Survey
* Your assessment is very important for improving the workof artificial intelligence, which forms the content of this project
* Your assessment is very important for improving the workof artificial intelligence, which forms the content of this project
CHAPTER 3 Protozoa in Wastewater Treatment: Function and Importance Wilfried Pauli 1, Kurt Jax 2, Sandra Berger 1 1 2 Institut für Biochemie und Ökotoxikologie, Freie Universität Berlin, Ehrenbergstr. 26–28, D-14195 Berlin, Germany, E-mail: [email protected] Zentrum für Ethik in den Wissenschaften, Universität Tübingen, Keplerstrasse 17, D-72074 Tübingen, Germany Protozoa constitute a major link between the highly productive and nutrient retaining microbial loop and the metazoans of the classical food web. Protozoa are efficient at gathering microbes as food, and they are sufficiently small to have generation times that are similar to those of the food particles on which they feed. They are, in quantitative terms, the most important grazers of microbes in aquatic environments, balancing bacterio-plankton production. Protozoa not only play an important ecological role in the self-purification and matter cycling of natural ecosystems, but also in the artificial system of sewage treatment plants. In conventional plants ciliates usually dominate over other protozoa, not only in number of species but also in total count and biomass. It is generally accepted that their feeding on bacteria improve the treatment, resulting in a lower organic load in the output water of the treated wastes. Due to their biodegradation potential some attempts have been made to use ciliates specifically in environmental biotechnology. As biosensors they could provide valuable information regarding adverse effects of environmental chemicals on this part of the biocoenosis essential for the effective operation of biological waste-water treatment processes. Keywords. Protozoa, Ciliates, Ecology, Sewage treatment, Environmental biotechnology 1 Ecological Role of Aquatic Protozoa with Special Regard to Ciliates Within the Microbial Food Web . . . . . . . . . . . . . 205 1.1 1.2 1.3 1.4 Introduction . . . . . . . . . . . . . . . . . . . . Traditional Food Webs and Microbial Food Webs The Role of Protozoa in Aquatic Food Webs . . . Outlook . . . . . . . . . . . . . . . . . . . . . . . 2 Protozoa in Wastewater Treatment . . . . . . . . . . . . . . . . . . 212 2.1 2.1.1 2.1.2 2.1.3 2.1.4 2.2 2.2.1 2.2.2 2.2.3 2.2.4 2.2.5 2.2.5.1 Background . . . . . . . . . . . . . . . . . . . . . . . Wastewater . . . . . . . . . . . . . . . . . . . . . . . Biological Treatment Processes . . . . . . . . . . . . Bacterial Biofilms . . . . . . . . . . . . . . . . . . . . Activated Sludge . . . . . . . . . . . . . . . . . . . . Protozoa in Biological Wastewater Treatment Plants Occurrence . . . . . . . . . . . . . . . . . . . . . . . Species Composition . . . . . . . . . . . . . . . . . . Plant Specific Basic Communities . . . . . . . . . . . Biomass . . . . . . . . . . . . . . . . . . . . . . . . . Ecological Framework . . . . . . . . . . . . . . . . . Sludge Loading . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 205 205 208 211 212 212 214 216 216 217 217 218 220 221 221 223 The Handbook of Environmental Chemistry Vol. 2 Part K Biodegradation and Persistence (ed. by B. Beek) © Springer-Verlag Berlin Heidelberg 2001 204 W. Pauli et al. 2.2.5.2 2.2.5.3 2.2.5.4 2.3 2.3.1 2.3.2 2.3.2.1 2.3.2.2 2.3.2.3 2.3.3 2.3.4 2.3.5 2.3.6 2.3.7 Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . pH-Value . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . O2-Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Significance of Protozoa for Wastewater Treatment . . . . . . . . Nutrition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reduction and Elimination of Suspended Particles and Bacteria Clearing Rate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Experimental Findings . . . . . . . . . . . . . . . . . . . . . . . . “Field”-Observations . . . . . . . . . . . . . . . . . . . . . . . . . Elimination of Dissolved Substances . . . . . . . . . . . . . . . . Flocculation and Composition of the Bacterial Community . . . Reduction of the Total Biomass . . . . . . . . . . . . . . . . . . . Influence of Protozoa on Bacterial Metabolism . . . . . . . . . . Filamentous Bacteria and Protozoa . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 223 223 224 225 225 227 227 228 231 232 232 235 237 239 3 Impairments of Protozoa: Consequences for Water Purification . 241 4 Environmental Biotechnological Aspects . . . . . . . . . . . . . . 243 4.1 4.2 Biodegradation Potentials of Ciliates . . . . . . . . . . . . . . . . . 243 Ciliates as Biosensors . . . . . . . . . . . . . . . . . . . . . . . . . . 245 5 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 246 List of Abbreviations BOD(5) COD dw fm biological oxygen demand (index: within 5 days) chemical oxygen demand dry weight sludge loading [g BOD (g MLSS day)–1 or g BOD (g MLVSS day)–1], also known as “food to micro-organism (F/M) ratio” F/M-ratio see fm MLSS Mixed-liquor suspended solids, sludge solids (g m–3; concentration of the suspended solids in an aeration tank including inorganic matter) MLVSS Mixed-liquor volatile suspended solids (g m–3; corresponds to the organic, i.e., combustible content of the sludge, which amounts to ca. 70% of the sludge solids: 0.7 MLSS≈MLVSS; this parameter is often used as indicator of microbial concentration, although it does not distinguish between biochemically active material and inert or dead material in the sludge) EC/LC50 50% effective and lethal concentration, respectively Protozoa in Wastewater Treatment: Function and Importance 205 1 Ecological Role of Aquatic Protozoa with Special Regard to Ciliates Within the Microbial Food Web 1.1 Introduction There is hardly any place on earth in which protozoa cannot be found. They are abundant in terrestrial as well as in aquatic systems. In the latter they are present in high numbers of species and individuals both in the oceans and in freshwater habitats. Some taxa live attached to solid substrates or within the sediment, some as part of the plankton. An overview of the data about the abundance of protozoa in aquatic habitats gives a first indication that these organisms are not negligible in aquatic environments – although in fact they are still often neglected. In the plankton of highly productive lakes, densities of small flagellates (< 20 mm body size) of more than 106 cells per ml were reported [1] and in studies on the periphyton of small bodies of waters maximum values of more than 1350 cells per cm2 of the much larger testate amoebae specimens were encountered [2]. However, these numbers do not make any statements about the ecological interactions in which the species are involved and the role they play within those processes which mostly are seen as the essence of ecosystem dynamics, namely the fluxes of energy and material. It is the objective of this paper to provide a short introduction to the current knowledge of these roles as regards aquatic environments. 1.2 Traditional Food Webs and Microbial Food Webs Traditionally, food webs in aquatic systems were illustrated as in Fig. 1. Going back to the limnologist August Thienemann, the different species within a body of water were characterized by the categories of producers, consumers of different order (primary consumers, secondary consumers and so on) and decomposers [3]. The latter live on the dead organic matter and mineralize the organic compounds to inorganic nutrients, e.g., phosphorus, nitrogen, etc. These categories were also the basis on which Raymond Lindeman [4] built his famous trophic dynamic concept of ecology which was the first implementation of Arthur Tansley‘s ecosystem concept [5]. Energy enters the system as light and is processed as organic matter along the food chain or food web until most of the energy is dissipated by respiration. In aquatic habitats these functional categories – trophic levels in Lindeman’s parlance – were commonly attributed to phytoplankton (producers), zooplankton (primary consumers), and different kinds of vertebrates on the higher trophic levels. Protozoa and particularly bacteria were seen as decomposers, mainly restricted to sediments and other surfaces, but of minor importance in the pelagic food web. This association of bacteria and protozoa with decaying matter was recognized and used for applied purposes rather early. Protozoa were used as bioin- 206 W. Pauli et al. Fig. 1. Diagram of the “classical” food web in lakes. Modified, according to [6] dicators for the saprobic states of natural and manmade freshwaters as early as 1908 (e.g., [7, 8]). Their dynamics in the process of decomposition of organic substances were clarified by the middle of the century. Meanwhile, classical studies on this topic were made by Bick and co-workers (e.g., [9, 10]) who investigated the succession of micro-organisms, in particular ciliated protozoa, in the course of the “self-purification” of water enriched with sewage and other organic substances. However, during the last two decades there have been some new insights which have broadened and fundamentally changed our way of looking at the water of lakes and oceans and which affect the role protozoa and other microorganisms are supposed to play within aquatic systems. These insights were initiated by the appearance of some new actors on the stage of the ecological theater which also radically changed the roles in which protozoa were perceived. In 1974 Pomeroy [11] presented a paper in which he developed new ideas about the interactions of the pelagic organisms. Although these ideas were first developed in connection with marine systems they were soon transferred to freshwater habitats. The main point made is that, besides and connected with the classical “macroscopic” food web, there exists a microbial food web. The reason why these microbial food webs were discovered so late can, to a high degree, be attributed to the development of new methods in aquatic ecology. 207 Protozoa in Wastewater Treatment: Function and Importance By the early 1970s it was recognized that an important part of the pelagic organisms had been neglected as a result both of the methods used and of the theories regarding interactions in the water. Using direct counts of bacteria with epifluorescence methods instead of plate counts, it turned out that the abundance of bacteria in the open water had been underestimated by orders of magnitude. Only 0.1–1% of the actual abundance had been counted [12]. Furthermore, most investigations of marine and freshwater plankton used plankton nets with a mesh size of 20 mm or even 60 mm, while all smaller organisms were thought to be of minor importance. Finally, the methods of conserving planktonic protozoa were inadequate and even larger protozoa were neglected or underestimated as components of the pelagic species assemblages [13]. What was collected and counted were those fractions of the plankton which we now call the micro- and macroplankton, i.e., organisms bigger than 20 mm (Table 1). Thus, not only all smaller organisms, the pico- and nanoplankton – consisting of bacteria, Cyanobacteria, small protozoa, and small eukaryotic algae [14] – but also many larger protozoans were to a large extent excluded from the quantitative sampling. However, it turned out that especially this small sized fraction of the plankton is of extreme importance in terms of energy- and material fluxes. New measurements revealed that the major part of the metabolic activity in plankton was displayed by the size fraction below 10 mm [15]. The most productive component of the pelagic food webs was not, as thought earlier, the planktonic eukaryotic algae of the microplankton, but the tiny Cyanobacteria, mostly of the genus Synechococcus, and some small eukaryotic algae. The percentage of primary production in terms of carbon varies between 1% and 90% in marine waters – with higher ratios in more oligotrophic conditions – and 16–70% in fresh waters [16]. For oligotrophic lakes 50–70% are documented, while the autotrophic picoplankton amounts to 10–45% of the total phytoplankton biomass (standing stock, measured as chlorophyll) [17]. Data for marine habitats give estimates of 20–80% [18]. Similarly, the abundance of heterotrophic picoplankton, i.e., heterotrophic bacteria, is much higher than previously thought and can approach 109 cells in highly eutrophic fresh waters [1]. However, the new theory incorporates some new links rather than just adding picoplankton to the classical food web. Figure 2 presents a very simple diagram of what a microbial food web might look like, given the current status of knowledge. Table 1. The size classes of planktonic organisms Picoplankton Nanoplankton Microplankton Macroplankton 0.2–2 mm Bacteria Cyanobacteria Algae Rhizopods Flagellates Ciliates 2–20 mm Algae Flagellates Ciliates Ciliates 20–200 mm Algae Rhizopods Crustaceans Rotatoria Nauplii > 200 mm Ciliates Rotatoria Fish larvae 208 W. Pauli et al. Fig. 2. The food web of the lake plankton. The classical food chain (open circles) is supplemented by the elements of the microbial loop (filled ovals and square). DOM: dissolved organic matter The earlier food chain from algae via macrozooplankton to fish still exists but is supplemented by a new section which is commonly called the microbial “loop.” This consists of the picoplankton (“algae,” i.e., Cyanobacteria and heterotrophic bacteria), protozoa, and a compartment of non-living material, i.e., dissolved organic matter (DOM). DOM is lost and excreted in substantial amounts by both algae and Cyanobacteria and constitutes the energy source for the heterotrophic bacteria. The rate of fixed carbon lost by phytoplankton cells may vary between 10% and 40% depending on the physiological status of the cells [13]. The picoplankton is grazed by protozoa which themselves are preyed upon by the metazoan zooplankton, thus coupling the microbial loop to the traditional parts of the food web. As cells with a size of up to 2 mm hardly get lost through sedimentation, the microbial loop not only adds some new links to the classical food web but keeps the nutrients (DOM and inorganic nutrients) within the water body and minimizes losses to the deeper, non-productive regions of the waters or even the sediment. This seems to be particularly important during the summer stratification of oligotrophic lakes, in which the epilimnion, the upper and photosynthetically active region of the lake – the euphotic zone – is temporarily cut off from the richer nutrient supply of the deeper waters [17]. 