Water Availability and Its Use in Agriculture
crops. East Asia, however, would need to import maize to meet the large increase in feed demand. In addition, the Middle East and North Africa would need to import food due to lack of suitable lands for rainfed production. Global food trade would increase from 14% to 17% of total production. The scenario analysis also demonstrates the risks inherent in a rainfed-based strategy. In the pessimistic scenario, with a low rate of adoption of water harvesting and only modest improvements in rainfed yields, the area of rainfed production must increase by 53% to meet future food demands (an additional 400 million hectares as compared with the optimistic yield scenario; see Figure 5). The Food and Agriculture Organization of the United Nations (FAO) estimates suggest ample capacity for increasing the area under cultivation, except in South Asia, the Middle East, and North Africa. In sub-Saharan Africa and Latin America, only one-fifth of the potential land area is already in use. Although there are significant amounts of land available for cultivation, more than half are now forested or protected areas (Alexandratos, 2005). Furthermore, some of these lands might be of marginal quality (Bruinsma, 2003) or not suitable for cereal crops. In the pessimistic yield scenario, countries without potential to expand rainfed areas – due to either
lack of suitable land or unreliable rainfall – must increase food importation. In this case, the Middle East and North Africa would import more than two-thirds of their agricultural needs. South and East Asia, due to land limitations, would become major importers of maize and other grains, importing 30–50% of their domestic needs. Latin America, developed countries, Central Asia, and Eastern Europe, having the potential to expand land in agriculture, would increase their exports. Globally, food trade would increase from 14% of total agricultural production to 22% in 2050. Large grain imports by East and South Asia would put upward pressure on food prices (the model results suggest an increase of 11%). There is a risk that poor countries may not be able to afford food imports, and household-level food insecurity and inequity might worsen. Climate change, which is expected to increase the variability and intensity of weather events, exacerbates the risks of rainfed production, particularly in semi-arid areas vulnerable to drought (Kurukulasuriya et al., 2006). Both the optimistic and pessimistic rainfed scenarios lead to substantial increases in soil water consumption. While the global average of rainfed cereal yield would improve by 72%, crop water productivity would improve by 35%. In the pessimistic yield scenario,
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Water Availability and Its Use in Agriculture
global rainfed cereal yields improve by 20% and water productivity by 10%, while total crop water consumption increases by 54% to 10 980 km3, an additional 3850 km3 after the year 2000. Increases agricultural evapotraspiration of that order of magnitude will have impacts on river flows and groundwater recharge, with implications for downstream water users and those relying on groundwater resources.
4.21.2 Productive Use of Agricultural Water The long-term sustainability of food production and agriculture depends on the efficient management of limited water resources. Moreover, as these resources come into greater demand, driven by a broadening range of applications and functions, the implications of agricultural water use will have increasing effect on other water users and the environment. In other words, the impact of agricultural water use is increasing. By analyzing agricultural water productivity, crop production can be assessed in terms of its social, economic, and ecological costs and benefits. The following section outlines several approaches for assessing water productivity and contextualizes the topic in relation to rainfed and irrigated agricultural systems. From here, future productivity scenarios are considered, and livestock and fisheries agriculture are discussed (this section draws from analysis presented in de Fraiture et al., 2007).
4.21.2.1 Water Productivity in Agriculture Water productivity is defined as the ratio of the net benefits from crop, forestry, fishery, livestock, and mixed agricultural systems to the amount of water required to produce these benefits (this section draws from Molden et al., 2007). In its broadest sense, water productivity reflects the objectives of producing more food, income, livelihoods, and ecological benefits at less social and environmental cost per unit of water used. Water productivity can be defined in several ways. Physical water productivity is the ratio of the mass of agricultural output to the amount of water used, while economic
water productivity is defined as the value derived per unit of water used (Figure 6). Other modes of analysis include crop water productivity, where specific crops are measured individually for comparative purposes, and livestock water productivity where the ratio of the net beneficial livestock-related products and services is calculated in relation to the volume of water depleted in production including the water to feed them (Peden et al., 2007). In areas of the world already exhibiting high physical water productivity, the scale for improvement is limited. Many rainfed, irrigated, livestock, and fisheries systems across the globe, however, do not exhibit high physical water productivity. Many farmers in developing countries could raise their water productivity by adopting better management practices. These include supplemental irrigation; soil fertility maintenance; deficit irrigation; small-scale water storage, delivery, and application; modern irrigation technologies (such as pressurized systems and drip irrigation); and soil water conservation through mulching zero or minimum tillage. Breeding technologies and biotechnology can also indirectly help agricultural systems become more efficient by reducing biomass losses through increased resistance to pests and diseases, enhancing the vigorous early growth for fast ground cover to reduce soil surface evaporation, and by reducing drought susceptibility for specific crops. However, water productivity gains are context dependent, and, in some cases, a gain for one group of people can mean a loss for the others. Water productivity, especially in physically water-scarce basins, can be properly assessed only by taking an integrated basin perspective where trade-offs between uses are considered. Employment opportunities, income generation, nutrition, and opportunities for women can all be linked to agricultural productivity; in this way, increasing values derived per unit of water is important to poverty reduction. However, carefully crafted programs are required to ensure that these gains reach the poor, especially rural women, and are not captured exclusively by wealthier or more powerful users. As described above, rising demand for livestock and fish products also leads to rising demand for water. In producing
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Figure 6 Agricultural water management: a continuum of practices. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
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these animal products, water productivity gains can be made by carefully considering feed sources and feeding strategies animal species produced (chicken use less water than cattle), improving the quality of produce, and integrating fisheries and livestock into farm production systems. Because freshwater fisheries are increasingly threatened by reductions in stream flows, basin water productivity analysis should consider the social and ecological values generated by fisheries before reducing river flows that support them. Several studies describe multiple uses of agricultural water and the ways in which these uses can improve productivity. Poor rural households use agricultural water in multiple ways (Laamrani et al., 2000; Moriarty et al., 2004; Jehangir et al., 2000). Agricultural water can be used for drinking, sanitation, home gardens, livestock, rural industries, and aquaculture. The increase of agricultural water productivity can result in many benefits such as:
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helping to meet food demand; contributing to poverty reduction and economic growth; and helping to reduce pressures to reallocate water from agriculture and to ensure that water is available for environmental uses.
Integrated and multiple-use systems – in which water serves crops, fish, livestock, and domestic purposes – can increase the value derived per unit of water used. Gains in crop production have often come, for instance, at the expense of fisheries. Values generated by fisheries, including ecosystem sustenance values, are routinely underestimated. Recognizing these values helps us to understand where there are win–win situations and what trade-offs will have to be made. However, these values are poorly recognized today and rarely influence the decision-making processes.
