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From Total Suspended Solids to Molecular Biology Tools-A Personal View of Biological Wastewater Treatment Process Population Dynamics

Posted on: Sunday, 21 September 2008, 03:00 CDT

By Jenkins, David

ABSTRACT: The development of the tools needed to study the population dynamics of biological wastewater treatment processes is traced from its beginnings in the early 1900s to today's use of molecular biology tools (Oerther and Love, 2003). Examples of the benefits of population dynamics research in improving the performance and aiding the design and operation of biological wastewater treatment processes are given. Some thoughts on future areas of study are presented.

Water Environ. Res., 80, 677 (2008).

KEYWORDS: biological wastewater treatment processes, population dynamics, molecular biology tools.

doi: 10.2175/106143008X276679

Keynote Address-Association of Environmental Engineering and Science Professors Lecture, 80th Annual Water Environment Federation Technical Exhibition and Conference, San Diego, California, Oct. 13- 17, 2007.

Introduction

Biological waste water treatment processes are enrichment cultures (Beijeirinck, 1901). In the broadest sense, these are cultures whose composition is determined by the type of substrate(s) fed to them and the physical and environmental conditions under which they are grown.

Enrichment cultures were first used as preconcentration devices for isolating individual microorganisms from complex samples. They have been exploited for centuries in commercial/technological processes (beer, wine, bread, sauerkraut, soy sauce, kimchi, etc.). A pertinent illustration of the effect that enriching conditions have on culture composition and activity is to consider what would happen if the same wastewater stream were to be split into four parts, with one being sent to an activated sludge plant, one to a trickling filter plant, one to an oxidation pond, and one to an anaerobic digester. Although they were all developed on the same substrate and seeded with the same set of organisms, vastly different microbial communities and activities would develop in these four technological enrichment cultures.

Early Days

One of my first summer jobs provided me with an intimate introduction to biological wastewater treatment process population dynamics. My late father, Dr. S. H. (Sam) Jenkins, was the Chief Chemist and in charge of plant operation at the Birmingham Tame and Rea District Drainage Board (a predecessor organization to today's Severn Trent Plc, Birmingham, United Kingdom). At its Minworth plant (United Kingdom), the Board operated 17.4 ha (43 acres) of low-rate trickling filters that caused a severe "filter fly" nuisance, especially in the spring and early summer. Methods for controlling this nuisance were being investigated. The first of these was to dose the insecticide Gamexane (Lindane; gamma-1, 2, 3, 4, 5, 6- hexachlorocyclohexane) to the primary effluent entering the filters (Hawkes, 1955b). It was my job to stand by a primary effluent channel and help with shoveling the insecticide powder into the wastewater. It didn't work. The fly larvae present in the filters were killed off. This allowed the biofilm to accumulate further. Then, the larvae came back with even more vigor and produced a fly nuisance that was a lot worse than usual. Hawkes then proposed that the fly larvae could be flushed out of the filters by increasing the instantaneous wastewater dosing rate. He reasoned that these larvae would be washed out of the filter and replaced by Lumbricillus spp (worms with no adult fly stage), because the Lumbricillus, with its strong curved setae, could cling to the filter media, while the fly larvae could not. To investigate this, Hawkes devised multiyear experiments, in which baskets containing stone spheres were set at various depths in the filter, to be removed at intervals, for an assessment of the populations on them. My job was to help dig the holes in the filters for placing these baskets. This experience taught me the power of fullscale experiments and the beauty of being able to abstract something of clarity and logic from such a complex and variable system. It was this experience that convinced me to try and follow in my father's footsteps. Hawkes' hypothesis was borne out (Hawkes, 1955a, 1959) and developed into an extremely successful fly and biofilm control method (Albertson, 1995; Anonymous, 1983) that is still in use today. Hawkes' work is a classic in the field of biological process population dynamics. It should be required reading for anyone who sets out on a career in biological wastewater treatment.

The Activated Sludge Process

A discussion of the development of activated sludge science and technology usually starts with the experiments conducted in Manchester, England, by two chemists-Edward Arden and William T. Lockett-and reported in the Journal of the Society of Chemical Industry in 1914 (Ardern and Lockett, 1914). These experiments were done at the time when there was interest in developing biological methods of wastewater treatment as alternatives to physical and chemical techniques, one of whose principles was to stop the natural decomposition of the wastewater. Biological methods, on the other hand, sought to encourage and take advantage of these decomposition processes. Treatment by filters containing various types of solid media had been used with good results, but the "plain aeration" of wastewater was uneconomical because of the long period of time it took to achieve the desired results.

Arden and Lockett (1914) conducted municipal wastewater aeration experiments, in which they took the amazingly counterintuitive step of retaining the "solids accumulations" that developed in the vessels used for aerating the wastewater. They decanted the treated (nitrified) liquid and transferred the accumulated solids to more wastewater, aerated it, and again retained the solids for treating more wastewater, and so on. The progressive accumulation of solids might have led lesser beings to conclude that the wastewater was getting more, rather than less, polluted, or that the accumulating suspended solids were a waste product that should be discarded. Consider what must have passed through their minds as the contents of their cylinders became darker and darker, more and more dirty looking-but imagine their delight when they found that these dirty grey-brown mixtures, when settled, were associated with higher and higher wastewater purification rates.