1.3 The Role of Protozoa in Aquatic Food Webs From this scheme the new role of protozoa within the food webs of aquatic systems seems obvious. They are not only – in the same way as bacteria – decomposers associated with the decay of organic material, but they are a link between Protozoa in Wastewater Treatment: Function and Importance 209 the highly productive and nutrient retaining microbial loop and the metazoans of the classical food web. Most microplankton organisms are unable to utilize particles smaller than 5 mm directly [18]. Protozoa “repack” the organic material into edible portions and thus make it available to crustaceans, rotatoria, and other metazoans. There is empirical evidence that planktonic protozoa graze effectively on picoplankton and also that protozoa constitute a valuable diet for crustaceans [19]. Thus both necessary links between picoplankton and metazoa have been established. The details of the microbial webs, however, are still the subject of research and discussion. The specific pathways and the number of steps over which energy and nutrients are transferred are subject to much variation. There is temporal variation, e.g., seasonally, [20] and there is spatial variation both within lakes and even more if different lakes are compared. The compartment of protozoa can be divided in several ecological relevant ways. Not only is there a taxonomic division between flagellates and ciliates, but also a physiological one, relating to the nutritional mode (autotroph, heterotroph, mixotroph, etc.) which does not correspond with the classic taxonomic or “trophic level” boundaries [21]. Furthermore the body sizes of the different taxa are important features for their position within the food webs. In many cases bacteria are grazed upon mainly by small heterotrophic flagellates, the heterotrophic nanoplankton (HNAN), which in most cases turned out to be the most efficient predators of bacteria that were able to control the bacterial populations even during their highest productivity (e.g., [1, 22]). Berninger et al. [1] found a clear correlation between the abundance of bacteria and HNAN in comparing samples from more than hundred freshwater sites of different trophic states. The numbers of the two groups of organisms differed by two or three orders of magnitude, with maxima of more than 106 specimen of HNAN and 109 specimens of bacteria per ml. They inferred predator-prey relationships between these groups. HNAN are sometimes grazed upon directly by metazoa, while in other bodies of water ciliates constitute the main predators [17, 23]. Heterotrophic flagellates, possessing high turnover rates, inhabit a central position in the transfer of organic carbon in most microbial food webs. But what about the ecological roles of ciliates? In some cases, especially in productive waters, ciliates can also graze effectively on picoplankton and can even be the most important bacterivores, taking a key position for the transfer of matter to the metazoan links [23]. However, smaller bacterivorous ciliates with high grazing efficiencies need a threshold abundance of bacteria to persist on this diet. Beaver and Crisman [24] gave an estimate that small ciliates (20–30 mm) were “largely excluded from lakes having <5 ¥ 106 – 5 ¥ 108 bacteria ml–1 – a concentration normally found only in more productive systems.” Large ciliates (> 50 mm), being mainly phytophagous and grazing on nanoplanktic algae, dominate the ciliate assemblages in oligotrophic lakes, with low bacterial abundance. Mixotrophic ciliates with endosymbiotic algae can even contribute substantially to pelagic autotrophic biomass in some lakes (15% of annual total [25]). The overall number of planktonic ciliates in lakes is correlated with the trophic state of the water bodies.While under oligotrophic conditions abundancies 210 W. Pauli et al. of 3–10 cells ml–1 were recorded, 90–215 cells ml–1 were recorded in hypereutrophic waters [25]. The length of the food chain originating from bacteria and Cyanobacteria and the identity of links involved is important to the still unresolved question as to whether the microbial loop is acting as a link or a sink for organic material. Adherents of the latter position argue that a microbial food chain with four steps will be unlikely to transfer any substantial amount of organic carbon to the metazoan part of the web [15, 26, 27]. The answer to this question is dependent on several variables. Besides the trophic states of the waterbodies, other abiotic variables such as temperature and acidity are relevant for the specific patterns of the microbial web [25] and also the species composition of the whole food web [28]. In some cases organic material is transferred from picoplankton via heterotrophic flagellates to larger ciliates and then to crustaceans or other metazoans. In other cases crustaceans may directly feed on nanoplankton, while ciliates are of minor importance [29]. Even though most metazoans cannot feed effectively on small particles of the order of few mm, some freshwater species, in particular cladocera of the genus Daphnia, can effectively control bacterial abundance (although they may not persist on bacteria alone), thus shortcutting the microbial loop [17, 28]. The presence or absence of a single species can thus change the pathways completely, deciding the coupling or decoupling of the microbial loop from the metazoan web. The proportion to which different groups of organisms contribute to different nutritional types in a lake is also seasonally variable [17, 20, 28]. In this regard, the scheme displayed in Fig. 3 comes closer to the perceived processes than many other representations, in that a multitude of pathways is possible which may be more or less important at different times. Fig. 3. Diagram of the food web in lake plankton. In contrast to the scheme in Fig. 2, the compartment of protozoa has been differentiated. Note that not all pathways are realized at any one time. See also text. DOM: dissolved organic matter Protozoa in Wastewater Treatment: Function and Importance 211 As mentioned above, the microbial loop is not only important for the transfer of energy in the form of organic carbon, but also for the cycling and retention of nutrients. This is especially important in oligotrophic situations, where nutrients like phosphorus and nitrogen are scarce – at least during certain times of the year. The phosphorus dynamics of the pelagic zone seem to be strongly determined by the interactions of algae, bacteria, and protozoan grazers. Algae and bacteria compete for P, with bacteria being more efficient in the uptake of P. Bacterial grazing by protozoa was demonstrated to enhance phosphorus turnover and mineralization [30]. As grazed bacteria populations grow faster their excretion of P also becomes stronger. Furthermore, protozoan grazers increase the amount of organic P by excretion, which seems to be of special importance for phytoplankton [31]. Although this compound is also excreted by micro- and macrozooplankton, the high metabolic rate of protozoa leads to higher excretion rate of this group of organisms. Buechler and Dillon [32] estimated that if ciliates only contribute 1% to the biomass of a zooplankton assemblage, they should be able to contribute 50% to the release of dissolved P. A similar situation exists with regard to nitrogen in cases where nitrogen is a limiting factor for the growth of algae and bacteria. Bacteria can also outcompete phytoplankton for N and thus serve as a sink for nitrogen within the food web. However, as has been demonstrated experimentally, the presence of bacterivorous protozoan grazers leads to a partial remineralization of N and allows an increase in algal biomass [33]. The degree to which this process is of importance depends on the carbon available for the bacteria. As Caron et al. [33] concluded: “the role of bacterivorous protozoa as mineralizers of a growthlimiting nutrient is maximal in situations where the carbon:nutrient ratio of the bacterial substrate is high”. 1.4 Outlook Most of the interactions described above were investigated in the pelagic part of aquatic habitats. However, as mentioned above, many protozoa are closely related to surfaces within the water bodies, be they sediments, plants, and stones, or even microscopic aggregates within the pelagic zone. In lakes or oceans the main metabolic activity is certainly associated with the pelagic zone. Regarding streams or small water bodies, the surface-related biota gain in importance for the fluxes of energy and materials. In streams, a true plankton only exists in the slow flowing lower reaches of large rivers. Thus, most organismic activities are found in and on the benthic parts. Many of the aspects discussed above will also be valid in these environments. However, there will surely be differences. Although some data is available on the numbers and production of protozoa in these microhabitats [34–36], our understanding of the complex web of interrelations is much less than for the open water. To a considerable degree this seems to be a consequence of the methodical difficulties. Benthic assemblages are highly heterogeneous in space and time and this heterogeneity, i.e., the small scale spatial arrangement of the different components, is by itself of importance for the nature of the interactions between protozoa and the other parts of these 212 W. Pauli et al. assemblages. Thus we are only just beginning to delve deeper into the complicated patterns and dynamics of those biofilms. There is now important evidence that these biofilms are also highly productive but also very retentive in regard to nutrients [37]. Nutrient pulses are retained much longer within the periphyton assemblages of streams than would be expected on the basis of a continuous water flow. There are certainly many other important ways in which protozoa are involved in the ecology of aquatic systems. For example, little is known about informational relations between protozoa and other members of the species assemblages, although there may be indications in this direction (e.g., [38]). Also, our view of microbial food webs may change during the next years with the new awareness that even the pelagic zone of lakes is not as homogenous as it seems at first sight. In addition to rather macroscopic stratifications of abiotic factors and the related stratifications of organisms, the role of tiny and – in the realm of human time-scales – fleeting aggregates of small detritus particles, bacteria, protozoa and algae come into prominence, the so called “lake snow.” These aggregates may turn out to be hot spots of microbial activity, and especially for the grazing activities of protozoa. There are data that indicate that ciliate bacterivory is especially high in lakes with high amounts of suspended organic matter [39]. Similar to biofilms on solid substrates, the microenvironment on, in, and around these aggregates can be chemically strangely different from the average water column data. It remains to be seen, what these new insights will bring about for the understanding of the ecological processes in freshwater habitats. 2 Protozoa in Wastewater Treatment 2.1 Background 2.1.1 Wastewater Wastewater includes municipal, industrial, and agricultural wastewater as well as rainwater. The relative proportions of wastewater for West Germany (1980) were 32% municipal, 47% industrial, and 1% agricultural wastewater, plus 20% rainwater run-off in areas with main drainage. All wastewater produced in towns and communities is termed municipal sewage. This expression covers domestic wastewater (50%), extraneous water (leachates 14%), and wastewater from industry and commerce (36%) [40]. Municipal sewage is treated as follows: – Initial mechanical purification or sedimentation – Biological purification or clarification – Further purification, e.g., elimination or reduction of the nitrogen, sulfur, or phosphate content, polishing, filtration – The treated wastewater is then discharged into the receiving stream (Fig. 4) Protozoa in Wastewater Treatment: Function and Importance 213 Fig. 4 a– c. Types of common sewage treatment plants – flow diagram of: a activated sludge plants; b, c biofilm processes (trickling filter and Rotating Biological Contactor, RBC, respectively). In the activated sludge process (a) the wastewater is exposed to a mixed microbial population in the form of a flocculent suspension. In fixed medium systems the wastewater is brought into contact with a film of microbial slime (b) on the surfaces of the packing medium, (the wastewater trickles through the bed, most commonly consisting of stacked stones), or (c) on a partly submerged support medium which rotates slowly on a horizontal axis in a tank through which the wastewater flows 214 W. Pauli et al. Table 2. Average contribution of settleable (sedimentation within 2 h) and non-settleable matter and their respective biochemical oxygen demand (BOD5) to the total organic load of municipal sewage, according to [157] Organic load (in total ca. 450 mg/l) Æ Æ Settleable: 33% (w/v) or 150 mg/l, 33% (BOD) Non-settleable: 67% (w/v) or 300 mg/l, 67% (BOD) – Æ Æ Dissolved: 83% (w/v) or 250 mg/l, 75% (BOD) Suspended: 17% (w/v) or 50 mg/l, 25% (BOD) All substances present in sewage are classified according to their significance for wastewater treatment plants. Organic content is of particular importance for degradation processes. It is quoted in terms of the chemical or biochemical oxygen demand (COD, BOD) of the organic substances. Furthermore, a differentiation is made between suspended and dissolved wastewater components. Approximately two-thirds of the total load (organic and inorganic) of municipal sewage is in solution. With regard to the organic load almost half is in solution, the rest consists of colloidal material (25%) or is bound to particles which sediment (75%). Similarly, about half of the oxygen demand of biochemically degradable organic compounds is attributed to the dissolved fraction, of the other half one third to floating and two thirds to particulate matter. After a 2 h sedimentation period, two-thirds of the total organic load remains in the supernatant (also two-thirds of the total BOD). About 25% of the dissolved organic load is bound to colloids and particles which do not sediment (Table 2). Carbohydrates are not usually present in municipal wastewater plants. They are metabolized on route in the sewage. Proteins are also hydrolyzed in the sewers. The main task of the wastewater treatment plant is then to eliminate fatty acids and the amino acids formed by protein hydrolysis. Municipal sewage averages an organic load of 300 mg BOD5 l–1 (ca. 450 mg l–1 organic content). Activated sludge plants aim for effluent values < 20 mg BOD5 l–1, i.e., a reduction in the organic content of more than 90% [41]. For industrial – as opposed to municipal – wastewater, no generalizations can be made regarding type and amount of load. Diverse organic and inorganic loads are produced by different industrial sectors. Even within a sector values vary according to the production methods and environmental requirements. Wastewater from the chemical industry often exhibits toxic or inhibitory effects. 2.1.2 Biological Treatment Processes It is well known that a microbial degradation of organic substances takes place in natural flowing waters. This natural, self-purifying capacity of water became overtaxed by the increase in population and industrialization. Attempts were then made to pre-treat partially or fully sewage by mechano-biological processes, before discharging it into the surface water. Protozoa in Wastewater Treatment: Function and Importance 215 A conscious use of biological degradation began after bacteria were discovered in the nineteenth century. Two principles were implemented: activated and fixed-bed processes. The latter have been in use since 1882 and utilize the slime growth of organisms in the receiving stream. The activated sludge process, which takes advantage of the self-purification properties of the suspended organisms in the receiving water body, was developed in 1913, and the first German plant was operational in 1926 [42]. Both methods are still in use today. In Germany the activated sludge technique has taken precedence, due to its higher performance capacity, particularly for extended wastewater treatment including nutrient elimination. However fixed-bed reactors in combination with activated sludge techniques are finding increased application today. As submerged aerators they increase the active biomass and the age of the sludge in activated sludge plants, making a positive contribution to the purification efficiency [43]. The underlying principle of biological wastewater treatment is to transform the majority of dissolved and suspended substances into biomass which can then be removed either by sedimentation (activated sludge) or by fixing (submerged aerator contactors). In this way, a nutrient concentration exceeding the degradation capacity of local surface waters, resulting in disruption or even destruction of natural biological systems, can be avoided: Direct discharge of substances would result in anaerobic or aerobic burdening of the sediment of surface waters; high oxygen consuming, organic content (BOD5) in the effluent can overtax the oxygen household of the water, through its rapid conversion by heterotrophic organisms; direct discharge of plant nutrients, particularly nitrogen compounds and phosphates, encourages algal growth, with negative effects on the water (larger pH- and O2-fluctuations, sludge formation). At the same time, however, the discharge of bacteria – used for the fixation of wastewater substances – should be kept to a minimum. All biological processes have in common that they involve sectors of natural metabolic cycles. In wastewater treatment plants, the only difference from natural processes is that part of the reaction chain is technically controlled. The performance is dependent not on one specific species with a high degradation capacity, but on the interaction of a wide range of different organisms. Over the last 20 years the traditional model of a vertical material and energy flow, starting from nutrients through to decomposers and primary producers and both primary and secondary consumers, has been replaced by a more complex ecological web, which takes into account the network of microbial systems and their significance for turnover of matter (see Sect. 1.2). In treatment plants, due to the high organic content of the wastewater, a biocoenosis of organisms forms, primarily made up of members of the group of decomposers, i.e., saprophytic bacteria. The majority of the bacteria degrade dead organic matter, in the presence of oxygen, to carbon dioxide and water. Nitrogen is released in the form of ammonia. Bacteria are significant in wastewater treatment due to their large surface area in relation to their body volume and their associated high metabolic and reproductive rates. Apart from these prokaryotic forms of life, protozoa (unicellular, animal organisms) are the next most important group of organisms in the wastewater biocoenosis. Together 216 W. Pauli et al. with bacteria they form a closely related microbial system which forms the basis of the so-called natural self-purification process. 