4.21.2.2 Rainfed Agriculture Productivity Rainfed agriculture includes both permanent crops (such as rubber, tea, and coffee) as well as annual crops (such as wheat, maize, and rice). For example, tubers, a staple crop for subSaharan Africa, have been all but uninfluenced by the technological developments of the green revolution. Rainfed farming constitutes 80% of the world’s cropland and produces more than 60% of the world’s cereal grains, generating livelihoods in rural areas while producing food for cities. In temperate regions with relatively reliable rainfall and good soils, rainfed agriculture generates high yields. Supplemental irrigation practices boost yields even higher. With rising concerns over the high cost of expanding largescale irrigation and the environmental impacts of large dams, upgrading rainfed agriculture is gaining increased attention (Rockstro¨m et al., 2007). Many people dependent on rainfed agriculture are highly vulnerable to both short-term dry spells and long-term droughts. Exposure to these risks can contribute to a reluctance to invest in agricultural inputs that could increase crop yields. Moreover, changing precipitation patterns resulting from climate change will compound this issue for many small farmers. There are several compelling reasons to invest in agricultural water management technologies and institutions
connected to rainfed agriculture (Rockstro¨m et al., 2007). To start, there is high potential to improve productivity, especially where yields are low. A majority of the rural poor are small holders who depend on rainfed rather than irrigated agriculture. Improving productivity in rainfed areas is therefore a way of supporting the poor. Boosting the potential of existing rainfed areas reduces the need for new large-scale irrigation development, which can generate adverse environmental and social impacts. Furthermore, the cost of upgrading rainfed areas is generally lower than the cost of constructing new irrigation schemes, particularly in sub-Saharan Africa. Even with these incentives, the potential contributions of rainfed agriculture to world food production are debatable, and forecasts regarding the relative roles of irrigated and rainfed agriculture vary considerably. Relying on rainfed agriculture also involves considerable risk. Water-harvesting techniques are useful for bridging short-term dry spells. Investments in water management are thus a way to decrease risk in rainfed agriculture. However, adoption rates of waterharvesting techniques are low, and extending successful local techniques over larger areas has proved to be difficult in the past. As longer dry spells may lead to crop failure, rainfed agriculture generally entails more risk than fully irrigated agriculture. Farmers adopt risk management strategies in line with the level of risk. There is a range of ’soft’ and hard’ measures that are available to integrate the risk related to climate variability in agriculture. Better control of water, either through full-fledged irrigation or supplemented irrigation, in costs or systems of crop insurances, can also integrate risk and provide farmers with better incentives to invest in their crop (Faures, 2010).
4.21.2.3 Irrigated Agriculture and Productivity The last 50 years have seen major investments in large-scale public surface irrigation as part of a global effort to increase staple food production, ensure food self-sufficiency, and to avoid devastating famine (this section follows Faures et al., 2007). From 1961 to 2008, for example, the world’s cultivated land increased approximately 12% (i.e., from 1368 to 1526 million hectares). At the same time, irrigated land area increased by 120%. The percentage of cultivated land equipped with irrigation rose from 10% in 1961 to 20% in 2008 (i.e., from 139 to 306 million hectares). Paralleling these global trends, irrigation investments in developing countries also accelerated rapidly in the 1960s and 1970s. On average, irrigated land in these regions grew by 2.2% per year reaching 155 million hectares by 1982. Widespread use of newly developed, high yielding, and fertilizer-responsive crops partially constituted the increased demand for water during this period. To achieve the higher yields now possible from these new crops, agriculture simply needed more water. In the developing world, other factors were important to the increased use of irrigation. Private and community-based investments in these countries, particularly programs aimed at groundwater pumping, grew from the 1980s onward. These projects were propelled by cheap drilling technology, rural electrification, and the availability of inexpensive small water pumps. Approximately 70% of the world’s irrigated land is in Asia (Figure 7), where it accounts for 34% of cultivated land.
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Figure 7 The area equipped for irrigation. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007).
China and India alone account for more than half of irrigated land in Asia. Over time, Asia, with its high population densities, has come to rely increasingly on irrigated agriculture to boost agricultural productivity and thus to ensure domestic food security. Sub-Saharan Africa is much different. Notwithstanding a few large commercial irrigation schemes developed during the colonial period and a relatively modest small-scale irrigation subsector, there is very little irrigation in sub-Saharan Africa where water application methods are largely surface irrigation based, and little has been done to improve water productivity. The 1990s, however, saw a substantial rise in private irrigated peri-urban agriculture in sub-Saharan Africa in response to higher demand from growing cities for fresh fruits and vegetables (FAO, 2005). Today, it is suggested that global harvested irrigated area, which includes double cropping (two crops are grown in the same year), is estimated at 340 million hectares, although new incomplete evidence suggests otherwise. According to some studies, global harvested irrigated area might actually be higher than previously calculated after adjusting for higher cropping intensity and unreported, often informal, groundwater, or private irrigation use (Thenkabail et al., 2006). By the mid-1990s, irrigation projects leveled off around the world. Before this, the rapid growth in irrigated area, along with the other technological advancements of the green revolution – such as improved crop varieties and substantial growth in fertilizer use – led to a steady increase in staple food production and a reduction of real-world food prices. Until very recently, food prices in developed countries have been kept low by agricultural subsidies (Rosegrant et al., 2002), and since the late 1970s, the annual growth rate of global irrigation development, particularly in large-scale public schemes, has decreased. Other factors also contributed to the post-green revolution slow down of irrigation development. Most areas best suited for dams and irrigation have been developed, and as a result, new dams for irrigation and the related infrastructure for water delivery will cost more to construct in less
ideal locations. As a result, these geographic and economic conditions have led to overall less economic incentives for the development of large-scale irrigation projects. Other recent factors have created disincentives for irrigation investments as well. Some research has shown that the underperformance of large-scale irrigation (Chambers, 1988) has reduced donor interest (Merrey, 1997). Attention to the negative social and environmental externalities of dams – particularly the displacement of residents in affected communities and the calls for increased in-stream flows for environmental purposes – has discouraged the lending markets for irrigation investment. More competition for water from other sectors (as mentioned above) has reduced the scope for further development of irrigation. Irrigation is particularly crucial in sustaining agriculture across the dry belt, a region that extends from North Africa, the Middle East, through Northern China to Central America and parts of the United States (Figure 8). The advent of affordable drilling and pumping technologies in India and Pakistan in the mid-1980s led to the rapid development of shallow tube wells and the combined or conjunctive use of surface water and groundwater (Shah, 1993; Palmer Jones and Mandal, 1987). These technologies enabled farmers to have direct, individual control over water resources. By harvesting water by way of groundwater pumping, drainage reuse, or direct pumping from ponds, canals, and rivers, small holders gained flexibility and reliability in water delivery. Large-scale surface distribution systems did not offer these advantages. Yet, these technologies also contributed new challenges to water management. The indirect subsidization of electricity enabled farmers to pump water at zero to little cost. As a result, water tables have fallen in many regions of the world. In 1995, 38% of cereals grown in developing countries were on irrigated land, accounting for just less than 60% of all cereal production (Ringler et al., 2003). Rainfed cereal yields averaged 1.5 Mt ha1 in the developing world in 1995, but irrigated yields were 3.3 Mt ha1 (Rosegrant et al., 2002). The
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Figure 8 Irrigated areas as a share of cultivated area by country, 2003. Source: Comprehensive assessment of water management in agriculture (2007), Water for Food, Water for Life (Earthscan, 2007). From FAO (2006a) AQUASTAT database. Rome.