Here is their account of these experiments:

"In a series of preliminary experiments, samples of Manchester raw sewage, contained in bottles of 80 oz. [2.4 L] capacity, were aerated until complete nitrification ensued . . .

. . .In the case of the first experiment, about 5 weeks' continuous aeration was required to obtain complete nitrification, as had been previously observed. At the end of this period, the clear oxidized liquid was removed by decantation, and a further sample of raw sewage aerated in contact with the original deposited matter until the sewage was again completely nitrified.

This method of treatment was repeated a number of times with the retention in each case of the deposited solids.

It was found that, as the amount of deposited matter increased, the time required for each succeeding oxidation gradually diminished, until, eventually, it was possible to completely oxidize a fresh sample of crude sewage within 24 hours.

For reference purposes and failing a better term, the deposited solids resulting from the complete oxidation of sewage have been designated activated sludge."On the basis of these papers, Ardem and Lockett (1914) have been credited with the development of the activated sludge process. However, its invention, in the legal meaning of the word, belongs to the firm of Jones & Attwood Limited (West Midlands, United Kingdom), who obtained (among others) British Patent No. 729 (1914), which referenced the recycle of solid(s) in an apparatus for the purification of wastewater or other impure waters and contained diagrams similar to the looped type of basins used by some currently marketed processes (Alleman, undated; Schnieder, 2007).

Collective Parameters

The first part of this paper deals with the development of substrate and microorganism parameters, because they form the foundation on which process population dynamics is built. Arden and Lockett's (1914) reference to the importance of deposited solids was probably the first reference to total suspended solids (TSS) as a measure for activated sludge (biomass) concentration. Collective measurement parameters for organisms, such as TSS, are "attractive", because they can be analyzed with relative ease, and "relevant", because they are the same parameters used to measure process loadings and effluent quality. However, they are "inexact", because they do not measure live organisms or their relevant activities in treatment processes. This inexactness has become increasingly important over the years, because biological wastewater treatment process objectives have expanded from biochemical oxygen demand (BOD) and TSS removals to include a multitude of other parameters.

Collective Organic Substrate Measurements. The collective parameters BOD and chemical oxygen demand (COD) have been widely used for process design and performance evaluation, even though they were originally conceived for measuring river water quality and later the strength of wastewater. These parameters are attractive, relevant, and inexact for the same reasons cited for TSS as a microorganism measurement. The inexactness of collective measures of substrates (and microorganisms) has led to lengthy, vigorous, and often heated discussion. Take, for example, the century-long debate about whether a biological method (i.e., BOD) (Adeney and Letts, 1908) or a chemical method (i.e., COD) is the more appropriate measure of limiting organic substrate concentrations in biological wastewater treatment processes. Before his death, my father gave me the originals of early correspondence on this topic between Professor W. E. Adeney of the Royal University of Ireland in Dublin and Dr. George McGowan, Chemist to the Royal Commission on Sewage Disposal (HMSO, 1901-1915). This correspondence concerned the validity of the following formula developed by McGowan for determining wastewater strength (McGowan et al., 1904): McGowan Strength = 4.5 N + 6.5 P

Where

N - (ammonia nitrogen + organic nitrogen), parts/10^sup 5^, and

P = N/8 permanganate value, parts/10^sup 5^ (a COD method [Tidy, 1879]).

These abstracts from a letter dated August 4, 1908, from Professor Adeney to Dr. McGowan, show Adeney's disdain of McGowan's formula:

"Burnham

Queen's Park,

Monkston, Co. Dublin

Aug. 4, 1908

My dear McGowan,

My objection to your formula is a fundamental one. There is no question in my mind of mathematical accuracy as you put it, but rather of incorrect statement as I have tried to explain to you in conversation. . . ."

"... I am afraid the table of results which you give does not afford much real support to the formula. Thus you give in all 14 analyses; the four which you mark with an asterisk cannot be regarded as sufficiently reliable for your purpose. Of the remaining 10, one only shows a concordance between the value obtained experimentally and that obtained from the formula. . . ."

". . . / have written frankly as you ask me, and because I genuinely dislike the formula.

With very kind regards,

Yours Sincerely,

W. E. Adeney"

Discussions of this sort have been held on many occasions during the almost 100 years since this letter was written. We have improved the COD method by making it able to almost completely degrade organic material (from the permanganate method to the dichromate method) (Moore, 1949), and we have improved the BOD analysis by making it more appropriate for measuring wastewater strength (by adding dilution water, nutrients, buffer, and nitrification inhibitor) (APHA et al., 2005).