2.1.3 Bacterial Biofilms In both fixed-bed and activated sludge processes, microbial biofilms – either as slime growth or flocs – are fundamental for the turnover of organic waste. The colonization of surfaces by bacteria is a widespread process in the environment. In natural biotopes, bacteria favor the colonization of suspended particles and sediment. By far the majority (99%) of all bacteria in the environment adhere to surfaces such as stones, sediment, and soil. Important physico-chemical processes, forming the basis for the biomass layer, precede the attachment of a biofilm. Dissolved organic molecules (polysaccharides, proteins, humic acids) accumulate spontaneously on the surface of very different materials forming a “conditioning film,” on which bacteria colonization follows. The cells are immobilized and produce extra-cellular polymeric substances which anchor the organisms to the surface and to each other. Embedded in this matrix, microbial communities of complex composition are built up, usually in several layers. Biofilms are not static systems, rather a dynamic equilibrium exists between freely suspended bacteria and those adhering to particles. From the moment a bacterial biofilm forms, a detachment of cells or cell-aggregates takes place [44], dependent on the prevailing conditions. Several bacteria species, dependent on their nutrient supply, can exist either freely suspended or mainly aggregated in both pure and mixed cultures [45]. 2.1.4 Activated Sludge Existing literature regarding protozoa and wastewater treatment deals mainly with aerobic processes, with the focus on activated sludge technology. This is due to the significance of this technology for wastewater treatment on the one hand and that suspended activated sludge is more easily accessible for biological investigations than slime-growth areas of fixed-bed reactors on the other. Activated sludge processes operate with typical sludge concentrations between 2–3 g l–1 [46]. About 70% of the activated sludge is organic content and 30% inorganic (clay: Si; Al; Fe; ferric oxide; calcium phosphate) [47]. Non- – or not easily – oxidizable organic matter makes up 20–25% of the sludge [41]. In a conventional activated sludge tank flocculate suspended material contains about 6 ¥ 109 bacteria ml–1, i.e., 1–3 ¥ 1012 bacteria g–1 dry weight [48]. They represent about 90% of the total biomass of the activated sludge. The proportion of living or metabolically active bacteria found in the flocs varies considerably, depending on the method of analysis. Estimates based on glucose, stearate and acetate uptake rates imply active proportions of 8–13%, 14–28%, and 5–10% of the total biomass, respectively [48]. More recently, direct measurements by fluorescence-microscopy indicate a proportion of 35–40% (de- 217 Protozoa in Wastewater Treatment: Function and Importance hydrogenase activity [49]) and 70% (rRNA directed oligonucleotide probes, [50]), whereby a similar level of activity was assumed for all zones of the floc [51]. 2.2 Protozoa in Biological Wastewater Treatment Plants 2.2.1 Occurrence Systematic investigations at a large number of wastewater treatment plants reveal protozoa as typical components of the biocoenosis (Table 3). Thus, for example, in all ten South African activated sludge plants studied by Bux and Kasan [52] “basic communities” of protozoa, typical for sewage plants were found. Similarly, Curds and Cockburn [53] found protozoa biocoenoses in 53 of 56 British activated sludge plants and all 52 biological percolation filter plants studied. In New Jersey, Chung and Strom [54] found protozoa in all the rotating disc contactors and according to Madoni and Ghetti [55], typical ciliate communities were detected in 38 of 39 activated sludge plants and 47 of 49 rotating disc contactors in the Emilia region of Italy. The presence of protozoa is closely associated with biofilms and restricted mainly to aerobic processes and therefore to certain areas of the wastewater treatment plant; only a few specialists among the protozoa take part in anaerobic processes. Thus protozoan communities can be typically encountered in activated sludge tanks as well as in the sedimentation tanks, whereas no protozoa are found in sludge digestion or in the supernatant of the sedimentation tank (effluent), with the exception of malfunctions [56]. Table 3. A survey of the protozoan fauna in sewage treatment plants (only microfaunistic in- vestigations based on ten and more plants are taken into consideration), according to [52–55] Type of plant Activated sludge No. of plants investigated (country) 56 (Great Britain) 39 (Italy) 10 (South Africa) Trickling filter 52 (Great Britain) Rotating biological 49 (Italy) contactor 10 (USA) a b Occurrence of typical protozoan communities Typical protozoan communities absent Protozoa absent Within 53 plants Within 38 plants b Within all 10 plants Within all 52 plants Within 47 plants b 2 plants a 1 plant b – – 2 plants b 1 plant ? – – ? Within all 10 plants – – No ciliates, but flagellates present. Only ciliates investigated, no comments on other protozoan groups such as flagellates and amoebae. 218 W. Pauli et al. 2.2.2 Species Composition The majority of microfaunal investigations confirms that all of the three main groups of protozoa – flagellates, ciliates, and amoebae (naked and shell) – can be found in wastewater treatment plants, whereby ciliates form the largest proportion with regard to biomass and number of species, both in activated sludge [53, 57–62] and in fixed-bed processes (percolation filters: [53, 59]; rotating disc contactors: [63–65]), compare Table 4. It should be noted, however, that the composition of the protozoan biocoenosis, as well as that of the total biomass involved in the purification process, is mainly dependent on the composition of the wastewater, together with physical conditions and factors arising from the process technology used. In the case of malfunctions, or in the initial stage of a plant, very different compositions can be encountered. Sydenham [57] observed 2 municipal activated sludge plants over a period of 12 months and identified amoebae as the dominant group with regard to biomass. In sludge with a high organic load, Curds and Cockburn [66] and Mudrack and Kunst [67] report high population densities of flagellates. The age of the sludge also has an effect on the composition of the protozoan community. Kinner and Curds [63] quote 6–12 months as the length of time required to establish a steady-state community of protozoa in a pilot rotating disc contactor plant supplied with domestic effluent. Bacteria were visible on the disc surfaces within one day of startup followed within a few days by flagellates and small amoebae. Free-swimming bacterivorous ciliates appeared within 8–10 days. Subsequently, sessile peritrichous forms accompanied by carnivorous ciliates, rotatoria, and large amoebae make up the stable community. Parallel to sludge aging, a typical chronological succession of dominant protozoa populations can also be observed in activated sludge plants. After the initial phase of 1–2 weeks where flagellates, naked amoebae, and free-swim- Table 4. Structure of the protozoan community in three urban activated-sludge plants, oper- ating at different organic loading rates and dissolved oxygen concentrations (observation over a one year period), according to [62]. Biomass calculation is based on data, given by [61] Organic load a O2-conc. (mg O2/l) Densities and biomass Ciliates Flagellates (< 20 mm) Naked Amoebae (<50 mm) a kg BOD5/(kg MLVSS) day. Plant 1 Plant 2 Plant 3 0.23–0.38 3.6–5.2 ind./ml mg/l 3000– 18–43 7400 43 000– 2.2–5.2 600 000 4000– 0.21–5.3 100 000 0.21–0.35 1.8–3.0 ind./ml mg/l 8600– 50–99 17 000 89000– 4.6–51 980 000 800– 0.04– 130 000 6.9 0.5–0.8 1.0–1.3 ind./ml 4500– 16000 38000– 1 600 000 77000– 101 000 mg/l 26–93 20–83 4.1–5.4 Protozoa in Wastewater Treatment: Function and Importance 219 ming ciliates predominate, more and more crawling and sessile forms appear, which remain dominant throughout the stabilization phase and can be regarded as typical representatives of mature sludge [62, 65, 68–70]; see also Fig. 5. Unlike the free-swimming forms, which arrive at the plant with the sewage and are flushed out at the end of the process, the existence of sessile and crawling forms is closely associated with the development of slime growth or sludge Fig. 5. Composition of the bacterivorous ciliate community during the establishment of a mature sludge. Stabilization, i.e., steady-state occurs after about 50 (activated sludge, above figure) and ca. 80 days (RBC, below figure), respectively. Bars lower than 100% indicate the additional presence of carnivorous and omnivorous ciliates, after [65] 220 W. Pauli et al. flocs. Bound to biofilms as fixed slime growths (fixed-bed) or as sedimentable sludge, they are retained in the treatment plant and can thus build up a stable community with the bacterial flora. Whereas characteristic population succession takes place in both plant types, in percolation filters, due to the unequal distribution of the organic load, a physical separation of the organisms is observed, dependent on the filter depth [68]. Figure 5a, b shows results from studies on the colonization behavior of ciliates in a pilot rotating disc contactor plant as well as in an operational activated sludge plant [65]. Both plants were fed with domestic wastewater. Whereas in the initial stage of the activated sludge plant ciliates make up between 0.17% and 0.44% of the total biomass, in the stabilizing phase they account for more than 9% of the sludge biomass. In the initial phase free-swimming forms from the wastewater dominate. After 10–15 days their numbers drop markedly and crawling (Aspidisca cicada, A. lynceus, Euplotes affinis, Chilodonella uncinata) as well as sessile (Vorticella convallaria, V. microstoma, Epistylis plicatilis, Opercularia coarctata) ciliates characterize the protozoan fauna. Similarly, in the rotating disc contactor plant, ciliates makes up only 4–5% of the slime biomass in the colonization phase, as opposed to 12–19% under steady-state conditions. Here, too, essential changes take place during the colonization of the submerged contact aerator and the typical ciliate biocoenosis develops in the plant itself. In the initial phase, free-swimming ciliates such as Paramecium putrinum and Uronema nigricans are present; in the stable phase sessile forms such as Opercularia coarctata and Vorticella convallaria dominate. Investigations by Madoni [64, 65] make it clear that in both types of plants (submerged contact aerator and activated sludge) a significant positive correlation exists between the increase of the sludge, biofilm and ciliate biomass (r2 = 0.927 and r2 = 0.853). This implies a close relationship between the size of the ciliate population and the bacterial biomass. 2.2.3 Plant Specific Basic Communities The relative abundance of an organism in a particular habitat can be considered as a measure for its significance within the ecological structure of the biological system concerned. Alongside amoebae and flagellates, Curds and Cockburn [53] identified 67 and 53 ciliate species in 56 activated sludge plants and 52 percolation filter plants in Britain, respectively. Madoni and Ghetti [55] detected 45 and 47 ciliate species in 39 activated sludge plants and 49 percolation filter plants in Northern Italy. Of note is that the British and Italian activated sludge plants revealed very similar ciliate fauna [55]. Nevertheless, not all species in the individual samples can be regarded as typical, as to their presence and population density, for the respective wastewater treatment process. The majority of the species are found only sporadically in a few samples and usually with a low population density. The overall picture of the ciliate population is determined by a few, primarily sessile (peritrichous) and crawling (hypotrichous), species most of which are bacterivorous (compare with Fig. 6). With cell counts of, on average, more than 104 ml–1, ciliate densities are 100–1000 times higher Protozoa in Wastewater Treatment: Function and Importance 221 Fig. 6. Examples of free swimming (holotrichous), crawling (hypotrichous), and sessile (peritrichous) ciliates in waste water treatment plants here, than in the plankton of oligotrophic (10 ml–1) and eutrophic (100 ml–1) waters [24]. Table 5 summarizes the dominant ciliate species in the “basis community” of each plant type identified by Curds and Cockburn [53] and Madoni and Ghetti [55]. The specific biocoenosis differs according to plant type and to the current operating conditions [55]: in areas with a high organic load an increase in free swimming species is observed [61, 66] along with a decrease in the diversity of species [71, 72]; with these limitations, the community forms given in Table 5 can be considered average for municipal plants. 2.2.4 Biomass In activated sludge plants a high proportion of the eukaryotic biomass is comprised of protozoa. Investigations carried out by Sydenham [57] revealed that protozoa made up over 90% of the total eukaryotic biomass of two municipal wastewater treatment plants. According to Aescht and Foissner [61], protozoa made up 99–100% of the eukaryotes in a pharmaceutical plant with a bacterial nutrient load. The average proportion of protozoa in relation to total solids (dw) is 5% [59, 73]. Ciliates alone make up 10% of the total biomass (pro- and eukaryotic dry weight). Even higher numbers of ciliates are encountered in municipal rotating disc contactors where proportions of about 20% of the total biomass of the slime-growth can be observed [64, 65]. 