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difference in productivity between irrigated and rainfed agriculture varies widely, depending on the climate, combination of crops, and technologies used. Typically, land productivity is 2–4 times higher in irrigated agriculture. Moreover, cropping intensity is typically higher under irrigation, with up to three rice crops per year in parts of Southeast Asia and two crops per year in most of the Asian subcontinent. Figure 9 shows the distribution of crops under irrigation worldwide.
4.21.2.4 Livestock Keeping livestock is one of the most important, complex, and diverse subsectors of world agriculture and for many people it
is a primary means of escaping poverty in rural areas. Modest amounts of meat in the diets of African children appear to improve mental, physical, and behavioral development (Sigman et al., 2005; Neumann et al., 2003). This suggests that meat production and water productivity must account for social and health values as well as produced food mass. However, current literature on livestock–water interactions does not address this important topic. Moreover, past research has underestimated the contributions of livestock to rural livelihoods in part because studies were predominantly concerned with food mass productivity. Limited consideration has been given to the nonmonetized products and services associated with livestock.
Water Availability and Its Use in Agriculture
Poor and subsistence households obtain multiple benefits from the use of livestock (Shackleton et al., 1999; Landefeld and Bettinger, 2005). Therefore, assessing the water resources used to support these animals must account for values beyond meat production. Livestock contribute to the livelihoods of at least 70% of the world’s rural poor and strengthen their capacity to cope with income shocks (Ashley et al., 1999). They provide milk, blood, manure, hides, and farm power essential for the cultivation and marketing of crops. Livestock assets are often an important source of wealth security. As mentioned above, livestock water productivity examines the net beneficial livestock-related products and services in relation to water use. As a systems concept, livestock water productivity attempts to account for the complex relationships among food production, livelihood, and water demands. The implications of livestock on water use have been generally overlooked by research. Animals ‘consume’ water directly for drinking purposes, but it is the food they eat that requires large quantities of water, as discussed earlier. The type, quality and origin of the feed used for animals, together with animal management practices can have major impact on livestock water productivity. Livestock water productivity differs from water or rainuse efficiency because it examines water depleted rather than applied or inflowing water. Four basic strategies help to increase livestock water productivity: improving water supply, feed sourcing, enhancing animal productivity water conservation, and spatially optimistic distributing of watering points, animal stocking rates, and pasture productivity (CA, 2007). Providing sufficient and adequate quality drinking water also improves livestock water productivity as it keeps the animal healthy and productive. However, it does not factor directly into the livestock water productivity equation because water that has been consumed remains inside the animal and thus within the production system, although subsequent evaporative depletion may follow. A balanced, site-specific approach that considers all four strategies will help increase the benefits derived from the use of agricultural water for the production of animal products and services. Children, women, and men often receive different benefits from animal keeping and have different roles in managing livestock–water interactions These are considerations that need to be taken into account in attempts to improve livestock water productivity. Livestock water productivity does not necessarily seek to maximize the number of livestock or the production of animal products and services. Rather it seeks to reach a higher level of animal products per unit of water consumed.
4.21.2.5 Aquaculture and Fisheries Inland fisheries and aquaculture contribute about 25% to the world’s fish production and are a fast growing sector (see Dugan et al., 2007). In addition, many important estuarine and coastal fisheries are closely linked to the ecological processes that occur in freshwater systems. Fisheries and aquaculture from lakes, reservoirs, rivers, ponds, and wetlands contributed about 25% (i.e., 34 million metric tons) of the reported world fisheries production in 2003 (FAO, 2004). However, catches in rivers and wetlands are easy to
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underestimate because the contributions of numerous fisheries on smaller tributaries and water bodies are generally overlooked (Coates, 2002). Reported harvests from river fisheries alone have been shown to account for only 30–50% of actual catch (Kolding and van Zwieten, forthcoming), and the contribution from inland fisheries is therefore believed to be underestimated. Aquaculture uses water in two ways. Blue water is needed for the fish ponds and the processes of aquaculture; blue and green water is also necessary for feed production. Water productivity in terms of aquaculture is defined as the mass or value of the aquaculture produce divided by the amount of water required for feed plus the amount of evaporation from the pond. On-farm water use in aquaculture can be as low as 500–700 l in super-intensive recirculation systems and as high as 45 000 l of water (evaporation plus seepage plus feed) per kg of produce in extensive ponds (Verdegem et al., 2006). Fish can often be integrated into water management systems with the addition of little or no water (Prein, 2002). Renwick (2001) found that the fisheries in irrigation reservoirs at Kirindi Oya, Sri Lanka, contributed income equal to 18% of the rice production in the system. Haylor (1994, 1997) assessed the potential for aquaculture in small- and large-scale irrigated farming systems in the Punjab, Pakistan. The study noted that aquaculture in the region was almost entirely focused on carp production using groundwater sourced from tube wells. It also concluded that there was economic justification for expanding such aquaculture using local shallow tube wells. The study also found that the revenue potential for cage aquaculture in irrigation canals was also attractive, but operational conflicts in the use of water for agriculture would need to be resolved. Murray et al. (2002) have pointed out that traditional power structures may undermine attempts to integrate aquaculture in irrigation systems and that changes in laws and regulations would be required from community to national levels. In coastal areas, aquaculture may severely degrade land and water quality and biodiversity, requiring special attention (Gowing et al., 2006). Fisheries in lakes, rivers, and wetlands present a special case for water productivity assessment because fish are only one of the many ecosystem services provided by aquatic ecosystems. The values and livelihood benefits of fisheries are high and often ignored or underestimated, but considering only the values of fish produced would grossly underestimate the value of water in these aquatic ecosystems. The water productivity of fisheries systems needs to be considered in terms of ecosystem services and livelihoods supported per unit of water. Thus, maintenance of wetlands and biodiversity should be considered as potential benefits for leaving water in these aquatic ecosystems.