Using BOD or COD as a measure of substrate concentration (and TSS as a measure of microorganism concentration) was appropriate as long as the major objectives of biological wastewater treatment were BOD and TSS removals. Now that much more is expected of biological wastewater treatment (i.e., removals of carbonaceous 5-day BOD [cBOD^sub 5^], total and soluble phosphorus, ammonianitrogen [NH^sub 3^-N], nitrate-nitrogen [NO^sub 3^-N], toxicity, metals, soluble organic nitrogen, and specific trace organics), it is not surprising that more precise substrate and biomass measurements are required.

Notable advances in the measurement of organic substrate in municipal wastewaters were the differentiation of soluble and paniculate BOD (or COD) and the concept of ease of biodegradability to better describe the organic matter function and effect in biological processes. Key contributions in this area were the early work of Symons and McKinney (1958) in defining biodegradable and nonbiodegradable soluble and paniculate substrates and the more recent studies by Ekama et al. (1986) and Marnais et al. (1993) in identifying and developing methods to measure the various categories of biodegradable organic matter (especially the readily biodegradable soluble fraction).

In the 1950s, there was a widespread introduction of synthetic anionic household detergents (HMSO, 1958). These detergents contained highly branched chain tetrapropylene benzene sulfonate surfactants, whose biodegradability was poor. Their residues persisted through biological treatment to create serious foaming at activated sludge plants and in receiving waters. The analysis of anionic surfactants in the influent and effluent of municipal wastewater treatment plants and the study of their biodegradability was the first time that a specific group of organics had been measured in municipal wastewater to evaluate its behavior during biological treatment. Nowadays, the measurement of biodegradability under biological wastewater treatment conditions is a standard testing procedure (OECD, 1993) required for new products that will be discharged to wastewater in significant quantities.

Collective Microorganism Measurements. As indicated previously, the use of TSS to measure microorganisms in activated sludge arose from its use as a pollutant parameter and for judging the accumulation of activated sludge in treatment plants. It was (and still is) an easy measurement and one that treatment plant laboratories are almost always set up to do. Moreover, plants typically use TSS measurements in determining the quantity of sludge to be treated and disposed. The use of TSS for determining the organic loading rate of activated sludge (i.e., kilograms BOD^sub 5^ per kilogram TSS per day) and later the solids residence time (SRT) were great successive improvements over volumetric loading parameters (i.e., kilograms BOD^sub 5^ per cubic meter per day [pounds BOD^sub 5^ per thousand cubic feet per day])-parameters that acknowledged the size of the aeration tank, but not that there were any organisms in it.

A small refinement of this crude microorganism mass measurement is the use of volatile suspended solids (VSS) instead of TSS. It was reasoned that VSS was more representative of microorganisms than TSS, because the latter included both inorganic and organic matter. The logic of this reasoning was never clear to me, because all living organisms contain both volatile matter and ash. The use of VSS is helpful in determining the relative biomass levels in activated sludges treating wastewaters with different inorganic suspended solids contents and for activated sludges operated at a wide range of organic loadings. Difficulties can still occur when using VSS as a biomass surrogate for wastewaters or processes that contain significant amounts of volatile, nonbiodegradable, or poorly biodegradable solids, for example, cellulose fibers in pulp and paper wastewaters and powdered activated carbon in the PACT process (Johnson and Lee, 1975; Kalinske, 1972). These special situations aside, the assumption that VSS is proportional to live microorganisms of unit activity has served the wastewater industry long and well.

Starting in the early 1960s, publications began to appear on attempts at refining microorganism mass and activity measurement beyond that provided by VSS measurement (Agardy, 1963; Genetelli, 1967). The major thrust of this work was to try and incorporate some index of the viability and biological activity of the activated sludge in the measurement. At that time in the United States, the major objective of secondary treatment processes was to remove the gross organic pollutants (measured by BOD and TSS), so that the appropriate sludge activity measures were those that reflected the overall activities of the aerobic heterotrophs in the culture. Thus, they included parameters such as specific oxygen uptake rate (milligrams oxygen per gram VSS per hour); organic matter removal rate (food-to-microorganism ratio [F/M], grams COD per gram VSS per day); dehydrogenase activity (milligrams triphenylformazan per gram VSS per day); and adenosine triphosphate (ATP) concentration (micrograms ATP per gram VSS) (Bucksteeg and Theile, 1959; Ford et al., 1966; Kotze, 1967). Weddle and Jenkins (1971) found that steady- state values of oxygen uptake rate (OUR), dehydrogenase activity, and ATP content all increased with activated sludge growth rate (decreased with SRT), if expressed on a VSS basis (i.e., activity per gram VSS per hour), but were fairly constant with activated sludge growth rate, if expressed on a viable cell basis (i.e., activity per gram viable cells per hour). In these experiments, the viable cell concentration of a sonically homogenized sample was measured by plate counting on an agar medium made up with the extract of autoclaved activated sludge.

Another of Weddle and Jenkins' (1971) conclusions was that, for virtually the entire range of growth rates over which full-scale activated sludge systems operate, the biomass viability could be considered constant and very low (10 to 20%). Recent investigations using molecular biology tools (MBTs) suggest that the vast majority of the microorganisms in natural systems, such as activated sludge, are nonculturable (Amann et al., 1995). This raises the question of whether the low viabilities found by Weddle and Jenkins (1971) were truly low or whether they were the result of the inability to culture all of the viable microorganisms present.