2.2.5 Ecological Framework The biocoenosis in wastewater treatment plants should not be regarded as a community with a rigid composition and constant characteristics but rather as 222 W. Pauli et al. Table 5. Ciliate species dominating and occurring with a high frequency in sludge samples of British (GB) and North Italian (I) sewage treatment plants, respectively, after [53, 55]. “Dominating” refers to the relative cell density, whereas “present” indicates the number of samples, in which the respective species – independent of its individual numbers – could be observed Dominant Present (%) (%) GB I GB I Life form Ecological type Nutrition a Activated sludge (GB and I) Aspidisca costata a Vorticella convallaria a Trachelophyllum pusillum a Opercularia coarctata a Carchesium polypinum a Vorticella alba (GB) Vorticella microstoma (GB) Euplotes moebiusi (GB) Vorticella fromenteli (GB) Euplotes affinis (I) Zoothamnium pygmaeum (I) Trochilia minuta (I) 35 19 15 12 11 11 10 5 4 – – 2 Crawling Sessile Free swimming Sessile Sessile Sessile Sessile Crawling Sessile Crawling Sessile Crawling Bacterivorous Bacterivorous Carnivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Filamentous b Trickling filter (GB) Opercularia micodiscum Carchesium polypinum Vorticella convallaria Chilodonella uncinata Opercularia phryganeae Opercularia coarctata Vorticella striata Aspidisca costata Cinetochilum margaritaceum 44 15 10 4 4 2 2 – – 81 62 83 90 90 56 52 56 54 Sessile Sessile Sessile Crawling Sessile Sessile Sessile Crawling Crawling Bacterivorous Bacterivorous Bacterivorous Filamentous b Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Rotating biological contactor (I) Euplotes moebiusi Paramecium caudatum Trachelophyllum pusillum Vorticella convallaria Opercularia microdiscum Opercularia coarctata Paramecium trichium Cinetochilum margaritaceum Chilodonella cucullulus 53 46 41 53 41 33 27 23 18 79 79 59 57 45 37 43 37 41 Crawling Free swimming Free swimming Sessile Sessile Sessile Free swimming Crawling Crawling Bacterivorous Bacterivorous Carnivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Bacterivorous Filamentous b a 85 77 30 23 26 – 10 5 – 59 33 23 69 58 64 54 25 38 75 35 31 11 – 12 90 84 58 25 28 – 10 7 – 69 33 25 Dominant both in British and Italian plants. Filamentous: ciliates with a specialized oral apparatus, enabling the ingestion of rod-shaped, filamentous bacteria. – not present. b Protozoa in Wastewater Treatment: Function and Importance 223 an artificial but biological segment of natural self-purification processes, the composition of which is influenced by ecological conditions and physico-chemical factors, thus differing from plant to plant and even within a plant over time. 2.2.5.1 Sludge Loading Sludge loads with fm-values between 0.2 and 0.6 [g BOD (g MLSS · day)–1] are considered optimal for the purification sequence at conventional municipal activated sludge plants (e.g., [47, 67]). Ciliate densities of 6000–30,000 ml–1 are found in sludge with these loads [71, 74]. However, similar concentrations of ciliates are also encountered in sludge with both higher and lower loads: Salvado and Gracia [71] observed a constant ciliate population density in a municipal plant with fm-values varying from 0.03 to 0.4. Experiments by Lee et al. [74] confirm only slight changes in ciliate counts at sludge loadings between 0.1–1.4 [g BOD (g MLVSS day)–1]. Only under very heavy loads [1.8–2.4 g BOD (g MLVSS day)–1], was a reduction in cell density observed. Although the population density remains constant over a wide range, the organic load influences the number of species and the composition of dominant ciliates in the basis community. The number of species present sinks with increasing organic content of the wastewater [66, 71, 72]. According to Curds and Cockburn [66], activated sludge with a relatively low organic load [fm = 0.1–0.3 g BOD (g MLSS day)–1] shows the greatest species diversification, whereby all three groups of ciliates – peritrichs (sessile), hypotrichs (crawling), and holotrichs (free swimming) – are represented with approximately the same number of species. In the medium load range of fm = 0.3–0.6, peritrichous species dominate and by high organic loads of fm = 0.6–0.9 equal portions of peritrichs and holotrichs are present (Fig. 7). 2.2.5.2 Temperature Temperatures in municipal plants are generally slightly above the outside temperature in winter and slightly below in summer. Performance is optimal between 10 °C and 25 °C [41]. No negative effects on ciliate fauna are found up to 30 °C; experimental activated sludge investigations reveal a decline in ciliates at temperatures above 30 °C and their disappearance above 40 °C [74]. The authors discuss the concomitant deterioration of the settling properties of the sludge as possibly resulting from the collapse of the ciliate population. 2.2.5.3 pH-Value Activated sludge has a relatively high buffer capacity. If no strongly acidic or alkaline effluents are introduced, mainly from industrial processes, pH-values generally fluctuate between 6.5 and 8 [41, 67, 75]. Therefore not only the tem- 224 W. Pauli et al. Fig. 7. Composition and species number of ciliates in activated sludge plants operated at different sludge loadings [food to micro-organism (F/M) ratio]. Results from an investigation of 52 British plants made by Curds and Cockburn [53]. Peritrichous, hypotrichous, and holotrichous ciliates represent sessile, crawling, and free swimming ciliates, respectively. In conventional municipal plants, treating domestic wastewater, a sludge loading between 0.2 and 0.6 is regarded to be optimum for the functioning of the sewage treatment process perature, but also the pH-values of municipal plants are in a favorable range for protozoan growth [75]. 2.2.5.4 O2-Content Conventional processes of biological wastewater treatment utilize the metabolism of the organic load, which is faster, more thorough, and easier to control under aerobic conditions. Aerobic conditions are also a prerequisite for a high incidence of protozoa. Few specialists can survive strictly anaerobic conditions and little knowledge is available regarding their distribution or function in anaerobic degradation processes. The number of facultative anaerobic protozoa is slightly higher, but almost all species seem to be able to survive low oxygen concentrations or even the absence of oxygen, at least for a short period [75]. Apart from plant malfunctions (e.g., breakdown of the aeration), this ability is also important in the normal cycle of activated sludge processes, where the organisms are constantly alternating between the aerobic activated sludge tanks and the sedimentation tanks, in which anaerobic conditions arise for short periods Protozoa in Wastewater Treatment: Function and Importance 225 in the deeper layers of the settling sludge (less than 4 h [41]). Only longer and repeated oxygen deprivation over several hours (continual alternation between 6 h aerated and 24 h without aeration) leads to a marked decline of the sessile ciliates Vorticella convallaria and Opercularia coarctata, typically found in wastewater treatment plants [76]. 2.3 Significance of Protozoa for Wastewater Treatment As already described (Sect. 2.2.2), the majority of protozoa in aerobic biological purification systems are sessile or crawling ciliates. Whereas free-swimming ciliates are flushed out with the clarified water, crawling and especially sessile forms are bound to bacterial biofilms (flocs and slime growths) [59]. In the case of fixed-bed plants they remain bound to the biofilms in the plant; in activated sludge processes they sediment with the sludge and are retained in the plant due to continual sludge recycling. To understand the role of protozoa and classify their position in the artificial system of biological wastewater treatment, the following characteristics have to be considered: type of motion (free swimming, crawling, or sessile); form of nutrition (e.g., filter-feeders, browsers); sources of nutrition (abiotic colloids and particles, bacteria, algae, other protozoa). From their form of nutrition and their trophic level, functional aspects important for wastewater treatment become apparent. New understanding of natural systems as well as experimental results on the physiology, energy budget, and nutrient cycling of both aquatic and terrestrial protozoa provide extensive information regarding the ecological role of this group of organisms, which, although quantitatively less significant than bacteria, make a considerable contribution to wastewater treatment. 2.3.1 Nutrition Several possibilities are open to ciliates for nutrient-uptake. On the one hand, similar to bacteria, substances can be transferred directly through the plasma membrane into the interior of the cell. Active and passive, carrier-mediated uptake mechanisms through the plasma membrane have been described for Tetrahymena for amino-acids [77–79], di-peptides [80], acetate, glucose [81, 82], and even for such complex nutrient solutions as proteose-peptone-yeast extract (PPY) medium [83]. Another method of nutrient uptake is pinocytosis [84, 85]. It describes the active transport of dissolved substances in sub-microscopic, particle-free vacuoles or vesicles from the plasma membrane to the cell interior, where they undergo normal lysosomal digestion processes. Finally, ciliates have a highly specialized oral apparatus for taking up particulate matter by phagocytosis. The particles are not simply ingested with the surrounding solution but rather undergo a highly efficient filtration process, facilitating the concentration of particulate matter from a large volume of liquid, prior to their intake in food vacuoles [85]. This process involves the production of a water current by cilia (Fig. 8) and the extraction of particles from the flowing water 226 W. Pauli et al. Fig. 8. Mechanisms of filter-feeding (ambiguously often referred to as ‘grazing’) used by protozoa. Water currents are created by flagella or the coordinated activity of cilia, that bring suspended food to the mouth region of the cell with the aid of a ciliary sieve, which retains – in the case of bacterivorous species – particles sized between 0.3 mm and 5 mm [85–87]. The particles, thus concentrated, are subsequently ingested. Apart from food, abiotic and even indigestible matter of the size of bacteria are efficiently ingested [86–89]. Paramecia concentrate food particles in this manner in their oral cavity up to 1000-fold [90]. A similarly high concentration capacity can be assumed for Tetrahymena: Whereas a volume of 50–80 nl is cleared of particles per hour and cell [87, 91], a more than 1000 times lower water volume of 36 pl h–1 and cell is actually ingested by the food vacuoles [92]. The efficiency of this form of nutrition is underlined by investigations comparing the growth kinetics of Tetrahymena pyriformis with particulate and dissolved substances as nutrient source, respectively [93]. While under monoxenic conditions with particulate bacterial substrate the half maximum growth rate is already attained with a bacteria content of 12 mg l–1 Klebsiella aerogenes (5.5 mg carbon l–1), 200 times that concentration of organic matter is required in case of dissolved nutrients (2.4 g l–1 proteose-peptone-yeast medium = 1.3 g carbon l–1). Protozoa in Wastewater Treatment: Function and Importance 227 2.3.2 Reduction and Elimination of Suspended Particles and Bacteria 2.3.2.1 Clearing Rate The volume of water cleared per individual and hour depends on cell size. Small protozoa with cell diameters of less than 5 mm, such as flagellates, filter less than 1 nl h–1 at temperatures between 9°C and 17°C [20]. Higher filtration rates are observed for larger ciliates. Sanders et al. [20] quote a yearly fluctuation range of 12–156 nl h–1 for the filtration performance of planktonic ciliates. In laboratory experiments with the ciliates Halteria grandinella (diameter: 25 mm) and Strombidium sp. (size: 15 ¥ 21 mm) filtration rates of 80–90 nl h–1 at 9°C and 120–140 nl h–1 at 17°C were determined. In the case of Vorticella microstoma (average cell dimensions: 60 ¥ 30 mm), a ciliate frequently present in wastewater treatment plants, filtration rates as high as 156 nl h–1 at bacteria densities of 106 ml–1 are reported. Tetrahymena (cell dimensions: 40 ¥ 20 mm), a species present but not dominant in wastewater treatment plants, has a filtration performance of 80 nl h–1 [91]. Fenchel [87] observed filtration rates of 50 nl h–1 and cell at 20–22°C for Tetrahymena pyriformis and 200–1000 nl h–1 for larger (100–200 mm) representatives of crawling and free-swimming ciliates such as Euplotes, Paramecium, or Blepharisma. Assuming average filtration rates of 100 nl h–1 and cell and ciliate densities of 10,000 ml–1 and above [61, 94–97], this implies that the entire liquid of an activated sludge plant can be filtered in less than 1 h. The enormous predator and selection pressure exerted on the bacteria is illustrated by the following examples. Many heterotrophic bacteria in activated sludge have the ability to divide every 20–40 min under optimal laboratory conditions [41, 48]. Under “field” conditions, such as those prevailing in wastewater treatment plants, their growth is generally much slower due to sub-optimal physical (temperature) and physiological (nutrients, pH-values) parameters. The actual bacterial division rates under constant operating conditions and good nutrient availability can be estimated from the ratio of the surplus (drawn off) sludge to the total sludge in the activated sludge plant [98]. For low to high organic loads (fm = 0.05–0.6 g BOD per g MLSS and day), growth rates can vary from 4–50% per day [41] or, expressed in other terms, the bacteria population in the sludge doubles every 48 h at most, i.e., in a time span by no means adequate to compensate for potential protozoan feeding. Highly loaded wastewater contains ca. 106 bacteria ml–1. The majority are medically harmless but others are pathogenic and bear health risks. Conventional wastewater purification involves an initial pre-clarification step of 20–30 min, after which the wastewater is fed into the activated sludge tank and aerated for 4 h. In the aerated and agitated system of the activated sludge tank the wastewater is brought into contact with a mixed microbial population in the form of a flocculent suspension. When the desired degree of treatment has been achieved, the flocculent microbial mass, known as the “sludge”, is separated for 2–4 h from the treated wastewater in a separate, specifically designed sedimen- 228 W. Pauli et al. tation tank. The supernatant from the separation stage is the treated wastewater, and should be virtually free of sludge. Most of the settled sludge from the separation stage is returned to the aeration stage to maintain the sludge concentration in the aeration tank at the level needed for effective treatment and to act as a microbial inoculum. Some of the sludge is removed for disposal, and is known as “waste” or “surplus” sludge. In both the activated sludge and the sedimentation tanks, the resident ciliate community has sufficient time to filter the entire wastewater several times, thus removing bacteria and abiotic particles of similar size (see Sects. 2.3.1 and 2.3.2). 