4.21.3 Environmental and Health Implications of Agricultural Water Use Agricultural water management has both negative and positive impacts on environment and health (this section follows Falkenmark et al., 2007). On the one hand, agricultural water management can improve health status through better
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nutrition, increasing the availability of drinking water, and controlling diseases such as malaria. On the other hand, extensive reporting shows many cases where agricultural water management has encouraged the spread of some waterborne diseases (Ersado, 2005); impacted upstream and downstream ecological services (Gichuki, 2004); affected water logging and salinization (Khan et al., 2006); and increased agrochemical usage, pollution, and eutrophication (Hendry et al., 2006).
4.21.3.1 Impact on Rivers, Wetlands, and Biodiversity Water management affects the physical and chemical characteristics of inland and coastal aquatic ecosystems, the quality and quantity of water, as well as direct and indirect biological change (Finlayson and D’Cruz, 2005; Agardy and Alder, 2005; Vo¨ro¨smarty et al., 2005). It has also affected terrestrial ecosystems through the expansion of agricultural lands and changes in water balances (Foley et al., 2005). Regulation of the world’s rivers has altered water regimes, with substantial declines in discharges to the ocean (Meybeck and Ragu, 1997). A long-term trend analysis (i.e., more than 25 years) of 145 major world’s rivers indicates that discharge has declined in one-fifth of the basins (Walling and Fang, 2003). Worldwide, large artificial impoundments hold vast quantities of water and cause significant distortion of flow regimes (Vo¨ro¨smarty et al., 2003).
4.21.3.1.1 Aquatic ecosystems Water diversion and the construction of hydraulic infrastructure have had the following negative effects: loss of local livelihood options, fragmentation, destruction of aquatic habitats, changes in the composition of aquatic communities, species loss, and health problems resulting from stagnant water. Improved flood control – an important agricultural mechanism for reducing risk – has led to the reduction of sedimentation and the deposition of nutrients on floodplains, as well as reduced flows and nutrient deposition to parts of coastal zones (Finlayson and D’Cruz, 2005). Inter basin transfers of water, particularly large transfers between major river systems have been in consideration in India and China, for example, are expected to be particularly harmful to downstream ecosystems (Gupta and Deshpande, 2004; Alam and Kabir, 2004) and will likely exacerbate pressures from hydrological regulation (Snaddon et al., 1999). Junk (2002) has highlighted the similar adverse consequences on water regimes expected from the construction of industrial waterways (i.e., hydrovias) through large wetlands, such as the Pantanal of Mato Grosso, Brazil. Shrinking lakes and rivers. There are many instances where consumptive water use and water diversions have contributed to the severe degradation of downstream ecosystem services by shrinking lakes and drying rivers. The degradation of the Aral Sea in Central Asia is an extreme case. Similarly, Lake Chapala, the world’s largest shallow lake, situated in the Lerma-Chapala Basin in central Mexico, is an example of consumptive water use upstream affecting lake-size downstream. From 1979 to 2001, water volume in the lake dropped substantially to about 20% of volume capacity due to excessive water extraction for agricultural and municipal needs.
Stream flow depletion is a widespread phenomenon in tropical and subtropical regions in river basins with large-scale irrigation, including the Pangani (IUCN, 2003), Yellow (He et al., 2005), Aral Sea tributaries, Chao Phraya, Ganges, Incomati, Indus, Murray-Darling, Nile, and Rio Grande (Falkenmark and Lannerstad, 2005). Smakhtin et al. (2004) have suggested that environmental flow (i.e., the stream flow required for aquatic ecosystem health) has already been overappropriated in many river basins. For example, in the United States the construction of dams and water diversions for irrigation and other purposes in the Colorado Basin, together with large-scale inter-basin transfers, have greatly reduced the flow of the river to the delta. As a result, a considerable portion of the delta has been transformed into mudflats, salt flats, and exposed sand. With the loss of the delta habitats, wetlands now exist mainly in areas where agricultural drainage has occurred (Postel, 1996). The Ganges is among the major rivers of South Asia that no longer discharges year round to the sea. As a result, there is a rapid upstream advance of saline water, with consequent changes in mangrove communities, fish habitat, cropping, and human livelihoods (Postel, 1996; Mirza, 1998; Rahman et al., 2000). On the Zambezi River in Southern Africa, damming for electricity and agriculture has reduced flows to the coast and led to a decline in shrimp production that could have been worth as much as $10 million a year (Gammelsrod, 1992). The regulation of rivers has brought many benefits to people, but the adverse impacts, especially those related to reduced downstream flows, have often failed to receive adequate and transparent consideration (WCD, 2000; Revenga et al., 2000; MEA, 2005a). Effects on wetlands. Water regulation and drainage for agricultural development are the main causes of wetland habitat loss and degradation (Revenga et al., 2000; Finlayson and D’Cruz, 2005) as well as the consequent loss of ecosystem services. By 1985, drainage and conversion of wetlands, mainly for agriculture, had affected an estimated 56–65% of inland and coastal marshes in Europe and North America and 27% in Asia (OECD, 1996). Drainage of wetlands often reduces important regulating ecosystem services, with such outcomes as increased vulnerability to storms and flooding and further eutrophication of lakes and coastal waters. The loss of small wetlands (regionally referred to as potholes) on the prairies of Canada and the United States through drainage and infilling has led to the loss of habitat for large numbers of migratory water birds (North American Waterfowl Management Plan, 2004). The loss of forested riparian wetlands adjacent to the Mississippi River in the United States was seen as an important factor contributing to the severity and damage of the 1993 flood in the Mississippi Basin (Daily et al., 1997). Changes in water quality. The use of fertilizers has brought major benefits to agriculture, and has also led to widespread contamination of surface water and groundwater through runoff. Over the past four decades, excessive nutrient loading has emerged as one of the most important direct drivers of ecosystem change in inland and coastal wetlands, with reactive nitrogen entering oceans at an increased rate of nearly 80% from 1860 to 1990 (MEA, 2005b). Phosphorus applications have also increased, rising threefold since 1960, with a steady increase until 1990 followed by a leveling off at approximately the application rates of the 1980s (Bennett et al., 2001). These
Water Availability and Its Use in Agriculture
changes are mirrored by phosphorus accumulation in soils, with high levels of phosphorus runoff. Excessive nutrient loading causes algal blooms, decreased drinking water quality, eutrophication of freshwater ecosystems and coastal zones, and hypoxia in coastal waters. In Lake Chivero, Zimbabwe, agricultural runoff is responsible for algal blooms, infestations of water hyacinth, and fish declines as a result of high levels of ammonia and low oxygen levels (UNEP, 2002). In Australia, extensive algal blooms in coastal inlets and estuaries, inland lakes, and rivers have been attributed to increased nutrient runoff from agricultural fields (Lukatelich and McComb, 1986; Falconer, 2001). Diffuse runoff of nutrients from agricultural land is considered a major cause for increased eutrophication of coastal waters in the United States as well as for the periodic development, often varying from year to year, of anoxic conditions in coastal water in many parts of the world, such as the Baltic and Adriatic Seas and the Gulf of Mexico (Hall, 2002). Extensive evidence shows that up to 80% of the global incidents of nitrogen loading can be retained within wetlands (Green et al., 2004; Galloway et al., 2004). However, the ability of such ecosystems to cleanse nutrient-enriched water varies and is limited (Alexander et al., 2000; Wollheim et al., 2001). Verhoeven et al. (2006) pointed out that many wetlands in agricultural catchments receive excessively high loads, with detrimental effects on biodiversity. Bioaccumulation, as a consequence of the wide use of agrochemicals, has had dire outcomes for many species that reside in or feed predominantly in wetlands or lakes where residues from pesticides have accumulated. The declined breeding success of raptors was a turning point in developing awareness about the dangers of pesticide use (Carson, 1962). An increasing amount of analytical and eco-toxicological data has become available for aquatic communities, and more recent research has also focused on risk assessments and the development of diagnostic tests that can guide management decisions about the use of such chemicals (van den Brink et al., 2003). Taylor et al. (2002) have highlighted the high levels of pesticide use and low levels of environmental risk assessment in developing countries. Vo¨ro¨smarty et al. (2005) reported that water contamination by pesticides has increased rapidly since the 1970s despite increased regulations, especially in developed countries, of xenobiotic substances (i.e., those chemical compounds foreign to a living organism). However, bans on the use of these chemicals have generally been imposed only two to three decades after their first commercial use, as with dichlorodiphenyltrichloroethane (DDT) and the common herbicide atrazine. Many of these substances are highly persistent in the environment, but because of the generally poor monitoring of their long-term effects, global and long-term implications of their use cannot be fully assessed.