The single early widespread exception to the statement that "the appropriate sludge activity measures were those that reflected the overall activities of the aerobic hetewtrophs in the culture" was the need to measure the nitrification rate in plants that were required to nitrify. Probably because this requirement was more common in Western Europe than in the United States, most of the pioneering work on nitrification was done at the British Water Pollution Research Laboratory (Downing and Hopwood, 1964; Downing et al., 1964; Knowles et al., 1965). This work laid the foundation for the development of much of our current understanding of the factors (i.e., dissolved oxygen, pH, temperature, ammonia concentration, and sludge growth rate) that affect nitrification rate in activated sludge. Recent work funded by the Water Environment Research Foundation (Alexandria, Virginia) has validated and refined these results (Melcer et al., 2002). The need to measure nitrification rate was just the harbinger of things to come. As biological treatment requirements expanded to include nitrogen and phosphorus removal, the need arose for determining other specific activities, such as denitrification rate (milligrams NO^sub 3^-N per gram VSS per hour) (Burdick et al., 1982), anaerobic phosphate (PO^sub 4^-P) release rate, and aerobic phosphate uptake rate (milligrams PO^sub 4^-P per gram VSS per hour) (Marnais and Jenkins, 1992). It should be noted that measurements of specific activities and rates are much more common in the industrial wastewater treatment field because of the many types of organics that have to be treated and because it is generally easier to identify and analyze for the individual compounds that must be removed.

It is reasonable to expect that, as treatment requirements for municipal wastewaters are further expanded, more specific biological treatment process activity rate measurements will be required. The rate of various refractory organic compound production (i.e., trihalomethane precursors) may be one of the next important rate measurements required.

Even with the ability to identify and measure specific activated sludge activities, such as those given above, one still does not obtain an intrinsic rate, because the denominator term in all these rates is still VSS or TSS. Rates expressed this way are acceptable for estimating the rates of general reactions carried out by the heterotrophic biomass, but they are not always appropriate for determining the specific rates of bacteria that are present in very small amounts in the biomass. For example, in an activated sludge treating typical domestic wastewater under conditions that allow full nitrification and enhanced biological phosphate removal (EBPR), the nitrifiers and polyphosphate-accumulating organisms (PAOs) would each make up less than 10% of the VSS. In situ genetic techniques for detecting, and, to some extent, enumerating, microorganisms have great promise for providing the ability to obtain specific rates of important activated sludge activities. Thus, with a gene probe, such as Nitri-VIT Vermicon AG (Munich, Germany), it is possible to determine whether a nitrification rate is low because the nitrifier population itself is low or because some factor in the wastewater or plant operation (i.e., toxicity) is decreasing the specific nitrification rate. Findings of this sort are of enormous economic importance in the design of biological nutrient removal activated sludge plants.

Process Models

The ability to better define the behavior of influent substrate fractions (i.e., organics, nitrogen forms, and phosphorus forms) and biomass fractions (i.e., heterotrophs, nitrifiers, denitrifiers, and PAOs) and their reaction stoichiometry and kinetics together with the power of the modern computer have made it possible to develop sophisticated dynamic models for the design, analysis, and performance prediction of complex biological treatment plants. The work of the International Water Association (London, United Kingdom) Task Group on Activated Sludge Modeling (IWA, 2000) is an excellent and continuing achievement of such modeling efforts that has resulted in the development, commercialization, and widespread use (and unfortunately misuse) of process models, such as Biowin (ASDM) (EnviroSim Associates Ltd., 2007), GPSX (Hydromantis Inc., 2004), STOAT (WRc Plc, 2007), WEST (MOSTforWATER, 2007), and SIMBA (Ifak System gmbH, Magdeburg, Germany). Biological treatment models, however, started long before this; the early workers in this area deserve recognition. The people who influenced me most with their pioneering papers on activated sludge modeling were Eckenfelder and McCabe (1960), Garrett and Sawyer (1952), McCarty (1971), and McKinney (1962).

When I ponder these modern models, I cannot help thinking back to the time of Adeney and McGowan. I believe that both of these remarkable gentlemen would be content with today's biological process models. McGowan would be vindicated to see that a chemical method (COD) had been adopted as the basic measure of organic matter, and Adeney would doubtless be consoled by the fact that the organic matter was functionally fractionated using biological methods.

Population Dynamics and Physical Properties of Activated Sludge

The abilities to settle/clarify and thicken activated sludge are key to the successful functioning of the process. The properties that determine the success of solids separation processes are an inextricable weave of the microbiological, chemical, and physical properties. Activated sludge settles slowly and compacts poorly because of an overabundance of either exocellular material or filamentous microorganisms. Dispersed growth, consisting of individual microorganisms or very small floes, is caused by chemical and/or physical disruption of flocculation mechanisms. Biological foaming is caused by the presence of microorganisms with hydrophobic cell walls.