2.3.2.2 Experimental Findings It has long been known that protozoa are present in wastewater treatment plants and that their species composition reflects the prevailing conditions in the plant. However, scientific opinion was less unanimous with regard to the actual contribution of protozoa to the purification process. Although Ardern and Lockett [99], Pillay and Subrahmanyan [100], Pillay et al. [101] and McKinney and Gram [102] referred to a connection between protozoa and the quality of the water discharged from the plant, proof of a causal relationship was lacking or inconclusive. Curds et al. [103] succeeded in selectively removing protozoa from activated sludge and further cultivating this protozoan-free sludge in bench-scale treatment plants over a long period. Through the subsequent re-introduction of typical sludge ciliates they observed, under various starting conditions, positive effects on a series of parameters describing the success of the purification process (Table 6). The principal observation of their experiments was that in the absence of protozoa the effluent of the plant was turbid, due to its high content of suspended bacteria; this turbidity almost disappears after re-introduction of the protozoa (Fig. 9). Similar findings are published by Sridhar and Pillai [104] and Macek [105] in protozoan-free, pasteurized sludge and in bacteria cultures isolated from activated sludge. The addition of sessile, crawling, and free-swimming ciliates Table 6. Effects of ciliated protozoa on the effluent quality of bench-scale activated-sludge plants. Results are given in mg l–1 unless otherwise noted; after [103] Effluent analysis Without ciliates With ciliates Mean reduction BOD COD Permanganate value (4 h) BOD after filtration COD after filtration Organic nitrogen Suspended solids Optical density at 620 nm Viable bacteria counts (106 ml–1) 53–70 198–250 83–106 30–35 31–50 14–21 86–118 0.95–1.42 160 7–24 134–142 62–70 3–9 14–25 7–10 26–34 0.23–0.34 1–9 75% 38% 30% 81% 39% 51% 71% 76% 97% Protozoa in Wastewater Treatment: Function and Importance 229 bacterial density Fig. 9. Influence of ciliates on the bacteria content in the effluent of a bench-scale activated sludge plant, after [103] (Epistylis articulata, Vorticella microstoma, Aspidisca cicada, Chilodonella uncinata, Stylonychia putrina, Colpidium camylum) reduces high COD values and suspended matter content. Farrah et al. [106] confirm the causal relationship between the presence of ciliates and a clear, almost bacteria-free effluent with a low organic content. Departing from a typical pro- and eukaryotic sludge biocoenosis, the authors show that a largely selective reduction of protozoa, by the addition of sodium fluoride (0.2 mol l–1) or sodium azide (6–20 mmol l–1) results in a notably higher content of freely suspended bacteria including streptococci. After application of the eukaryotic cell toxins, the total count of fecal streptococci increases about threefold and the proportion of suspended bacteria, as compared to those bound to flocs, increases from 0.3% to 64%. Kakiichi et al. [107] made essentially the same observations. The effects of two amphoteric detergents (orthodichlorobenzene and polyhexamethylene biguanide hydrochloride) with known effects on bacteria and protozoa were studied and a causal relationship between poor quality of the outflow (increased turbidity and COD values) from batch cultures of activated sludge and the inhibitory effect (reduction in population density) on the protozoa was observed. A correlation between the effluent quality and the population density of protozoa is also implied by Lee et al. [74]. Studies with a bench-scale activated sludge plant (organic load: 0.1–0.4 g BOD per g MLSS and day) show that the selective decline of the ciliate population density, due to running temperatures of 36°C and over (see Sect. 2.2.5.2), corresponds to a more than twofold increase in suspended matter in the effluent. Experiments with bacteria-free synthetic wastewater (e.g., [103]) exhibit that freely suspended bacteria, originating from the autochthonous microflora of the activated sludge itself, are substantially reduced in the presence of protozoa. 230 W. Pauli et al. Furthermore Curds and Fey [108] observed that bacteria originating from the influent wastewater are also effectively removed in the presence of protozoa. After mechanical destruction of the protozoan population, by means of a ball mill, the authors determined concentrations of 6.5 ¥ 105 culturable E. coli ml–1 in the effluent of a continuously operating bench-scale activated sludge plant; after re-inoculation of the activated sludge with ciliates (Opercularia coarctata, Vorticella microstoma, Hypotrichidium conicum, Tetrahymena pyriformis) and the establishment of a stable protozoan community, this count was reduced tenfold to 6.3 ¥ 104 ml–1. The half life of E. coli in the activated sludge was reduced from 16 h to 1.8 h. Filter-feeding ciliates in wastewater treatment plants are, in principal, not selective consumers. Along with harmless bacteria, a series of pathogenic strains causing, for example, diphtheria, cholera, typhoid, and streptococcal infections are also phagocytosed (for reviews [75, 109]. Investigations by Farrah et al. [106], with activated sludge in batch cultures, illustrate the significance of this elimination of pathogenic bacteria from wastewater treatment plants. After selective reduction of the protozoan fauna by sodium fluoride (200 mmol l–1) or sodium azide (20 mmol l–1), cultures of Salmonella typhimurium and E. coli, added in densities of 105 ml–1 almost treble within 24 h (S. typhimurium and sodium fluoride) or only decrease by ca. 50% (E. coli and sodium azide), whereas in untreated controls with protozoa, both bacteria are reduced to less than 5% of their initial density. Moreover, under conditions of aerobic sludge stabilization, the authors show that even low densities of protozoa (660 ml–1) lead to a substantial elimination of bacteria. Figure 10 shows results with Streptococcus faecalis. Fig. 10. Effect of sodium azide (6 mmol l–1, selectively reducing protozoan activity) on Streptococcus fecalis in activated sludge (laboratory scale), after [106]. CFU: colony forming units Protozoa in Wastewater Treatment: Function and Importance 231 2.3.2.3 “Field”-Observations “Field” observations leave no doubt that the results found in the laboratory microcosms are transferable to pilot and full-scale plants and that the presence of a typical protozoan community is reflected by the improved quality of the plant effluent. First, a close negative correlation is observed between the population density of, mainly crawling and sessile, ciliate populations and the proportion of suspended matter in the effluent of wastewater treatment plants (domestic and municipal wastewater [56, 110–114] and brewery wastewater [115]). Results from a three-year investigation of three activated sludge plants with different organic loads in Spain [114] reveal – on average for all plants – a highly significant correlation coefficient between total ciliate population density and biological oxygen demand of r = – 0.868. In the presence of protozoa the effluent BOD ranges from 4 mg l–1 to 18 mg l–1, rising to values of up to 67 mg l–1 in their absence. An almost identical correlation between effluent quality (COD) and the population density of typical activated sludge ciliates was observed by Sudo and Aiba [111] for six municipal wastewater treatment plants in Tokyo. Mean COD values of 10 mg l–1 were found with ciliate densities of ca. 104 ml–1; these increase to 40 mg l–1 when ciliate densities drop to 102 ml–1 (Fig. 11). Second, according to Curds and Cockburn [53], plants without ciliates can be recognized by the low quality of their effluent: 3 out of 53 activated sludge plants were selected due to the high content of suspended material in their effluent. In one of these plants no protozoa could be found at all, in the other two no ciliates, only small flagellates, could be detected. The highest BOD values measured in the three plants occurred in the plant with no protozoa. Fig. 11. Relationship between protozoan densities and effluent COD, observed in municipal activated sludge plants of Tokyo, after [111]. Symbols represent different plants 232 W. Pauli et al. Finally, as compared to normal activated sludge processes, the clarifying effect of protozoa in activated sludge processes with submerged fixed-bed filters – a technology which creates additional surfaces for slime growth and primarily sessile ciliates [116–118] – improves, which is basically due to low bacteria and suspended matter content in the fixed-bed plant effluent [117, 119]. (Evidently, protozoa find optimum living conditions on the filter installed in the activated sludge tank, an adequate oxygen supply and plenty of food, so that the dense population of mainly ciliates even crowds out attached bacterial growths. In contrast to the common activated sludge process, where ciliates contribute to about 10% of the total, bacteria dominated biomass, an almost inverse relation of 68% protozoan and 32% bacterial biomass (dw) is found for the biofilms of submerged fixed-bed filters [116].) 2.3.3 Elimination of Dissolved Substances The bulk of dissolved substances entering the wastewater treatment plant are amino-acids, products of protein hydrolysis in the sewage system, and fatty acids. Carbohydrates are usually completely degraded in the sewage before reaching the plant. Although many protozoa can take up organic substances [85, 89, 120, 121], their contribution to the degradation of these substances in wastewater treatment plants is negligible: For these substances the essential activity comes from the bacteria population. They dominate the biomass and possess a higher metabolic efficiency as a result of their high surface to volume ratio [41, 46–48, 67]. An impression of the different degradation efficiencies can be gathered from measurements of amino-acid uptake by Escherichia coli and T. pyriformis [122]. Even under the assumption that all ciliates present in wastewater treatment plants can metabolize not only bacteria but also dissolved substances similar to T. pyriformis, the experiments reveal an 80-times higher uptake of amino-acids by bacteria. Results from Hrudey [123] can also be well interpreted in the light of the significantly higher degradation rate of dissolved substances by bacteria. After addition of peptone, a protein hydrolysate rich in amino-acids, an immediate rise in the bacterial biomass was observed, whereas ciliates were scarcely able to convert the available peptone into their own biomass and could only reproduce substantially after the bacterial content increased considerably. 2.3.4 Flocculation and Composition of the Bacterial Community Apart from the feeding activity of protozoa, another factor is discussed as contributing to the reduction of the content of suspended matter and bacteria in bench and full-scale plants. In the presence of protozoa, freely suspended, single bacteria form compact flocs, which then settle [59, 105, 106, 111, 124–128]. This is attributed, on the one hand, to polymer, particle-aggregating excretion products (polysaccharides) from protozoa [59, 125], which are possibly released into the media to facilitate a more effective uptake of particles [24, 129]. Protozoa in Wastewater Treatment: Function and Importance 233 On the other hand, this flocculation is believed to be associated with the exocytosis of indigestible, originally finely dispersed material as a digested bundle [130, 131], which in turn could serve as a settlement surface for solitary bacteria [132–134]. However, wastewater itself contains a high proportion of chemically complex particles of differing sizes, and bacteria themselves, dominant with regard to their biomass in wastewater treatment plants, produce extracellular polymeric substances (polysaccharides), to which they can effectively adsorb [135, 136]. For these reasons, protozoa, by excretion of digested remains and polymers, probably play only a minor role in floc formation in wastewater treatment plants. Bacteria feeding itself seems, not only quantitatively but also qualitatively, a significant stimulus for complex bacterial growth forms. As a result of the predator-prey relationship between protozoa and bacteria, a collapse of the bacteria population in the activated sludge and a reduced elimination efficiency of the system as a whole would be expected (see Sects. 2.1.2 and 2.3.3). Such collapses or phase-shifted oscillations between predator and prey can be observed in model systems [111, 137–141] and in natural ecosystems [20, 142–146] and led originally to the view that protozoa are harmful for the clarification process [147]. Only a few protozoa, e.g., amoebae, mostly present at low densities in wastewater treatment plants, are principally capable of taking up larger particles, due to their ability to entrap their prey. Ciliates, typical representatives of protozoa found in wastewater treatment processes, possess a highly specialized oral apparatus for highly efficient filtration, which at the same time exclude particles of several micrometers in diameter [86]. The ability of bacteria to develop larger forms, to grow collectively, or to merge as micro-colonies protects them against the predator pressure from the protozoa [148–153]. The development of growth forms resistant to filter-feeding can thus be seen as an essential process in the evolution of bacterial flocs and biofilms [45, 111, 127, 148]. To what extent a qualitative selection of floc and biofilm forming bacteria is possible [148], and what could be gained from a quantitative shift within a species to larger or aggregating phenotypes [45, 149], cannot be decided in the light of the present literature. Güde [148] observed selection of bacteria populations which aggregate in pilot wastewater treatment plants. On the other hand Shikano et al. [149] find that, in the presence of the ciliate Cyclidium sp., phenotypes of considerably larger dimensions appear within a bacteria species. Gurijala and Alexander [154] provide evidence of lower feeding pressure by the ciliate Tetrahymena thermophila on bacteria with hydrophobic surfaces – in other words on phenotypes with water-repellent properties – which enhance their adhesive, i.e., aggregation, ability [155]. Many bacteria are also capable of organizing themselves spontaneously into biofilms in the absence of protozoa, thus forming flocs [102, 135, 156]. Nevertheless, the extent and persistence of the flocs seem to be influenced by the presence of protozoa. Farrah et al. [106] show that in the absence of protozoa autochthonous aerobic bacteria and cultures of Salmonella typhimurium and E. coli introduced into the sewage sludge are predominantly freely suspended (43–68%). In the presence of protozoa, the proportion of freely suspended 234 W. Pauli et al. bacteria drops significantly to 1–15%. The majority can now be found in or adhering to flocs (85–99%); compare also Fig. 10. Experiments in model wastewater treatment plants [105] show that different ciliate species induce flocculation to different degrees. With the exception of the crawling Aspidisca costata, which, even at low population densities, induces good flocculation when added to protozoan free (pasteurized: 50°C, 5 min) sewage sludge, the tendency of bacteria to aggregate in the laboratory fermenter varies considerably for free-swimming (Colpidium campylum), crawling (Chilodonella uncinata, Stylonichiaputrina), and sessile (Vorticella microstoma) forms, essentially independent of their population density. Ciliates feed selectively, not only – as shown for Tetrahymena – with regard to the physico-chemical surface structure of their prey [154], but also regarding the size of the phagocytosed particles: This was shown by Fenchel [86] with filter-feeding ciliates, characteristic for the ciliate fauna in wastewater treatment plants. Each ciliate species can only filter specific size ranges of food particles, i.e., different ciliate species feed in their respective – sometimes distinct – niches (Fig. 12). Dependent on the selection mechanism, different effects on the composition of the bacterial populations and the development of more or less aggregated growth forms become apparent. Fig. 12. Clearing rate (volume of water the organisms can clear of particles per unit time at low particle concentrations, here in multiples of the ciliates own volume per h) for three ciliate species as function of particle size, from [86] 235 Protozoa in Wastewater Treatment: Function and Importance 2.3.5 Reduction of the Total Biomass In order not to exceed a sludge concentration favorable for the purification performance of the wastewater treatment plant, an amount equal to the daily production must continuously be drawn off. This excess sludge is subsequently concentrated, digested, and drained and must finally be disposed of as a potentially pathogenic and frequently toxic waste product. Excess sludge is therefore an economic factor, even within the wastewater treatment plant itself. A reduction in sludge production corresponds to savings in personnel, energy, and running costs. Since the function of the sedimentation tank is merely to separate the biomass from the purified water, the effluent concentration must already be attained in the well-mixed activated sludge tank. The organisms therefore live in an environment with low nutrient concentrations, resulting in slow growth [46, 67]. The average age of activated sludge (sludge residence time) for organically burdened municipal sewage, where the main emphasis is on the elimination of the carbon compounds, is 4 days [41, 46]. If nitrification is an objective, the