4.21.3.1.2 Terrestrial ecosystems In many parts of the world, extensive sheet wash and gully erosion, due to poor land management practices, have had significant environmental effects. Large tracts of land have been devastated resulting in reduced agricultural productivity. Erosion has also contributed to the rapid siltation of reservoirs
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and increased sediment loads in many rivers (CA, 2007). On a regional scale, some reservoirs in Southern Africa are at risk of losing more than a quarter of their storage capacity within 20–25 years (Magadza, 1995). While many Australian and Southern African waters are naturally silty, many have experienced increased silt loads as a result of agricultural practices (Davies and Day, 1998). Zimbabwe’s more than 8000 small- to medium-size dams, for example, are threatened by sedimentation from soil erosion, while the Save River, an international river shared with Mozambique, has been reduced from a perennial to a seasonal river system in large part due to increased siltation caused by soil erosion. While it is not always easy to differentiate natural erosion from human-induced erosion. The high sediment loads carried by Asian rivers are partly a consequence of land-use practices, particularly land-clearing practices for agriculture that lead to erosion, a situation likely to continue as a consequence of the expansion of agriculture in Africa, Asia, and Latin America (Hall, 2002). Changes in the water table. Water builds up in a soil profile when the rate of input exceeds the rate of throughput (e.g., when irrigation volumes are greater than crop water consumption by way of evapotranspiration). This can cause water logging and salinization, which are extensively described for irrigated agriculture (Postel, 1998). Excessive irrigation can result in soil salinization in areas where the water table rises close to the surface and evaporation leaves salts behind in the soil profile. Salt-affected soils in irrigation schemes are often related to poor soil and water management, in addition to the unsuitability of many soils for irrigation. Clearing woody vegetation for pastures and crops can also lead to dryland salinization. Tree-covered landscapes provide an important regulating service by consuming rainfall through high evapotranspiration, limiting groundwater recharge, and keeping the groundwater low enough to prevent salt from being carried upward through the soil. Australia has had major problems with soil salinization as native woody vegetation was cleared in the 1930s for pastures and agricultural expansion (Farrington and Salama, 1996). Consumptive water use has declined there, the water table has risen, and salt has moved into the surface soils so that large tracts of land have become less suitable or unusable for agriculture (Anderies et al., 2001; Briggs and Taws, 2003). The overall trend, however, in irrigation is one of increased pumping and reduced water levels, but good salt management increases a critical issue in particular in arid regions. Moisture recycling. Increased irrigation and land clearing for agriculture have modified green water flows across the globe, reducing them by 3000 km3 through forest clearing and increasing them by 1000–2600 km3 in irrigated areas (Do¨ll and Siebert, 2002; Gordon et al., 2005). Changes in land cover affect evapotranspiration and ultimately impact the hydrologic cycle. It has been suggested that large-scale deforestation can reduce moisture recycling, affect precipitation (Savenije, 1995, 1996; Trenberth, 1999), and alter regional climates, with indications of global impacts (Kabat et al., 2004; Nemani et al., 1996; Marland et al., 2003; Savenije, 1995). Pielke et al. (1998) concluded that the evidence is convincing that land cover changes can significantly influence weather and climate and are as important as other human-induced changes for
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the Earth’s climate. Regional studies in West Africa (Savenije, 1996; Zheng and Eltathir, 1998), the United States (Baron et al., 1998; Pielke et al., 1998), and East Asia (Fu, 2003) have illustrated the ways in which changes in land cover affect green water flows, with impacts on local and regional climates. There are also indications that increased vapor flows through irrigation can alter local and regional climates (Pielke et al., 1997; Chase et al., 1999). The conversion of dry lands to irrigated croplands in Colorado resulted in a 120% increase in evapotranspiration, contributing to higher precipitation, lower temperatures, and an increase in thunderstorm activity (Pielke et al., 1997).
4.21.3.2 Health Impacts Among the agrochemicals that pose the greatest threats to domestic use of groundwater are nitrate and biocide residues. In addition, arsenic contamination in groundwater has emerged as a major health issue in Asia recently. Other health aspects concern malnutrition and vector-borne diseases. Many of the rural poor in Asia obtain water for drinking and household use from shallow aquifers under agricultural land. Irrigated rice fields can serve as breeding sites for mosquitoes, snails, and other intermediate hosts capable of transmitting human parasites. In particular, before transplanting and after harvest, puddles in rice fields are attractive breeding grounds for the mosquito Anopheles gambiae, Africa’s most efficient malaria vector. The conditions for mosquito breeding in rice fields have been identified and management practices, such as alternate wetting and drying of fields, exist to mitigate the problem. Moreover, countries such as Sri Lanka have made great strides in controlling epidemics through broad-based public health campaigns. Japanese B-encephalitis is highly correlated with rice irrigation in Asia, especially where pigs are also reared, as in China and Vietnam. Again, alternate wetting and drying can help reduce the breeding of disease vectors (Keiser et al., 2005a). Nitrate leaching from flooded rice fields is normally negligible because of rapid denitrification under anaerobic conditions (the following section follows from Bouman et al., 2007). In the Philippines, for example, nitrate pollution of groundwater under rice-based cropping systems exceeded the 10 mg l1 limit for safe drinking water only when highly fertilized vegetables were included in the cropping system (Bouman et al., 2002). In the Indian Punjab, however, an increase in nitrate of almost 2 mg l1 was recorded between 1982 and 1988, with a simultaneous increase in nitrogen fertilizer use from 56 to 188 kg ha1, most of it on combined rice–wheat cultivation (Bijay-Singh et al., 1991). These may lead to the blood disorder methemoglobinemia in human populations, especially in babies. Mean biocide use in irrigated rice systems varies from some 0.4 kg active ingredients per hectare in Tamil Nadu, India, to 3.8 kg ha1 in Zhejiang Province, China (Bouman et al., 2002). In the warm and humid conditions of the tropics, volatilization is the major process of biocide loss, especially when biocides are applied on water surfaces or on wet soil. Relatively high temperatures favor rapid transformation of remaining biocides by photochemical and microbial degradation, but little is known about the toxicity of the residual components.