The early work of Butterfield (1935) suggested that activated sludge was made up of floe-forming bacteria (i.e., Zoogloea spp.), and the studies of Lackey and Wattie (1940) suggested that the presence of trichome (filament)-forming organisms (i.e., Sphaerotilus natans) could cause poor settling. The microbiologically unsophisticated among the activated sludge community grasped the importance of these results, but missed an important detail stated by Lackey and Wattie-"Since bulking of activated sludge may occur in the absence of Sphaerotilus, it is not the cause of all bulking". In spite of this caution, all floe- forming bacteria became Zoogloea ramigera, and all filamentous organisms became Sphaerotilus natans.

Two investigations were important in dispelling these myths. Alien (1944) vigorously homogenized activated sludge (with a cream maker) before enumerating the microorganisms in it. Finding much higher plate counts than anyone before him, he concluded that there were many types of microorganisms present in activated sludge floes and that those native to natural waters and soil environments were far more important members of the activated sludge floe community than those from the guts of warm-blooded animals (enrichment culture principles suggest that activated sludge, an aerobic environment at ambient temperatures, will select "soil and water bacteria" over the anaerobic microorganisms growing in the guts of warm-bodied animals).

Eikelboom's work in identifying the filamentous organisms in activated sludge (Eikelboom 1975, 1977; Eikelboom and van Buijsen, 1981) provided an important tool for understanding filamentous bulking. They developed a filamentous organism key based on in situ morphological characteristics and staining reactions. This method dispensed with the traditional approach dictated by Koch's postulates (i.e., "observe conditions, isolate organism, grow organism in pure culture, reintroduce organism to environment, confirm observed condition is the same as that from which organism was isolated") (Koch, 1890). Eikelboom's pragmatic approach must have seemed heretical to classical microbiologists, but was invaluable to the wastewater treatment community. In retrospect, an in situ key for filamentous organisms in activated sludge was an excellent choice over using isolation, growth, and characterization of pure cultures, because we now know that the following activated sludge filamentous organisms have not yet been isolated in pure culture and grown on traditional media: types 0092, 0041, 0675, 0914, 0961, and 0581 (Jenkins et al., 2003).

Eikelboom and others all over the world have related the occurrence of various filament types to the conditions under which they were observed (Blackbeard et al., 1986; Eikelboom and van Buijsen, 1981; Kristensen et al., 1994; Richard et al., 1982; Rossetti et al., 1994; Seviour et al., 1990; Strom and Jenkins, 1984; Wagner, 1982; Wanner et al., 1998). In effect, these workers observed the enrichment conditions and related them to the filaments that they enriched. Having this information, it was possible to "turn the tables" and use the observation of large numbers of a particular filamentous organism(s) in a bulking activated sludge as a diagnostic tool for determining the cause of the bulking.

The most important conditions influencing the growth of filamentous organisms causing bulking were dissolved oxygen concentration, nutrient deficiency, sulfide concentration, aeration basin configuration, and the presence of initial unaerated zones in the aeration basin.

Palm et al. (1980) showed that the dissolved oxygen concentration required to prevent low-dissolved-oxygen bulking was a function of the organic loading (OUR) of the activated sludge; these findings were confirmed by the pure culture experiments of Lau et al. (1984a and b) and Richard et al. (1985) on two species of Sphaerotilus. Shimizu et al. (1985) showed that, under nitrogen deficiency, filamentous organism type 02IN competed successfully, because it could activate a higher ammonium uptake rate than a floe-forming organism.

Chudoba and coworkers (Chudoba, 1985; Chudoba et al., 1974, 1982, and 1985; Chudoba, Grau, and Ottova, 1973; Chudoba, Ottova, and Madera, 1973) and Lee et al. (1982) showed the importance of reactor compartmentalization on sludge-settling characteristics, van Niekerk et al. (1987, 1988) validated these results with competitive chemostat growth experiments of a pure culture of type 02IN and a floe-forming microorganism. Davidson (1957, 1959) was issued United States patents for activated sludge with an initial anaerobic zone to improve settling properties. Heide and Pasveer (1973) confirmed this claim for oxidation ditch activated sludge. Gabb et al. (1991) and Shao and Jenkins (1989), respectively, defined the filamentous bulking control capabilities of anaerobic and anoxic selectors. In the United States, the first widespread application of the results of the studies referred to above was by the manufacturers of oxygen activated sludge systems (Albertson et al., 1970) and by Barnard (1976) and Spector (1977). The concepts that compartmentalized aeration basins with initial unaerated (anoxic/anaerobic) zones and with sufficient dissolved oxygen throughout the aerated zones typically produce better settling sludge than completely mixed and uniformly aerated systems have gradually gained acceptance in the United States. Now, happily, these ideas seem to be firmly entrenched in most of the engineering and operations communities in the United States.