sludge residence time increases to 8–10 days [157]. This means that the sludge biomass doubles after 4 days, at the earliest. Generation times in this range imply not only stationary growth for the majority of heterotrophic bacteria but also sub-optimal, reduced growth rates for the ciliate fauna having generation times of 5–15 h; see Table 7. In principal, a lengthening of the food chain results in a reduction of the originally available energy. In every heterotrophic link, part of the assimilated food is converted into biomass. The remaining carbon compounds are used as energy source for metabolic processes. When the chain becomes longer, less energy will remain locked into biomass. This means more carbon-mineralization and less biomass production. Table 7. Doubling times of activated sludge and ciliates isolated from activated sludge plants Activated sludge a Aspidisca costata b Aspidisca lynceus b Vorticella microstoma b Vorticella convallaria b Carchesium polypinum b Opercularia spec b Epistylis plicatilis b Colpidium campylum b Tetrahymena pyriformis b Paramecium caudatum b a b [158]. [111]. Doubling time (h) Temperature ( °C) 3.3–10 13.6 12.4 5.0 7.6 9.3 5.0 10.2 4.7 4.5 12.0 20 20 20 20 20 20 20 20 20 20 20 236 W. Pauli et al. Protozoa assimilate about 85% of readily exploitable nutrients after uptake. They are converted into individual biomass or respired for energy purposes. The remaining 15% are eliminated as compact digestion bundles (exocytosis) or dissolved substances (excretion) [159]. Under optimal growth conditions, ca. 50% of the nutrients taken up by protozoa are converted into individual biomass, which corresponds to the metabolic efficiency of prokaryotes [160]. Different circumstances are encountered under inhibited or stationary growth conditions. Here the emphasis is on basal metabolism, not growth: The metabolic performance is reduced and energy consumption, as mineralized carbon in the form of CO2, increases [128, 161]. This diminished ability to utilize available nutrients for biomass production as a result of reduced growth rates is demonstrated by Ratsak et al. [128] with Tetrahymena pyriformis. At a high growth rate (generation time of 5.5 h near the optimum of 3.4 h), 51% of phagocytosed bacterial biomass (Pseudomonas fluorescens) are converted into ciliate biomass, whereas at a low growth rate with a generation time of 17 h only 39% of the prey is converted into predator biomass. At the same time the ratio of respired mineralized carbon to that converted into cell biomass increases from 0.65 to 1.2. In municipal activated sludge plants ciliates are present in densities of 104 ml–1 and over [61, 75, 94–97]. The number of bacteria required to maintain this ciliate population can be estimated based on data from Macek [162]. Under steady-state conditions (20°C) and generation times of 5 days, free-swimming ciliates such as Colpidium campylum and sessile forms such as Vorticella microstoma at densities of 1.3 ¥ 104 ml–1 and 0.59 ¥ 104 ml–1 consume, over the 5-day period, 2.5 ¥ 109 and 2.1 · 109 bacteria ml–1 (450 and 420 mg COD l–1), respectively. The bacterial content of sewage arriving at the plant is on average 106 ml–1. With flow-through times of 2 h or more in municipal activated sludge plants [41], no more than 0.5 ¥ 106 bacteria are available per ml and hour for the ciliates. Based on the findings of Macek [162], however, a typical ciliate density of 104 ml–1 would require more than 17 ¥ 106 bacteria ml–1 and hour (2–3 ¥ 105 ml–1 in 5 days). Therefore, the suspended bacterial content in the influent sewage cannot essentially contribute to the production of protozoan biomass. To supply adequately the protozoan population a 30-times higher bacterial content in the influent would be required. It is known that bacterivorous species are capable of effective filtration and ingestion of abiotic particles with diameters of 0.3–5 mm [85–87; see also Sect. 2.3.1] and exploiting them, if possible, for cell reproduction or to increase individual biomass. Thus in bench-scale plants, the addition of emulsified lipids, which form suspended particulate fat droplets, leads to a rapid increase in sessile ciliates, which can accumulate these lipids in their cytoplasm [123]. Similarly, Tetrahymena is able to convert particulate suspended skimmed-milk for reproduction, thereby attaining high population densities [163, 164]. It is unclear however to what extent particulate abiotic organic materials (e.g., protein rich colloids from feces) in municipal sewage are suitable, in terms of chemical composition, size, and content, to be utilized in the biomass production of typical sewage plant protozoan fauna. The composition of the ciliate community in wastewater treatment plants is primarily made up of bacterivorous filter-feeding organisms which efficiently Protozoa in Wastewater Treatment: Function and Importance 237 concentrate and ingest particulate matter the size of bacteria from the surrounding liquid (see Sects. 2.2.3 and 2.3.1). Bacteria occur both in activated sludge and fixed-bed processes as complex, aggregated cell formations (flocs and slime growth). Firmly embedded in these structures, they are protected against their protozoan predators. However, there is a dynamic equilibrium between flocculation and de-flocculation (see Sects. 2.1.3 and 2.3.4) which, in the presence of protozoa, shifts towards more complex micro-colonies and, in their absence, leads to high concentrations of single suspended bacteria (see Sects. 2.3.2 and 2.3.4). That ciliates indeed can exploit the micro-flora of the sludge itself as their primary source of nutrition is confirmed by experiments with sterile synthetic wastewaters, e.g., [103, 123]. Activated sludge with an almost exclusively bacterial biomass was supplied with sterile synthetic wastewater as nutrient source (see Sect. 2.3.3) and nonetheless, a typical protozoan biocoenosis is developing. The average sludge concentration at municipal plants is quoted as 2–3 g (dw) l–1 [46], which corresponds to ca. 6 ¥ 109 bacteria ml–1 [48]. In conventional plants this bacterial mass is reproduced in 4 or more days (sludge residence time). Referring to data from Macek [162], typical ciliate populations in activated sludge consume 1.5–2.9 ¥ 109 bacteria ml–1. In other words, even at shorter retention times in a plant aimed primarily at the elimination of carbon compounds, a considerable proportion (25–48%) of the bacteria can be phagocytosed by ciliates: This corresponds to a 10–19% reduction of the accumulated sludge, based on a mineralization of around 40% of the bacterial food [128]. Observations with submerged fixed-bed filters in activated sludge plants reveal a similar picture with regard to the reduction of the accumulated sludge by protozoa. In the activated sludge tank (volume 756 m3) a contact aerator, whose slime-growth makes up almost 18% of the biomass (dw) of the tank, leads to a reduction of the BOD sludge accumulation of about 25% [117]. Such submerged fixed-beds are primarily colonized by protozoa whereby ciliates dominate [116, 117, 165, 166], comprising around 68% of the total biomass [116]. Based on these data, an additional biomass of 12% consisting exclusively of ciliates (18% additional biomass, 68% of it ciliates) effects a sludge reduction of 25%. A transfer of these results to conventional activated sludge plants would mean that the autochthonous ciliate fauna, as the second link in the food chain and representing 9% (dw) of the total biomass [64, 65], is in a position to reduce sludge accumulation by 19%. 2.3.6 Influence of Protozoa on Bacterial Metabolism A series of studies on degradation efficiency in bench-scale wastewater treatment plants show that in the presence of protozoa – in spite of their antagonistic effects as bacteria predators – the physiological performance of the bacteria is maintained or even increased: In bench-scale plants, ciliates show no effects on the nitrification bound to flocs [103, 126, 127]. The degradation of nitrilotriacetic acid by bacteria is equally unaffected by the presence of ciliates; however, here a shift from single suspended to complex aggregate growth forms is observed [126, 127]. 238 W. Pauli et al. Clear indications of an increase in bacterial metabolic activity were found by Curds et al. [103]: Under experimental conditions they observed, in the presence of protozoa, an increased degradation (BOD, COD) of the dissolved, nonfilterable portion of organic materials, attributed almost exclusively to the activated sludge flora (see Sect. 2.3.3 and Table 6). Findings by Wiggins and Alexander [167] also imply a positive influence of protozoa on bacterial degradation processes with regard to the organic pollutants 2,4-dichlorophenol (2,4DCP) and 2,4-dichlorophenoxyacetic acid (2,4-D). Although protozoan feeding reduced the mixed culture of freely suspended bacteria by more than one order of magnitude – leading to delayed degradation compared to protozoa-free cultures – after 15 days the environmental chemicals 2,4-DCP and 2,4-D were mineralized in the presence of protozoa to 70% and 90%, respectively: Whereas in cultures where the protozoa were inhibited by nystatin and cycloheximide, degradation of only 40% (2,4-DCP) and 10% (2,4-D) were observed over the same period (Fig. 13). In biocoenoses other than wastewater, i.e., in microcosms with pure and mixed cultures of typical aquatic and terrestrial bacteria, an increased bacterial metabolism in the presence of protozoa is observed. Various explanatory attempts emphasize the direct influence of the protozoan metabolic activity; others attach more importance to bacteria feeding and its indirect consequences on the size and composition of bacteria populations and some correlate the micro-currents, generated by the ciliates, with an improved food and oxygen supply of the bacterial flocs or multi-layer biofilms. Protozoa are capable of metabolizing bacterial metabolic products such as acetic-acid, butyric acid, and ethanol [77, 168] and could thus avert end-product inhibition [169]. On the Fig. 13. Effects of protozoa on the mineralization of 0.1 mg l–1 2,4-dichlorophenol and 2,4- dichlorophenoxyacetic acid (2,4-D) in sewage. Cycloheximide (250 mg l–1) and nystatin (30 mg l–1) were used to suppress protozoa; from [167] Protozoa in Wastewater Treatment: Function and Importance 239 other hand, protozoa release a series of organic substances such as amino-acids [170] and “growth factors,” having chemical structures not characterized in detail [22, 109, 171–175], into the surrounding medium, leading to activation of bacterial metabolism and growth. Furthermore, protozoa have the highest excretion rate of inorganic phosphate and nitrogen, relative to biomass, within the zooplankton [176]. In addition, in the presence of protozoa, an accelerated bacterial phosphorus mineralization is observed [177]. This mutually advantageous interaction by nitrogen and phosphorus re-mineralization is emphasized by many authors [22, 145, 177–183]. To what extent these additional organic and inorganic substances introduced into the wastewater cycle play a part in wastewater treatment processes is controversial, but due to the composition of the wastewater, rather unlikely [75, 184]. On the one hand, municipal sewage itself is a complex nutrient solution with a heavy organic load; on the other hand, nitrogen and phosphorus are present in excess in wastewater treatment plants, in contrast to most limnic, marine, and terrestrial ecosystems (a BOD:N:P ratio of 100:5:1 is considered to be the optimal substrate composition of sewage – compared to this nutrient balance, municipal sewage with average BOD:N:P ratios of 100:17:5 [41] contains an excess of nitrogen and phosphorus). However, in the case of commercial and industrial wastewaters with high carbon loading and comparatively low concentrations of nitrogen and phosphorus (e.g., vegetable processing businesses, fiberboard works, paper and cardboard factories, coking plants, as well as chemical and pharmaceutical industries [41, 185]) catalytic effects on bacterial metabolism by interactions with N and P set free by protozoa are quite conceivable. It is not self-evident that bacteria feeding and their subsequent reduction of bacterial populations should have positive effects on bacterial metabolic turnover. A possible cause could be the qualitative shifting of the selection conditions for the bacteria and therefore the composition of mixed bacteria populations and their organizational forms. The success of a bacteria population is not only dependent on its adaptation to the nutrients on offer but also on whether it is edible for protozoa [148]. As discussed in Sect. 2.3.4, the selection of feeding-resistant bacterial growth forms can be viewed mainly as a result of phagocytic activity of protozoa: Freely suspended bacteria are succeeded by aggregated, sessile growth forms [136, 148, 186]. That this shift can be accompanied by a simultaneous intensification of the microbial metabolic processes is shown by studies on marine bacteria, which as adherent cells in biofilms (“marine snow”) display faster growth (incorporation of thymidin into DNA [187]), an increase in electron transport (reduction of tetrazolium salts to formazan [188]), and higher hydrolytic activity [189], than as freely suspended single cells. 2.3.7 Filamentous Bacteria and Protozoa Filamentous bacteria are present in the bacterial flora of almost all activated sludge. Due to their large surface area, they are well-equipped for the adsorption and metabolism of organic compounds. At low densities, they contribute to 240 W. Pauli et al. the stabilization of activated sludge flocs. However processing problems arise if mass reproduction of filamentous bacteria occurs in the activated sludge tank. The enlarged surface area of the flocs hinders the settling and thickening processes in the sedimentation tank which can, in extreme cases, due to the formation of light, fluffy, poorly settling flocs, result in the discharge of sludge into natural waters. This phenomenon, known as “bulking sludge,” used to be caused by “high load bacteria” such as Sphaerotilus sp. or filamentous types 1863 and 0961. Today, however, many of the filamentous bacteria found in sewage treatment are adjusted to low carbon concentrations [low F/M ( food:micro-organism) bulking], e.g., types 0041, 0675, 0092, 1851, or Microthrix parvicella. Experience shows that putrid wastewater, rich in H2S or with high carbohydrate or short-chain organic acid content (i.e., wastewater from food processing, paper and textile industries), as well as low nitrogen, phosphorus, or oxygen con- Fig. 14. Degeneration of bulking sludge (decrease of sludge volume index: SVI) in the aera- tion tank of an activated sludge plant and in laboratory experiment by the filamentous predacious protozoan Trochilioides recta; from [194] Protozoa in Wastewater Treatment: Function and Importance 241 tent, stimulates the development of bulking sludge. Various chemical (e.g., liming, chlorination, addition of H2O2 , iron salts, and nitrogen and phosphorus compounds) and physical (e.g., increased oxygen supply) methods are implemented to combat bulking. Sometimes even operational conditions of plants are altered (e.g., increasing the return-flow rate, by-passing the pre-clarification, aerobic, and anaerobic selectors) [67, 185]. In principal, autochthonous ciliates appear suited to counteract abundant development of filamentous bacteria. However, only a few species capable of taking up filamentous bacteria are present in activated sludge plants, e.g., Trithigmostoma cucullus (Chilodonella cucullus), Trochilioides recta, Trochilia minuta, and Chilodonella uncinata. If these ciliates attain a high population density, a pronounced decline in filamentous bacteria and degeneration of bulking sludge is observed within a few days, both in bench-scale and operational plants [190–193]; see Fig. 14. Effective cell densities for Trochilioides sp. are quoted as 1000 ml–1 [190] and for Trithigmostoma cucullus and Trochilioides recta as 2000 ml–1 [193]. 