In case studies in the Philippines, mean biocide concentrations in groundwater under irrigated rice-based cropping systems were one to two orders of magnitude below the single and multiple biocide limits for safe drinking water (i.e., 0.1 and 0.5 mg l1), although temporary peak concentrations of 1.14–4.17 mg l1 were measured (Bouman et al., 2002). Biocides and their residues may be directly transferred to open water bodies through drainage water that flows overland from rice fields. The potential for water pollution from biocides is greatly affected by field water management. Different water regimes result in different pest and weed populations and densities, which farmers may combat with different amounts and types of biocides. Agricultural use of untreated wastewater can affect human health through exposure to pathogens, parasite infections, and heavy metals. Leafy vegetables, eaten raw, can transmit contaminations from farm fields to consumers. Hookworm infections are transmitted by direct exposure to contaminated water and soils. A survey along the Musi River in India revealed the transfer of metal ions from wastewater to cow’s milk through fodder irrigated with wastewater. About 4% of grass samples showed excessive amounts of cadmium and all samples showed excessive lead levels. Milk samples were contaminated with metal ions ranging from 1.2 to 40 times permissible levels (Minhas and Samra, 2004). Farmers and their families using untreated wastewater are exposed to health risks from parasitic worms, viruses, and bacteria. Many farmers cannot afford treatment for some of the health problems caused by exposure. Generally, farmers irrigating with wastewater have higher rates of parasite infections than farmers using freshwater do, but there are exceptions (Trang et al., 2006). In addition, skin and nail problems occur more frequently among farmers using wastewater (Van der Hoek et al., 2002).
4.21.3.3 Environmental and Health Mitigation Agricultural water use is closely linked with health and environmental impacts. For health, the negative impacts of irrigation development can be mitigated through better design and operation of new and existing systems, especially through the multiple uses of irrigation water. Integrated approaches have taken many forms, including integrated river basin management, integrated land and water management, ecosystem approaches, integrated coastal zone management, and integrated natural resources management. These management strategies often seek to do the following: address the integration of a broad range of benefits and costs associated with land-use and water decisions, including effects on ecosystem services, food production, and social equity; involve key stakeholders at cross-institutional levels; and address interconnectedness across subbasin, river basin, and other biophysical scales. The MEA (2003) has provided a major advancement in understanding the links between the provision of ecosystem services and human well-being. Increased awareness is still needed on several different levels. The scientific knowledge of how ecosystem services contribute to human well-being within and between different sectors of society, and the role of water in sustaining these services, needs to be improved.
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Dissemination of information on these issues and dialog with stakeholders should be enhanced. Civil society organizations can help to ensure that appropriate consideration is given to the voices of individuals and social groups and to nonutilitarian values in decision making. Minority groups and disadvantaged groups, such as indigenous people and women, in particular, need to be heard. Women play a critical and increasing role in agriculture in many parts of the developing world (Elder and Schmidt, 2004). Mitigation measures are modifications to the design or operation of agricultural water development projects to reduce negative environmental and health impacts. However, present levels of understanding mean that very often some negative impacts are not foreseen prior to project implementation. Consequently, there should be a constant reevaluation of the need for mitigation measures throughout the life of a project. This requires monitoring so that measures can be introduced retrospectively when necessary. Monitoring enables health authorities to target resources and treatment interventions at times and locations of greatest need. A wide range of technical mitigation measures to prevent environmental damage has been developed for formal irrigation schemes. Measures that promote high water-use efficiency also tend to mitigate negative environmental and health impacts. For example, good water management and drainage are prerequisites to preventing water logging, decreasing habitats for mosquitoes, and minimizing salt accumulation in soils. Over the last 20 years, considerable progress has been made in the development of methods to determine environmental flows downstream of dams and extractions for irrigation. Increasingly, these techniques are taking a systems approach that includes holistic ecosystem assessments and works to predict the impacts of different flow regimes on the livelihoods of water users (Dyson et al., 2003). For example, downstream response to imposed flow transformation (DRIFT) is a scenario-based environmental flow assessment process designed specifically for use in negotiations over water resources. It is designed to quantify the linkages between changing river conditions, and the social and economic impacts for riparian people who rely on rivers for their livelihoods (Brown and King, 2000). Hydrological environmental flow assessment methods are being developed for use in areas where insufficient ecological data exist for conclusive analyses (Smakhtin and Shilpakar, 2005). Such approaches help define environmental targets and thus facilitate the design of mitigation measures by specifying desired environmental conditions. Concerning disease vectors, the primary approach is to design irrigation or pastoral water systems that do not provide habitats for vectors, while also conducting health education. In these cases, good design and construction of canal and drainage systems, and proper leveling of fields, can ensure that water is fast flowing and stagnant pools do not occur. Other options include direct vector control using chemicals or biological methods. However, with chemical control, care is required in application to ensure that the vectors do not develop resistance. Physical removal of habitats, for example, through manual or mechanical cutting of weeds is also possible. Good cleaning and preventative maintenance of all infrastructures, including canals, cattle troughs, hydraulic
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structures, and drains reduce the breeding of vectors and intermediate hosts, as well as improve irrigation performance. In several sub-Saharan African countries, weed control in canals has been applied as an effective method of vector control. However, routine cleaning work can itself be a health hazard. In the Gezira scheme in Sudan, canal cleaning personnel became the group most infected with the disease schistosomiasis (Fenwick et al., 1982). In some places, attempts have been made to minimize risk exposure by adapting the time of cleaning activities to the cycle of the parasite or by providing alternative tools (Euroconsult, 1993). A recent review of 40 largely pre-DDT interventions suggests that environmental modification can be a very effective malariacontrol strategy (Keiser et al., 2005b). Adapting water management to modify the vector habitat is another approach that has often been proposed in biomedical studies as an easy and cheap measure for vector control. However, there are very few examples where this type of environmental manipulation has been applied in practice (Matsuno et al., 1999; Laamrani and Boelee, 2002; Boelee and Laamrani, 2004). This is because, in reality, it is neither simple nor cheap to change established water management patterns. Water management interacts not only with vector breeding or disease transmission, but also with the irrigation system itself. Changes to water distribution often require modifications in design, notably the sizing of canals and type of structures. For example, if continuous delivery is replaced by rotation of the water flow to disrupt breeding sites, the discharge in the canals alters from constant low flows to intermittent high flows, requiring larger canals. At the same time, the wider human environment is influenced. With water flowing in the canals continuously, farmers can irrigate their crops at any time. With rotation, the flow has to be divided over time between users, demanding a higher level of organization. Water scheduling to meet crop water requirements is complicated, especially when conflicting interests between higher water-use efficiencies and farmers demanding flexibility have to be accounted for. If disease-control measures have to be observed as well, scheduling and management become very complicated (Boelee, 1999). Adaptive water management in rice fields may result in reduced vector breeding and hence reduced transmission of Japanese encephalitis and malaria (van der Hoek et al., 2001; Keiser et al., 2005b). However, these studies mainly report from Asia. In an African context, with constraints on resources and capacity, it may be especially difficult to achieve the required water deliveries and level of water management (Mutero et al., 2000). In reality, effective health interventions require an integrated approach that simultaneously implements avoidance and mitigation measures in collaboration between the water and health authorities.