There is still one area of concern; activated sludge design and wastewater treatability information is often obtained from operating batch-fed (fill-and-draw, sequencing batch reactor [SBR]) laboratory or pilot units. Typically, these units are mixed with the air that provides the aeration, and often there is no dissolved oxygen control. This type of unit most likely experiences a zero dissolved oxygen concentration immediately after feeding, followed by a considerable period of time, during which, the dissolved oxygen concentration is at, or very close to, saturation. This combination of conditions tends to produce activated sludge with excellent settling properties, because its feeding method simulates an initial anaerobic/anoxic zone, followed by a plug-flow aeration basin, and its aeration regime discourages the growth of low-dissolved-oxygen filamentous organisms. From the information provided by this type of laboratory system, full-scale plants are often designed, in which the aeration basins are complete-mix, the feed is continuous, and the dissolved oxygen concentration is controlled to levels far below saturation. This combination of conditions is the most favorable for filamentous organism growth. The result is that the full-scale plant sludge settleability is much poorer than that predicted from the laboratory experiments. In addition, the rates of substrate uptake in batch-fed systems can be higher than those in complete-mix reactors, so that these too are overestimated. Together, these factors can have serious consequences for the owners and designer (van Niekerk et al., 1987).

The lessons learned from studies of the relationship of filamentous organism growth to reactor configuration, anaerobic/ anoxic feed zones, and aeration-zone dissolved oxygen concentration lead us to the conclusion that laboratory- and pilot-scale systems used for plant design must faithfully simulate the enrichment conditions that will be used in the full-scale plant, if they are to accurately predict its sludge-settling characteristics.

Molecular Biology Tools

With the development and application to activated sludge of modern molecular biological, microscopic, and analytical methods (Table 1), we are now at the beginning of what could be another great leap forward in our understanding of biological wastewater treatment processes, which could lead to significant advances in their application, design, and operation. These methods allow the specific, in situ characterization of morphologically indistinguishable and nonculturable members of the activated sludge community, determination of their distribution throughout the culture, assessment of individual and culture-wide metabolic capabilities, and measurement of expressed individual organism and culture enzymatic activities.

With these abilities at hand and with the prospect of much more in view, I would like to quote some wise words written approximately 40 years ago by Professor Wesley O. Pipes (Pipes, 1966).

"... Research on the activated sludge process should be directed toward providing solutions to practical problems. The objective of attempts to identify the organisms present in activated sludge should be not to compile a list of organisms, but to determine the role of organisms in the process...

... There is another, more profound, reason that engineers have not made better use of the information available about the ecology of activated sludge ... Engineers need and expect to have quantitative data upon which to base their calculations, and most will not even read the results of qualitative studies. Many of the ecological studies of activated sludge have resulted only in descriptive material. The great challenge in this field of research at present is to produce useful quantitative data ...."

The messages in this quotation are still relevant today. However, the recent literature contains many papers on biological wastewater treatment process cultures that consist largely of lists of microorganisms with their positions on the "tree of life" elegantly worked out. On this topic, Rittman et al. (2006) aptly state the following:

"After 15 or so years of profitable 'stamp collecting', microbial ecology and environmental biotechnology are poised to advance to a higher plane of molecular-based research."

It is important to urge modern microbiologists and environmental engineers to work together, so that the information produced will be both scientifically credible and practically relevant.

Application of Population Dynamics Research

The penultimate section of this paper provides four examples of the advances in activated sludge process design and operational control that can be made by the practical application of population dynamics research.

Nocardioform Foaming. Activated sludge plants are commonly plagued by the occurrence of high TSS content, brown foams that are enriched over the mixed liquor with microorganisms having hydrophobic cell walls. These microorganisms include a group referred to as nocardioform organisms (nocardioforms, Gordonia amarae-l&s organisms [Lechevalier and Lechevalier, 1974]) and an organism named Microthrix parvicella. Only the nocardioforms are discussed here.

The mechanism of foam formation seems to be flotation, in which aeration air bubbles attach to the hydrophobic nocardioform cell walls and float them (and associated floe material) to the surface. At the surface, the interfacial film of the air bubble is stabilized by the hydrophobic particles, and film drainage can be impeded by them. When the air bubble finally collapses, it leaves the floated hydrophobic solids at the surface.

To study the nocardia foaming problem, it is necessary to have a quantitative method for measuring the amount of nocardioforms in activated sludge. Unlike the use of settling tests, such as sludge volume index, for indirectly estimating the amount of bulking filamentous organisms in activated sludge, it is not possible to use a foaming test to estimate nocardioform levels, because foaming is influenced by factors other than the quantity of nocardioforms (such as the presence of surfactants) and because laboratory foam tests are insensitive estimators of nocardioform levels. The nocardioform enumeration method (Pitt and Jenkins, 1990; Vega-Rodriguez, 1983) consists of evenly spreading a known volume of blended activated sludge on a microscope slide, air drying, Gram-staining, and then counting the number of Gram-positive branched filaments longer than 1 urn that intersect lines drawn on the slide.