3 Impairments of Protozoa: Consequences for Water Purification Ciliated protozoa are very numerous in all types of aerated biological treatment systems (compare Sects. 2.2.3 and 2.2.4). They play an important role in the purification process removing, through predation, the major part of dispersed bacteria that cause highly turbid, i.e., low quality effluent. It has been generally recognized that changes in the population density and community structure of ciliates affect the food web of this artificial ecosystem, thus influencing the performance of plants. Excess influx of toxic wastes with detrimental effects on ciliates would prevent clarification, thereby severely threatening the degradation process. A variety of chemicals can limit growth of ciliates. As with organisms from other taxonomic, functional, and trophic levels, the toxicological effects induced by organic and inorganic chemicals on ciliates vary widely, i.e., EC50values ranking from some mg l–1 to some g l–1 (reviewed by [194, 195]). Substances having toxic effects which diminish or even paralyze the purification performance frequently find their way into wastewater treatment plants with commercial and industrial wastewater. Risks are particularly great from metalfinishing works with electrochemical processes and wastewater from iron and steel pickling plants, accumulator-charging stations, stereotype, photocopy, photographic and printing works, dry-cleaning premises, industries producing pesticides, herbicides, and disinfectants, as well as tanneries, leather goods manufacturers and coking plants. In order to estimate the hazard potential and to lay down maximal concentrations, in addition to bacterial tests, biological tests with ciliates are indispensable to reflect potential risks of hazardous substances on the biological system of wastewater treatment as a whole. Tests with typical wastewater protozoa have been carried out for a number of toxic substances. Gracia et al. [196] observed effects of copper (sulfate) in concentrations of 1 mg Cu2+ l–1 on species diversity and population density – especially of the ciliates – of natural sludge samples. Madoni et al. [197] determined 242 W. Pauli et al. the 50% lethal effect concentrations of Cu<Hg<Cd<Pb<Cr<Zn (1 mg l–1 – 50 mg l–1) on various ciliates isolated from activated sludge, whereby the authors report differences in species sensitivities of up to two orders of magnitude, dependent on the heavy metal tested. Kakiichi et al. [107, 198–200] report inhibitory effects of disinfectants and surfactants on typical activated sludge ciliates. A comparison of the effect potential of 4 disinfectants towards the wastewater bacteria Alcaligenes faecalis and the wastewater ciliate Colpoda aspera reveals an almost 10-fold higher sensitivity of the ciliates [200]. Higher sensitivities of ciliates in comparison to bacteria were also found by Yoshioka et al. [201] for 32 wastewater relevant environmental chemicals. Results from the OECD activated sludge respiration test (RI Test, [202]) – considered as an indicator for acute effects of chemicals on heterotrophic bacterial flora – and growth tests with Tetrahymena, a ciliate typical in polysaprobic surface waters, but also found in activated sludge and submerged contactor plants [53, 55, 111, 112, 115, 203–209] were compared: 50% effect concentrations were, on average, 10 times lower with the ciliate test. Furthermore, certain substances proved highly toxic in the Tetrahymena test, and showed only weak effects in the respiration test; out of a total of 32 substances, just 6 cases had a (toxic) effect potential of less than 100 mg l–1. The weak correlation of r2 = 0.17 confirms the discrepancy between the two tests (Fig. 15). Similar observations of a low correlation were made by Pauli and Berger [210]. Figure 16 illustrates toxic responses of 4 ciliate species and standard tests with activated sludge towards industrial chemicals (data taken from the International Uniform ChemicaL Information Database, IUCLID, including toxic data of a wide variety of industrial chemi- Fig. 15. Acute effects of chemicals on the bacterial flora of activated sludge (OECD Respiration Inhibition Test) in comparison to those on the ciliate Tetrahymena pyriformis (growth inhibition) and on fish (OECD lethality test with Oryzias latipes); after [202] Protozoa in Wastewater Treatment: Function and Importance 243 Fig. 16. Comparison of results from standard activated sludge respiration tests and bioassays with ciliates (data from IUCLID); from [210] cals). Although a generally higher sensitivity of ciliates cannot be observed for this data set, the random distribution of points around the bisector confirms the dissimilarity of ciliate and activated sludge toxicities (r2 < 0.01, n = 35). Evidently ciliates are not only sensitive to pollutant induced stress, but test results reflect a series of additional toxic interactions, not represented by tests with bacteria in activated sludge. That this different toxic profile is probably due to the more complex cell-physiological – eukaryotic – organizational structure of the ciliates is implied by QSAR studies for heterogeneous chemical classes [211], which revealed a high correlation between the LC50 values found in the widely accepted fish lethality test (r2 = 0.78) with Tetrahymena growth, but not with bacteria test. 4 Environmental Biotechnological Aspects 4.1 Biodegradation Potentials of Ciliates Although it is well known that ciliate grazing on bacteria fulfills important tasks in the biological purification of sewage (compare Sects. 2.3.2.2 and 2.3.2.3) and that a number of technical methods and plant operation parameters obviously improve the purification efficiency by favoring ciliate growth (see Sects. 2.2.5, 2.3.2.2, and 2.3.2.3); only recently some pioneering attempts have been made to specifically use ciliates in biodegradation processes. Generally, large amounts of biosludge are formed in biological wastewater treatment processes and the separation, dewatering, treatment and disposal of this sludge represents major investment and operating costs. One of the poten- 244 W. Pauli et al. tially useful assemblies for reducing the sludge yield is the two-stage cascade used in many experiments for the study of ciliate-bacterial interactions, e.g., [140, 212–215]. The technique of a two-stage system enables one to manipulate the artificial ecosystem of conventional treatment processes so that dispersed bacteria are growing in the first part of the process and being consumed by protozoa in the last. Whereas in conventional treatment due to the growth of floc or film forming bacteria most of the bacterial biomass is protected against predation (see Sect. 2.3.4), dispersed bacteria can be readily taken up and metabolized by protozoa (see Sect. 2.3.2), resulting in a lower sludge yield (see Sect. 2.3.5). Operating the first part of the treatment process as an aerated tank reactor without biomass retention and at an hydraulic retention time short enough to prevent a significant growth of protozoa is a simple way to stimulate this growth of dispersed bacteria. Cultivations using synthetic wastewater and defined cultures of bacteria and ciliates in a two-stage chemostat cascade have shown that protozoan grazing can result in a considerable biomass reduction [128]. By introducing a “predation trap” (second stage) it was possible to obtain a decrease of 12–43% in biomass yield in comparison with a system without ciliate grazing. Studies of Lee and Welander [216, 217] confirm this potential of a two-stage system to reduce the sludge yield. Employing synthetic wastewater and mixed cultures of bacteria, protozoa and metazoa from activated sludge they observed a sludge yield around 30–50% of the yields typically obtained in conventional aerobic processes [216]. If authentic instead of synthetic wastewater was used as bacterial food supply the sludge production was also considerably lower than in conventional treatment [217]. Cox and Deshusses [218] developed a strategy to control biomass growth in biotrickling filters for waste air treatment by engineering predation of bacteria by protozoa. It was shown that clogging of bench-scale biotrickling filters could be slowed down with the use of protozoa. Interestingly, it was found that the reactor with protozoa had a shorter start-up time, possibly because of bacterial growth factors secreted by the protozoa. For the biodegradation of whey, the ciliate Tetrahymena had been chosen by Bonnet at al. [219] as a micro-organism capable of degrading and modifying the whey biologically in order to diminish its pollutant effect (whey is the aqueous phase that separates from the curds during cheese making or casein production). Disposal of crude whey completely arrested operation of lagoon pilots serving as example of receptor media, whereas the effects of biodegraded whey were only temporary, and normal operation was recovered within a few days. The authors stress that this method could be a valuable tool for small dairy farms, being unable to use complex industrial treatment technologies to forestall pollution by waste whey. Clearly, optimal conditions for protozoan activity need to be further evaluated and pilot scale experiments have to be performed to prove the influence of biomass predators in real treatment systems. Nonetheless these findings are auspicious, suggesting that specific use of ciliates can be made to improve biodegradation processes. Protozoa in Wastewater Treatment: Function and Importance 245 4.2 Ciliates as Biosensors As a constitutive group within the microbial food web, ciliates not only play an important ecological role in the self-purification and matter cycling of natural aquatic ecosystems, but also in the artificial system of sewage treatment plants. Their feeding on bacteria improve the treatment, resulting in higher transparency, i.e. lower organic loads in the output water of the treated wastes (see Sects. 1 and 2). This status of ciliates as an important functional group, improving the process in municipal sewage treatment, and furthermore that‚ “values from ciliate growth inhibition tests are relevant for the risk assessment for sewage treatment plants” has been recently acknowledged by a Technical Recommendation of the EEC [220]. There is a broad consensus in ecotoxicology that taxonomic similarity (i.e., close relationship, in terms of phylogeny) generally implies similar toxicological responses, e.g., [221, 222]. This is reflected in aquatic toxicology by selecting certain fish, crustacean, and algae species to represent trophic and taxonomic levels as a whole. A transferability of toxicological data for ciliates is also indicated. Although there exists an extraordinary amount of evolutionary distance between different genera and even between species of the same genus [223, 224], comparisons between the ciliates Colpidium, Colpoda, Paramecium, Tetrahymena, Uronema, and Vorticella reveal an almost comparable toxicological susceptibility [210]. Despite the lack of standardized ciliate test protocols, only 2 substances out of 13 exert a toxic effect differing by a factor of more than 100, whereas for the rest of the chemicals the deviations lie within about one order of magnitude (Fig. 17). The early use of ciliates in toxicity testing was reviewed by Persoone and Dive [225]. Among the ciliates, the organism of choice in aquatic toxicity testing has become the common freshwater hymenostome Tetrahymena [195, 226, 227]. Many features have contributed to making Tetrahymena – particularly the species T. pyriformis and T. thermophila – favorite models in cell biology and facilitated their modern day use as aquatic toxicity test organisms. It is worth mentioning, not only that these unicellular organisms can be grown under axenic, i.e., bacteria-free conditions, but also that they combine important advantages from two groups of organisms. Indeed, they belong to the higher cells, the eukaryotes, but they can be cultured both easily and economically like the prokaryotic bacteria. An innovative tool with the potential of a wide application has recently been offered by the introduction of a commercialized microtoxicity test kit with Tetrahymena (Protoxkit F, Creasel Ltd., Belgium). The test is specially designed for the use of environmental samples, thereby providing a helpful means to assess risks of sewage contaminants and their possible detrimental effects on the performance of waste water treatment plants. Following the concept of readyto-use microbiotests, with the test kit a ciliate multi-generation (growth) assay can be conducted by non-experts without sophisticated sample preparation and expensive equipment. Growth impairment tests with Tetrahymena have generally reached the highest degree of acceptance and standardization [195, 227, 228]: Based on an inter- 246 W. Pauli et al. Fig. 17. Comparison between toxic effects on ciliates from different genera (data from IUCLID, effect of methanol on T. pyriformis: own measurement). The arrows indicate cases where the ciliate data deviate by a factor of more than two orders of magnitude, from [210] national pilot ring test, a growth test with the ciliate Tetrahymena is recommended by the German Federal Environmental Agency for ecotoxicological risk assessment [229]. A final ring test to establish an internationally recognized Test Guideline has been initiated – an important step to include a traditionally untested, but ecologically important group of organisms in comprehensive ecotoxicity test batteries. 5 References 1. 2. 3. 4. 5. 6. 7. 8. 9. Berninger U-G, Finlay BJ, Kuuppo-Leinikki P (1991) Limnol Oceanogr 36:139 Jax K (1992) Limnologica 22:299 Thienemann A (1926) Verh Dtsch Zool Ges 31:29 Lindeman RL (1942) Ecology 23:399 Tansley AG (1935 Ecology 16:284 Bick H (1989) Ökologie. Gustav Fischer, Stuttgart Kolkwitz R, Marsson M (1908) Ber Dtsch Bot Ges 26a:509 Kolkwitz R, Marsson M (1909) Int Rev Gesamten Hydrobiol 2:126 Bick H (1964) Die Sukzession der Organismen bei der Selbstreinigung von organisch verunreinigtem Wasser unter verschiedenen Milieubedingungen. Habil-Schr, Düsseldorf 10. Bick H (1973) Am Zool 13:149 11. Pomeroy LR (1974) BioScience 24:499 Protozoa in Wastewater Treatment: Function and Importance 247 12. Porter KG, Paerl H, Hodson R, Pace M, Priscu J, Riemann B, Scavia D, Stockner J (1988) Microbial interactions in lake food webs. In: Carpenter SR (ed) Complex interactions in Lake Communities. Springer, Berlin Heidelberg New York, p 209 13. Graham JM (1991) J Protozool 38:66 14. Sieburth JM, Smetacek V, Lenz J (1978) Limnol Oceanogr 23:1256 15. Pomeroy LR, Wiebe WL (1988) Hydrobiologia 159:7 16. Stockner JG, Antja NJ (1986) Can J Fish Aquat Sci 43:2472 17. Stockner JG, Porter KG (1988) Microbial food webs in freshwater planktonic ecosystems. In: Carpenter SR (ed) Complex interactions in lake communities. Springer, Berlin Heidelberg New York, p 69 18. Sherr EB, Sherr BF (1991) Trends Ecol Evol 6:50 19. Porter KG, Pace ML, Battey JF (1979) Nature 277:563 20. Sanders RW, Porter KG, Bennet SJ, DeBiase AE (1989) Limnol Oceanogr 34:673 21. Sanders RW (1991) J Protozool 38:76 22. Bloem J, Bär-Gilissen M-JB (1989) Limnol Oceanogr 34:297 23. Pace ML, Orcutt JDJ (1981) Limnol Oceanogr 26:822 24. Beaver JR, Crisman TL (1982) Limnol Oceanogr 27:246 25. Beaver JR, Crisman TL (1989) Microb Ecol 17:111 26. Ducklow HW, Purdie DA, Williams PJL, Davies JM (1986) Science 232:865 27. Gifford DJ (1991) J Protozool 38:81 28. Pace ML, McManus GB, Findlay SEG (1990) Limnol Oceanogr 35:795 29. Carrick HJ, Fahnenstiel GL, Stoermer EF, Wetzel RG (1991) Limnol Oceanogr 36:1335 30. Hamilton FT, Taylor WD (1987) Can J Fish Aquat Sci 44:1038 31. Currie DJ, Kalff J (1984) Limnol Oceanogr 29:298 32. Buechler DG, Dillon RD (1974) J Protozool 21:339 33. Caron DA, Goldman JC, Dennett MR (1988) Hydrobiologia 159:27 34. Schönborn W (1982) Limnologica 14:329 35. Baldock BM, Sleigh MA (1988) Arch Hydrobiol 111:409 36. Harmsworth GC, Sleigh MA, Baker JH (1992) J Protozool 39:58 37. Lock MA, Wallace RR, Costerton JW, Ventullo RM, Charlton SE (1984) Oikos 42:102 38. Christensen ST, Wheatley DN, Rasmussen MI, Rasmussen L (1995) Cell Death and Differentiation 2:301 39. Neill WE (1994) Spatial and temporal scaling and the organization of limnetic communities. In: Giller PS, Hildrew AG, Raffaelli DG (eds) Aquatic ecology: scale, pattern and process. Blackwell Scientific, Oxford, p 189 40. Pöpel F (1980) Lehrbuch für Abwassertechnik und Gewässerschutz. Deutscher Fachschriften, Wiesbaden 41. Abwassertechnologie (1988) Deutsche Gesellschaft für Technische Zusammenarbeit (ed) Springer, Berlin Heidelberg New York 42. Böhnke B (1980) Wissenschaft Umwelt 1:27 43. Schlegel S (1995) Korrespondez Abwasser 8:1343 44. Meyer-Reil L-A (1994) Mar Ecol Prog Ser 112:303 45. Costerton JW, Lewandowski Z, Caldwell DE, Korber DR, Lappin-Scott HM (1995) Annu Rev Microbiol 49:711 46. Winkler M (1981) Biological treatment of waste-water. Ellis Horwood, Chichester 47. Hartmann L (1989) Biologische Abwasserreinigung Springer, Berlin Heidelberg New York 48. Pike EB (1975) Aerobic bacteria. In: Curds C R, Hawkes HA (eds) Ecological aspects of used-water treatment. Academic Press, London, p 1 49. Griebe T, Schaule G, Secker J, Flemming H-J (1996) Bestimmung der stoffwechselaktiven Bakterien im Belebtschlamm. In: Lemmer H, Griebe T, Flemming H-C (eds) Ökologie der Abwasserorganismen. Springer, Berlin Heidelberg New York, p 155 50. Wagner M, Amann R (1996) Die Anwendung von in situ-Hybridisierungssonden zur Aufklärung von Struktur und Dynamik der mikrobiellen Biozönosen in der Abwasserreinigung. In: Lemmer H, Griebe T, Flemming H-C (eds) Ökologie der Abwasserorganismen. Springer, Berlin Heidelberg New York, p 93 248 W. Pauli et al. 51. 