4.21.4 Water Governance Improving agricultural water productivity in the future will take careful attention to the ways in which water use is linked to economies, social well-being, and ecological systems. Implementing these changes will require more effective water governance, which means rethinking how water is managed. Beyond national-scale water legislation, water governance
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encompasses various bodies of formal and informal regulations and institutions and also the way decision-making and political power is exercised. In short, it refers to the social mobilization and other actions designed to promote ownership, co-investment, capacity building, incentives for participation, and willingness to pay for services at the community level (UNDP, 2004). The following section outlines various forms of water governance and points to future challenges water governance will face.
4.21.4.1 Definition Governance is the way authority is organized and executed in society, and often includes the normative notion of the necessity for good governance (Merrey et al., 2007). 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. Effective water governance builds institutional capacity from the local level upward and empowers stakeholders with knowledge and the ability to make decisions about matters that directly affect their lives. It promotes the equal participation of women and men in decision making. Water governance is critical for resource planning and allocation among riparian states (those sharing a water basin) and vital for conflict resolution to defuse upstream–downstream tensions and balance the needs of different groups sharing water resources. Good water governance determines the appropriate role for the government in service delivery (i.e., as a facilitator or as a service provider) and ensures that water and sanitation services provided by both public and private actors meet the needs of the people they serve and do not fall prey to corruption. Good water governance corrects market distortions, perverse incentives, and pricing that shuts out the poor (UNDP, 2004).
basin, essential functions are partly or completely carried out, with their sum constituting basin governance (Table 3). Much attention has been given to the ideal organizational model for river basin management, while much less emphasis has been placed on the process of developing, managing, and maintaining collaborative relationships for river basin governance. More fundamentally, the essential function in river basin management – allocating water between competing uses and users, including the environment – has not received sufficient attention, although it is at the heart of integrated water resources management. Moreover, agricultural water and land practice, such as rainfed agriculture, livestock and fisheries practices often do not feature strongly. There are two main trends in basin governance. One trend concerns watersheds, or sub-basins of a limited size (typically from 10–1000 km2), where local stakeholders and agencies attempt to solve their land- and water-related problems. A second trend concerns the management of wider river basins. This trend has three salient aspects (Svendsen and Wester, 2005). First is the consensus that integrated water resources management should be carried out at the river basin level. This, together with the desire to realize the promise of integration, has placed river basin management on the agenda of
Table 3
Functions for river basin management
Function
Description
Plan
Formulation of medium- to long-term plans for managing and developing water resources in the basin. Activities executed for the design and construction of hydraulic infrastructure. Activities executed to maintain the serviceability of the hydraulic infrastructure in the basin. Mechanisms and criteria by which water is apportioned among different use sectors, including the environment. Activities executed to ensure that allocated water reaches its point of use. Activities executed to monitor water pollution and salinity levels and ensure that they remain at or below accepted standards. Flood and drought warning, prevention of floods, and development of emergency works, drought preparedness, and coping mechanisms. Provision of space or mechanisms for negotiation and litigation. Priorities and actions to protect ecosystems, including awareness campaigns. Harmonization of policies and actions undertaken in the basin by state and nonstate actors relevant to land and water management.
Construct facilities Maintain facilities
Allocate water
Distribute water
4.21.4.2 Types of Governance for River Basin Management The growing pressure on water resources and the increasing hydrological, social, and ecological interdependencies in closing river basins have led to widespread recognition of the need for holistic approaches to water management. There is a renewed emphasis on river basins as the most appropriate spatial unit for water management. The decision to manage water on the basis of river basins is a political choice, and river basins thus become a scale of governance in which tensions arise among effectiveness, participation, and legitimacy (Barham, 2001; Schlager and Blomquist, 2000; Wester and Warner, 2002). Progress in establishing adaptive, multilevel, collaborative governance arrangements for river basin management has been weak, with undue emphasis on form (setting up river basin organizations) over process. Although there may not be a central basin manager, this does not mean that river basins are not managed (Schlager and Blomquist, 2000). This can be accomplished by identifying the roles of various actors engaged in river basin water management, asking who does what, where, to what end, and how well. In any river
Monitor and enforce water quality Preparedness against water disasters
Resolve conflicts Project ecosystems Coordinate
Note: The functions listed here subsume supporting functions such as data collection and resource mobilization, which are not ends in themselves, but rather facilitate the higher-level functions listed. From Svendsen M, Wester P, and Molle F (2005) Managing river basins: An institutional perspective. In: Svendsen M (ed.) Irrigation and River Basin Management: Options for Governance and Institutions. Wallingford: CABI Publishing.