Using this method, the effect of SRT on nocardioform levels was determined using reactors fed with wastewater from the San Francisco (California) South East Plant. This work validated the existing practice of using low SRT to control foaming, but, surprisingly, it was impossible to completely wash out the nocardioforms at very low SRTs and at very low temperatures (Pitt and Jenkins, 1990). When washout experiments were repeated in reactors at the Sacramento (California) Regional Plant, nocardioform washout at low SRT was complete. The washout SRT was temperaturedependent, and maximum nocardioform levels were much lower than those in the reactors receiving San Francisco wastewater (Cha et al., 1992). These differences were all related to the foam removal and disposal methods in the full-scale plants from which the reactor feed wastewater was drawn. At San Francisco, removed foam was recycled to the headworks, thereby seeding (i.e., bioaugmenting) the reactors, elevating their nocardioform populations, and preventing complete nocardioform washout. At Sacramento, removed foam was disposed of in a way that did not return it to the plant. Without the seeding, complete nocardioform washout was possible, and nocardioform populations were much lower.

The Sacramento studies also showed the effect of foam trapping on nocardioform populations. In activated sludge systems with subsurface aeration basin draw-off and secondary clarifier surface scum baffles, nocardioform levels were five times higher than in a system with an overflow aeration basin draw-off and no clarifier surface baffles (Cha et al., 1992).

Besides elevating nocardioform populations, foam trapping encourages dispersed rather than floe-bound nocardioform growth. In chemostats with overflow effluent draw-off, Gordonia amarae grew as pellets, but, with subsurface effluent withdrawal (Koopman et al., 1980), it grew in a dispersed form (Blackall et al., 1991). For hydrophobic organisms, an overflow outlet favors clumped growth over dispersed growth, because dispersed organisms can float and pass into the effluent, while the clumped organisms can sink and be retained in the reactor. Conversely, a subsurface overflow favors the dispersed over the clumped growth form, because dispersed organisms will not be lost if they float, and growth in this form provides greater access to the substrate.

These findings suggested that nocardioform growth in activated sludge would be strongly influenced by reactor configuration. This supposition was confirmed by the studies of Narayanan (2003) and Shao et al. (1997). Shao and coworkers showed, at full-scale, that cationic polymer addition eliminated nocardioform foam, but only slightly reduced total nocardioform counts. It was hypothesized that the polymer flocculated the dispersed nocardioforms into the floes, converting them to floe-bound nocardioforms. Dispersed nocardioforms produce more foam than the floe-bound form, because they expose more of their hydrophobic surface to the air bubbles. This hypothesis was tested by Narayanan et al. (1993), who developed a nocardioform enumeration method that could distinguish between dispersed and floe- bound organisms (Narayanan, 2003). Both a cationic polymer and polyaluminum chloride reduced foam levels and dispersed nocardioform counts. Narayanan and coworkers also showed that trapping reactors produced significant dispersed nocardioform levels, while there were no dispersed nocardioforms in reactors without surface trapping. Finally, foam levels were found to be a function of only the amount of dispersed nocardioforms and were not related to the amounts of total or floe-bound nocardioforms. Thus, nocardioform foaming can be explained in largely physical terms, as follows: surface traps retain foam and increase both the total nocardioform population and the fraction of it that is the foamcausing, dispersed form, thereby further exacerbating foaming. The two most effective foam control methods also rely on physical rather than biological phenomena. Already mentioned is the way in which cationic polymer eliminates foam by incorporating dispersed nocardioforms to the floes by flocculation. Surface removal of some or all of the waste activated sludge provides an outlet for dispersed nocardioforms and prevents their accumulation, even in systems with surface traps (Pagilla and Jenkins, 1996; Parker et al., 2003; Pretorius and Laubscher, 1987).

This case study shows that, in addition to the metabolic and physiological activities of microorganisms, their population dynamics can be determined by physical, chemical, and hardware factors in the real world of wastewater treatment.

Dispersed Growth and Toxicity. The inability to form strong floes is the third most common solids separation problem after filamentous bulking and foaming. Higgins and Novak (1997) proposed that microorganisms in activated sludge floes are held together by exocellular polymers (largely negatively charged proteins and carbohydrates), which are ionically bridged through the adsorption of divalent cations, such as calcium and magnesium. They further proposed that high influent monovalent cation concentrations, leading to high monovalent-to-divalent cation ratios, caused floe dispersion through destruction of the divalent cation bridging between exocellular biopolymers. This occurred because cation adsorption can be likened to ion exchange, in which a favorably adsorbed divalent cation can be displaced by a high solute concentration of a less well-adsorbed monovalent cation. Excellent results in decreasing dispersed growth in full-scale activated sludge plants treating a high-sodium industrial wastewater were obtained by adding a source of divalent cation (magnesium sulfate [MgSO^sub 4^]) to decrease the monovalent-to-divalent cation ratio of the wastewater (Murthy et al., 1998).