52. 53. 54. 55. 56. 57. 58. 59. 60. 61. 62. 63. 64. 65. 66. 67. 68. Wagner M, Aßmus B, Hartmann A, Hutzler P, Amann R (1994) J Microsc 176:181 Bux F, Kasan HC (1994) Water S A (Pretoria) 20:61 Curds RC, Cockburn A (1970) Water Res 4:225 Chung JC, Strom PF (1991) Res J Water Pollut Control Fed 63:35 Madoni P, Ghetti PF (1981) Hydrobiologia 83:207 Varma MM, Finley HE, Bennett GH (1975) WPFC J 47:85 Sydenham DHJ (1971) Hydrobiologia 38:553 Hughes DE, Stafford DA (1976) Critical Reviews in Env Control 6:233 Curds CR (1982) A Rev Microbiol 36:27 Madoni P (1982) Acta Hydrobiol 24:223 Aescht E, Foissner W (1992) Arch Hydrobiol 90:207 Salvado H (1994) Water Res 28:1315 Kinner NE, Curds CR (1987) Water Res 21:481 Madoni P (1994) Water Sci Technol 29:63 Madoni P (1994) Bioresource Technology 48:245 Curds RC, Cockburn A (1970) Water Res 4:237 Mudrack K, Kunst S (1994) Biologie der Abwasserreinigung. G Fischer, Stuttgart Curds CR (1992) Protozoa and the water industry. Cambridge University Press, Cambridge Klee O (1968) Ger Mikrokosmos 57:231 Klimowicz H (1970) Acta Hydrobiol 12:357 Salvado H, Gracia MP (1993) Water Res 27:891 Salvado H, Gracia MP (1994) Verh Internat Verein Limnol 25:1950 Eikelboom DH (1988) Extra toepassingsmogelijkheden voor protozoa en metazoa bij de zuivering van afvalwater. TNO, Delft, Nr R88/286 Lee EGH, Mueller JC, Walden CC (1975) Tappi 58:100 Curds CR (1975) Protozoa. In: Curds CR, Hawkes HA (eds) Ecological aspects of usedwater treatment. Academic Press, London, p 203 Toman M, Rejic M (1988) Z f Wasser- und Abwasserforschung 21:189 Hill DL (1972) The biochemistry and physiology of Tetrahymena. Academic Press, London Hoffmann EK, Rasmussen L (1972) Biochim Biophys Acta 266:206 Orias E, Rasmussen L (1979) J Cell Sci 36:343 Rasmussen L, Zdanowski MK (1980) Experentia 36:1044 Seaman GR (1955) Metabolism of free-living ciliates. In: Hutner SH, Lwoff A (eds) Biochemistry and physiology of protozoa, vol 2. Academic Press, London, p 91 Cirillo VP (1962) J Bacteriol 84:754 Andersen AP, Hellung-Larsen P (1989) J Cell Biochem 41:125 Nilsson JR (1979) Phagotrophy in Tetrahymena. In: Lewandowski M, Hutner S (eds) Biochemistry and physiology of protozoa, 2nd edn, vol 2. Academic Press, New York, p 339 Sleigh M (1989) Protozoa and other protists. Edward Arnold, London Fenchel T (1980) Microb Ecol 6:1 Fenchel T (1980) Microb Ecol 6:13 Holz GG (1973) The nutrition of Tetrahymena: essential nutrients, feeding, and digestion. In: Elliott AM (ed) Biology of Tetrahymena. Dowden Hutchinson, Stroudsburg, p 89 Rasmussen L, Modeweg-Hansen L (1973) J Cell Sci 12:275 Fenchel T (1986) Progr Protistol 1:65 Hatzis C, Sweeney PJ, Srienc F, Fredrickson AG (1993) Biotechnol Bioeng 42:284 Seaman GR (1961) J Protozool 8:204 Curds RC, Cockburn A (1968) J Gen Microbiol 54:343 Baines S, Hawkes HA, Hewitt C H, Jenkins SH (1953) Sewage Indust Wastes 25:1024 Ministry of Technology (1968) Not Wat Pollut 43:1 Bark AW (1972) Annls Stn Limnol Besse 6–7:241 Augustin H, Foissner W, Bauer R (1989) Acta Hydrochim Hydrobiol 17:375 69. 70. 71. 72. 73. 74. 75. 76. 77. 78. 79. 80. 81. 82. 83. 84. 85. 86. 87. 88. 89. 90. 91. 92. 93. 94. 95. 96. 97. Protozoa in Wastewater Treatment: Function and Importance 98. 99. 100. 101. 102. 103. 104. 105. 106. 107. 108. 109. 110. 111. 112. 113. 114. 115. 116. 117. 118. 119. 120. 121. 122. 123. 124. 125. 126. 127. 128. 129. 130. 131. 132. 133. 134. 135. 136. 137. 138. 139. 140. 141. 142. 143. 144. 249 Curds CR (1971) Wat Res 5:1049 Ardern E, Lockett WT (1928) Manchester Rivers Dept Ann Rep 1:41 Pillay SC, Subrahmanyan V (1942) Nature 150:525 Pillay SC, Wadhwani TK, Gurbaxani MI, Subrahmanyan V (1944) Nature 154:179 McKinney RE, Gram A (1956) Sewage Ind Wastes 28:1219 Curds RC, Cockburn A, Vandyke JM (1968) Wat Pollut Control 67:312 Sridhar MKC, Pillai SC (1974) Environ Pollut 6:195 Macek M (1991) Single-species ciliate cultures controlling bacterial flocs distribution. In: Madoni P (ed) Proc Int Symp – Biol Approach to Sew Treatment Process. Perugia 1990, p 109 Farrah SR, Scheuerman PR, Eubanks RD, Bitton G (1985) Water Sci Technol 17:165 Kakiichi N, Kamata S, Ito O, Komine K, Otsuka H, Uchida K (1991) Anim Sci Technol (Jpn) 62:32 Curds CR, Fey GJ (1969) Wat Res 3:853 Mallory LM, Yuk CS, Liang LN, Alexander M (1983) Appl Environ Microbiol 46:1073 Pitman AR (1975) Water Pollut Control 74:688 Sudo R, Aiba S (1984) Adv Biochem Eng 29:117 Esteban G, Tellez C, Bautista LM (1990) Environ Technol 12:381 Fernandez-Leborans G, Moro P (1991) Bioresour Technol 38:7 Salvado H, Gracia MP, Amigo JM (1995) Water Res 29:1041 Sasahara T, Ogawa T (1983) Monatsschrift für Brauwissenschaft 11:443 Middeldorf JM (1989) Korrespondenz Abwasser 10:1165 Schlegel S (1986) Wasser – Abwasser 127:421 Hu HY, Fujie G, Urono K (1993) Wat Sci Technol 28:179 Schlegel S (1988) Korrespondenz Abwasser 2:120 Reilly M (1964) J Protozool 12:109 Weekers PHH, Vogels G D (1994) Journal of Microbiological Methods 19:13 Glaser D (1988) Microb Ecol 15:189 Hrudey SE (1982) J Water Pollut Control Fed 54:1207 Witthauer DP (1980) European J Appl Microbiol Biotechnol 9:151 Clarholm M (1984) Microbes as predators or prey. In: Klug MJ, Reddy CA (eds) Current perspectives on microbial ecology. American Society for Microbiology, Washington, DC, p 321 Macek M, Hartmann P, Škopová I (1993) Int Revue ges Hydrobiol 78:557 Macek M, Hartmann P (1991) Stud Environ Sci 42 (Environ Biotechnol):113 Ratsak CH, Kooi BW, Verseveld HW van (1994) Water Sci Technol 29:119 Taylor WD, Berger J (1980) Microb Ecol 6:27 Curds CR (1963) PhD Thesis, London University Curds CR (1963) J Gen Microbiol 33:357 Fletcher M, Loeb GI (1979) Appl Environ Microbiol 37:67 Costerton JW (1992) Int Biodeter Biodegrad 30:123 Lappin-Scott HM, Costerton JW (1995) Biofouling 1:323 Decho AW (1990) Microbial exopolymer secretions in ocean environments: their role(s) in food webs and marine processes. In: Barnes M (ed) Oceanogr Mar Biol Annu Rev, vol 28. Aberdeen University Press, p 73 Stehr G, Zörner S, Böttcher B, Koops HP (1995) Microb Ecol 30:115 Gause GF (1935) J Exp Biol 12:44 Curds CR (1970) Proc Symp on Methods of Study of Soil Ecology. UNESCO, Paris, France Watson PJ, Ohtaguchi K, Fredrickson AG (1981) J gen Microbiol 122:323 Swift ST, Najita IY, Ohtaguchi K, Fredrickson AG (1982) Biotechnol Bioeng 24:1953 Huber HC, Huber W, Ritter U (1990) Zbl Hyg 189:511 Hapte M, Alexander M (1975) Microbiol 29:159 Hapte M, Alexander M (1977) Microbiol 113:181 Ibanez F, Rassoulzadegan F (1977) Ann Inst Oceanogr 53:17 250 W. Pauli et al. 145. 146. 147. 148. 149. 150. 151. 152. 153. 154. 155. Clarholm M (1981) Microb Ecol 7:343 Fenchel T (1982) Mar Ecol Prog Ser 9:25 Fairbrother TH, Renshaw A (1922) J Soc chem Ind Lond 41:134 Güde H (1979) Microb Ecol 5:225 Shikano S, Luckinbill LS, Kurihara Y (1990) Microb Ecol 20:75 Sime-Ngando T, Bourdier G, Amblard C, Pinel-Alloul B (1991) Microb Ecol 21:211 Jürgens K, Stolpe G (1995) Freshwater Biol 33:27 Jürgens K, Pernthaler J, Schalla S, Amann R (1999) Appl Environ Microbiol 65:1241 Sommaragu R, Psenner R (1995) Appl Environ Microbiol 61:3457 Gurijala KR, Alexander M (1990) Appl Environ Microbiol 56:1631 Kjelleberg S (1984) Effects of interfaces on survival mechanisms of copiotrophic bacteria in low-nutrient habitats. In: Klug MJ, Reddy CA (eds) Current perspectives in microbial ecology. ASM, Washington, p 151 Jenkins SH (1942) Nature 150:607 Imhoff K (1993) Taschenbuch der Stadtentwässerung, 28 Aufl. Oldenburg, München – Wien Horan NJ (1990) Biological wastewater treatment systems, theory and operation. Wiley, Chichester Schönborn (1992) Arch Protistenkd 141:181 Calow P (1977) Biol Rev 52:385 Pirt SJ (1965) Proceedings of the Royal Society B 163:224 Macek M (1989) Int Rev Ges Hydrobiol 74:643 Kiy T, Tiedke A (1992) Appl Microbiol Biotechnol 37:576 Pauli W, Kühnel S, Berger S (1995) Neue biotechnologische Verfahren zur Gewinnung von Wertstoffen: Grundlagenuntersuchungen für die Überführung von Verfahren zur Erzeugung von Wertstoffen mittels Ciliaten in den klein- bzw. halbtechnischen Maßstab. Abschlußbericht zum Forschungsvorhaben des Bundesministers für Forschung und Technologie (Förderkennzeichen 0317383B) Lang H (1981) Wasserwirtschaft 71:166 Eberhard H (1984) Wasserwirtschaft 74:47 Wiggins BA, Alexander M (1988) Can J Microbiol 34:661 Elliott AM (1973) Biology of Tetrahymena. Dowden, Hutchinson and Ross, Stroudsburg, Pennsylvania Chudoba J (1985) Wat Res 19, 197–200 Andersson A, Lee C, Azam F, Hagstrom A (1985) Mar Ecol Prog Ser 23:99 Hervey RJ, Greaves JE (1941) Soil Sci 51:85 Nicoljuk VF (1969) Acta Protozool 7:99 Taylor GT, Iturriaga R, Sullivan CW (1985) Mar Ecol Prog Ser 23:129 Fenchel T (1988) Ann Rev Ecol Syst 19:19 Henkinet R, Couteaux M-M, Billes G, Bottner P, Palka L (1990) Soil Biol Biochem 22:555 Sherr BF, Sherr EB, Hopkinson CS (1988) Hydrobiologia 159:19 Barsdate RJ, Prentki RT, Fenchel T (1974) Oikos 25:239 Woods LE, Cole CV, Elliot ET, Anderson RV, Coleman DC (1982) Nitrogen transformations in soil as affected by bacterial-microfaunal interactions. Soil Biol Biochem 14:93 Bloem J, Starink M, Bär-Gilissen MJB, Cappenberg TE (1988) Appl Environ Microbiol 54:3112 Coleman DC (1985) Through a ped darkly: an ecological assessment of soil-root-microbial-faunal interactions. In: Fitter AH, Atkinson D, Read DJ, Usher MB (eds) Ecological interactions in soil. Blackwell, Oxford, p 1 Coleman DC, Crossley DAJ, Beare MH, Hendrix PF (1988) Agric Ecosyst Environ 24:117 Fenchel T, Harrison P (1976) The significance of bacterial grazing and mineral cycling for the decomposition of particulate detritus. In: Anderson JM, MacFadyen A (eds) The role of terrestrial and aquatic organisms in decomposition processes. Blackwell, Oxford, p 285 Rutherford PM, Juma NG (1992) Can J Soil Sci 72:217 Ratsak CH, Maarsen KA, Kooijman SALM (1996) Wat Res 30:1 156. 157. 158. 159. 160. 161. 162. 163. 164. 165. 166. 167. 168. 169. 170. 171. 172. 173. 174. 175. 176. 177. 178. 179. 180. 181. 182. 183. 184. Protozoa in Wastewater Treatment: Function and Importance 251 185. Lemmer H (1996) Ursachen und Bekämpfung von Blähschlamm. In: Lemmer H Griebe T, Flemming H-C (Hrsg) Ökologie der Abwasserorganismen. Springer, Berlin Heidelberg New York 186. Suwa Y, Imamura Y, Suzuki T, Tashiro T, Urushigawa Y (1994) Wat Res 28:1523 187. Alldredge AL, Cole JJ, Caron DA (1986) Limnol Oceanogr 31:68 188. Jeffrey WH, Paul JH (1986) Appl Environ Microbiol 51:1177 189. Karner M, Herndl GJ (1992) Mar Biol 113:341 190. Seguchi K, Koga M (1983) Proceedings of the 20th Annual Meeting of Sewage Works Researches. Tokyo, Japan 191. Hashimoto R (1985) J Jpn Sewage Works Assoc 22:61 192. Nitta T, Sakai Y, Mori T (1987) Appl Microbiol Biotechnol 26:195 193. Inamori Y, Kuniyasu Y, Sudo R, Koga M (1991) Water Sci Technol 23:963 194. Gilron GL, Lynn DH (1997) Ciliated protozoa as test organisms in toxicity assessment. In: Wells PG, Lee K, Blaise C (eds) Microscale testing in aquatic toxicology. CRC Press, Boca Raton 195. Sauvant MP, Pepin D, Piccini E (1999) Chemosphere 38:1631 196. Gracia MP, Salvado H, Rius M, Amigo J-M (1994) Acta Protozool 33:219 197. Madoni P, Davoli D, Gorbi G (1994) Bull Environ Contam Toxicol 53:420–425 198. Kakiichi N, Kamata S, Komine K, Uchida K (1989) Jpn J Zootech Sci 60:857 199. Kakiichi N, Matsui M, Kamata S, Komine K, Ito O, Hayashi M, Otsuka H, Uchida K (1990) Jpn J Zootech Sci 61:924 200. Kakiichi N, Yamamoto T, Kamata S, Otsuka H, Uchida K (1993) Anim Sci Technol (Jpn) 64:1013 201. Yoshioka Y, Nagase H, Ose Y, Sato T (1986) Ecotox Environ Saf 12:206 202. OECD (1983) OECD Guideline for Testing of Chemicals “Activated Sludge, Respiration Inhibition Test” Draft 1.8.83, No 210 203. Guhl W (1987) Korrespondenz Abwasser 34:1076 204. Poole J E P A (1987) Water Pollut Control 86:116 205. Luna-Pabello V M, Mayen R, Olvera-Viascan V, Saavedra J, Duran De Bazua C (1990) Biological Wastes 32:81 206. Al-Shahwani SM, Horan NJ (1991) Water Res 25:633 207. Esteban G, Tellez C (1992) Water Air Soil Pollut 61:185 208. Ratsak CH, Kooi BW, Kooijman B (1995) J Euk Microbiol 42:268 209. Martin-Cereceda M, Serrano S, Guinea A (1996) FEMS Microbiology Ecology 21:267 210. Pauli W, Berger S (1999) A new Toxkit microbiotest with the protozoan ciliate Tetrahymena. In: Persoone G, Janssen C, de Coen W (eds) New microbiotests for routine toxicity screening and biomonitoring. Kluwer Academic/Plenum Publishers, New York, p 169 211. Jaworska JS, Schultz TW (1994) Ecotoxicol Environ Safety 29:200 212. Curds CR, Cockburn R (1971) J Gen Microbiol 66:95 213. Jost JL Drake JF, Frederickson AG Tsuchia HM (1973) J Bacteriol 113:834 214. Ashby RE (1976) J Exp Mar Biol Ecol 24:227 215. Drake JF, Tsuchia HM (1977) Appl Environ Microbiol 34:18 216. Lee NM, Welander T (1996) Biotechnol Lett 18:429 217. Lee NM, Welander T (1996) Wat Res 30(8):1781–1790 218. Cox HHJ, Deshusses MA (1997) Annual Meeting and Exhibition of the Air and Waste Management Association. Toronto, Canada 219. Bonnet JL, Bogaerts P, Bohatier J (1999) Chemosphere 38:2979 220. ECB (1988) Effects assessment for micro-organisms in sewage treatment plants: consideration of protozoa toxicity data. Document European Chemicals Bureau 4/TR1/98, Technical Recommendation, TGD chap 3, sect 4 221. Suter GW (1982) Extrapolation of ecotoxicity data: choosing tests to suit the assessment CONF-821048–7 Environmental Protection Agency, USA 222. Volmer J, Kördel W, Klein W (1988) Chemosphere 17:1493 223. Schlegel M, Eisler K (1996) Evolution of ciliates. In: Hausmann K, Bradbury PC (eds) Ciliates, cells as organisms. Gustav Fischer, Stuttgart 252 W. Pauli et al. 224. Brunk CF, Kahn RW, Sadler LA (1990) J Mol Evol 30:290 225. Persoone G, Dive D (1978) Ecotoxicol Environ Safety 2:105 226. Schultz TW (1996) Tetrahymena in aquatic toxicology: QSARs and ecological hazard assessment. In: Pauli W, Berger S (eds) Proceedings of the International Workshop on a Protozoan Test Protocol with Tetrahymena in Aquatic Toxicity Testing. Umweltbundesamt-Texte 34/96, Berlin, Germany, p 31 227. Gilron GL, Lynn DH (1997) Ciliated protozoa as test organisms in toxicity assessment. In: Wells PG, Lee K, Blaise C (eds) Microscale testing in aquatic toxicology. CRC Press, Boca Raton 228. Pauli W, Berger S (1996) Proceedings of the International Workshop on a Protozoan Test Protocol with Tetrahymena in Aquatic Toxicity testing. Umweltbundesamt-Texte 34/96, Berlin, Germany 229. Heger W, Jung S, Martin S, Rönnefahrt I, Schiecke U, Schmitz S, Teichmann H, Peter H (1998) Chemikaliengesetz Heft 11, Ökotoxikologische Testverfahren mit aquatischen Organismen. Texte 58/98, Umweltbundesamt, Berlin, Germany