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governments and international funding agencies and has led to many new river basin initiatives. Second, the number of public and private sector actors involved in, or concerned with basin planning and management, is increasing, from environmental agencies and civil society or interest groups to regulatory bodies and service providers for agricultural, municipal, and industrial water users. With rising standards of living, urbanization, and continuing environmental deterioration, more diverse stakeholders and worldviews need to be integrated. Third, organizations associated with basin planning and management have become more specialized and differentiated into regulators, resource managers, and service providers (Millington, 2000). Regulation and standard setting are carried out in the public interest and are necessarily functions of government, but other tasks may be fulfilled by commercial or hybrid public–private organizations. River basin organizations cover a wide gamut of organizations with quite varied roles and structures. At first this may seem a source of confusion, but it also suggests that both the nature of the problems faced (e.g., development or management) and the particular history and context of each basin reflect on each river basin organization. The following typology can be inferred from a broad-brush review of river basin organizations, keeping in mind that there are no clear-cut definitions and that there are large variations in roles and power, even within the same category. In other words, the generic terms may not correspond to particular bodies. Basin authorities are autonomous executive organizations with extensive mandates for their river basin, undertaking most water-related development and management functions. They serve as regulator, resource manager, and service provider all in one. The Damodar Valley Corporation in India, the Mahaweli Authority in Sri Lanka, the Companhia de Desenvolvimento dos Vales de Sa˜o Franciscoe do Parnaiba in Brazil, and the Confederaciones Hidrograficas in Spain are examples of such basin authorities. Authorities generally exhibit poor responsiveness to local demands and are often undermined by bureaucratic conflict because they infringe on the competence of other government agencies and line ministries. Some of these authorities receive basin-wide, multifunctional mandates covering various domains but are not endowed with the legal, political, or administrative power to achieve them. They generally end up focusing on construction projects and dam management (mostly for hydropower or flood control). Examples include the Damodar Valley Corporation in India (Saha, 1979), the River Basin Development Authorities in Nigeria (Adams, 1985), and the China River Commissions (Millington, 2000). Some authorities were designed to ensure regional infrastructure development (the early River Basin Commissions in Mexico), others endured as powerful manager/operators (Brantas basin in Indonesia, Tarim in China), while others shrank and were confined to one issue or degenerated into powerless parallel structures with narrow scope and erratic funding (A. Dourojeanni, personal communication). Basin commissions or committees focus on policy setting, basin-wide planning, water allocation, and information management, with varying degrees of stakeholder participation. They are usually endowed with authority to manage water resources (allocating permits, defining taxation,
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negotiating water allocations, and defining effluent standards), and sometimes to plan future developments, but are not involved in operation or construction. Examples include the Delaware Commission in the United States, the Murray–Darling Commission in Australia, the British water authorities, and the French Agences de l’eau. Coordinating councils are deliberative decision-making bodies incorporating public and private stakeholders and integrating policymaking across different policy areas. They are not organizations in the strict sense, but rather bring together stakeholders from various agencies and water-use sectors. Their role is coordination, conflict resolution, and review of water resources allocation or management. Examples include the river basin councils in Mexico (Wester et al., 2005), the proposed catchment management agencies in South Africa (Waalewijn et al., 2005), the Zimbabwean catchment councils (Jaspers, 2001), the river basin committees and users commissions in Brazil (Lemos and Oliveira, 2004), and several river commissions in the United States. International river commissions are unique because coordination is achieved between countries rather than among stakeholders and because political dimensions are pervasive. They were frequently established as part of a treaty signed among riparian countries or to manage dams on shared rivers (e.g., Senegal, Volta, or Zambezi rivers) (Barrows, 1998). They mediate water conflicts through consultation and cooperation and may also manage common databases, and their work may lead to concrete agreements. From a governance perspective, institutional arrangements for river basin management may be distributed along two axes, one that distinguishes between state-driven and stakeholder-driven functioning, and the other that contrasts centralized and decentralized modes. This yields four models for basin governance: unicentric (state-driven, centralized), deconcentrated (state-driven, decentralized), coordination (stakeholder-driven, centralized), and polycentric (stakeholder-driven, decentralized). Under the unicentric model, a basin authority or line ministry manages the river basin. In the polycentric model, the actions of existing organizations, layers of government, and stakeholder initiatives are coordinated to cover an entire river basin or sub-basin.
4.21.4.3 Basin Governance Challenges River basin governance is about the emergence of the appropriate blend of government, civil society, and markets in decision making and regulation. In addition to greater control, rigor, and openness for water resource planning and allocation, as just described, integrated river basin management demands adequate governance. This brings out two main challenges: ensuring that all stakeholders, including the environment, have a voice, and coordinating uses and policies within the basin. Although frequently advocated as a key to achieving effective water management (Rogers and Hall, 2003), stakeholder participation in river basin management is not straightforward, and achieving substantive stakeholder representation have proved to be elusive in practice (Wester et al., 2003). Emphasizing participation in river basin management may draw attention away from the very real social and economic differences among people and the need for
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redistributing resources, entitlements, and opportunities. This is unlikely to happen without challenges, and decision makers committed to social equity need to devise mechanisms that strengthen the representation of marginal groups in river basin management and empower them. Stakeholder platforms, whether river basin councils, catchment management agencies, or watershed councils, democratize river basin management by giving voice to multiple actors. However, much depends on the institutional arrangements from which these river basin management institutions emerge, as many roles, rights, and technologies and physical infrastructure for controlling water are already in place. Stakeholders have different levels and types of education, differ in access to resources and politics, hold different beliefs about how nature and society function, and often speak different languages (Edmunds and Wollenberg, 2001). If these differences are not taken into account when creating new rules, roles, and rights, the institutional outcome can easily privilege those who are literate and have access to the legal system and eventually institutionalize inequality and power differentials instead of giving voice to marginal groups (Wester and Warner, 2002). This review of basin governance patterns identified the various types of organizations and arrangements for basin management. A strong civil engineering body capable of planning, designing, and constructing infrastructure to tap available water is useful and effective when resources are plentiful and management is not a strong requirement. In the later phases of basin closure, however, experience shows that large civil engineering organizations (and agricultural or other line agencies) are not well suited to deal with the challenges of basin governance. They have limited experience in political negotiation or interaction with key stakeholders and lack the breadth of experience in dealing with complex, broad-based issues, and multiple values. Further, they often tend to adopt stances based on vested interest in continuing infrastructure development, a position antagonistic to that of stakeholders with ecosystem concerns. Countries that have strong civil engineering organizations reluctant to cede any power will face intense negotiations and struggles before an acceptable form of river basin coordination emerges that is capable of undertaking the key tasks required. However, wherever the scope for construction is reduced and societal values have changed, the trend is likely to follow that of countries such as Australia and the United States, where engineering bodies have contracted and evolved into environmental agencies. Decision makers should not infer from the integrated water resources management message that river basin management needs a strong centralized organization. Basins facing complex problems of conflicting societal values and pressure on resources will probably not be well managed by a single body. Nested or polycentric patterns of basin governance, in which user and community organizations, layers of government, and stakeholder initiatives are coordinated at the basin level, perform better and can be especially effective in settings where participation and democratic practices are well established. Moving toward sustainable river basin management requires much more emphasis on developing, managing, and maintaining collaborative relationships for river basin governance, building on existing organizations, customary practices, and administrative structures.
Acknowledgments The material from this chapter was drawn largely from the Comprehensive Assessment of Water Management in Agriculture which drew together hundreds of researchers and practitioners to find water solutions for tomorrow.
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Relevant Websites http://www.maweb.org Millennium Ecosystem Assessment.