Floe dispersion leading to the production of dispersed bacteria often occurs shortly after activated sludge has received toxic materials. Bott and Love (2002) showed that a series of physiological responses to sublethal, toxicity-induced stress caused the efflux of potassium (K+) from cells to the interfloc solution, elevating its monovalent-to-divalent cation ratio, causing floe breakup, and producing dispersed cells. Increases in bulk-solution potassium concentration occurred when activated sludge was deflocculated by sodium hypochlorite (NaCXTl) addition (Wimmer and Love, 2004). Measurement of bulk potassium concentration may be a simple, sensitive monitoring method for the early detection of influent toxicity (Gillam et al., 2005).

This example shows how knowledge of microbial metabolic mechanisms, floe structure, and instrumentation can be combined to diagnose and solve activated sludge operating problems to provide sophisticated monitoring methods.

Bioaugmentation. Until recently, the use of bioaugmentation in biological wastewater treatment processes has not been readily accepted. I could not understand how it was possible to add relatively small quantities of organisms to a heavily seeded enrichment culture and expect them to compete with the organisms being continuously selected by the enrichment conditions in the process. I had never been shown a properly controlled, sufficiently documented, long-term study whose results showed any difference between the control and test systems. Then, there was the fact that it was necessary to keep adding organisms (albeit at a lower rate) to maintain the desired effect. This indicates that the added organisms were not competing very well.

Three types of processes that could be considered to be bioaugmentation changed my mind. These were the following:

(1) The use of sidestream nitrification of streams produced by dewatering anaerobically digested sludge to aid nitrification in the mainstream aeration basins (Kos, 1998).

(2) Short-term bioaugmentation with commercially available cultures for "kick-starting" biological processes, such as nitrification.

(3) The recycle of nocardioform and M. parvicella foams back through the activated sludge system that produced more severe foaming (i.e., negative bioaugmentation).

While these cases support the utility of bioaugmentation, they do not change my viewpoint of the conditions necessary to make it work in activated sludge treating a heavily seeded wastewater like municipal wastewater. Continuous bioaugmentation will only succeed if the appropriate form of the desired organisms is added in sufficient quantities to overcome competition from the resident microorganisms derived from the influent and return sludge streams. Sufficient quantities of augmenting microorganisms must be available at an affordable price. Sidestream nitrification of liquid from the thickening of anaerobically digested sludge is a good example of this.

Wagner (1982) cautions that, in using bioaugmentation for kickstarting nitrification, it is important to choose a seed containing the appropriate microorganisms. For example, both Nitmsococcus and Nitrosonuwnas can be active in ammonium (NH4+) oxidation, and Nitmbacter and Nitmspira can both be important nitrite (NC>2~) oxidizers. Bouchez et al. (2000), using fluorescence in situ hybridization (FISH), showed that the bacteria used to bioaugment the denitrifying capacity of a laboratory-scale SBR were very rapidly eaten by stalked ciliated protozoa, because the added bacteria were in the dispersed form. Wagner (1982) suggested that polymer-embedded microorganisms could survive protozoan grazing activities.

Nitmspira and Nitmbacter seam to differ in their affinity for nitrite, with Nitmspira favored by low nitrite concentrations and Nitmbacter by high nitrite levels (Schramm et al., 1999). It is possible that "nitrite lock" can be explained by these observations. In this phenomenon, difficulty is experienced in returning a system to the complete nitrification of ammonium to nitrate (NO^sub 3^^sup - ^) once it has, for some reason, only been partially nitrifying to nitrite (Muirhead and Appleton, 2007). If Nitmbacter become established at the high nitrite levels, they may not be able to reduce the nitrite levels down to those where Nitmspira, with its ability to produce low nitrite levels, becomes the dominant nitrite oxidizer. In such cases, bioaugmentation with a culture rich in Nitmspira would be indicated.

Enhanced Biological Phosphate Removal. This process has been a particularly fruitful area for the use of MBTs in correcting mistaken identities. Activated sludge, with an initial zone devoid of dissolved oxygen and nitrate, can remove more phosphate than needed for microbial growth through cellular storage of inorganic polyphosphates, which allows the anaerobic uptake and storage of organic substrate. Using information from culture-dependent methods, PAOs were first thought to be Acinetobacter (Fuhs and Chen, 1975). Acinetobacter was studied in this context for many years (i.e., Hao and Chang, 1987). Jenkins and Tandoi (1991) noted that Acinetobacter did not take up organic substrate anaerobically-one of the requirements for a PAO. Using MBTs, Crocetti et al. (2000) and Hesselmann et al. (1999) showed that a novel microorganism in the family Rhodocyclaceae with the proposed name Candidatus 'Accumulibacter phosphatis' was an important PAO, and Crocetti et al. (2000) related its population in an EBPR activated sludge to the sludge phosphorus content. It is likely that there are other important PAOs in activated sludge. For example, Actinobacteria have been found in abundance in EBPR systems by several workers (Bond et al., 1999; Christensson et al., 1998; Kong et al., 2005; Liu et al., 2001).

In recent work, Martin et al. (2006) presented the almost complete genome for two laboratory-scale EBPR cultures enriched with Candidatus 'Accumulibacter phosphatis'. Genes were present for coding two types of phosphate transport enzymes that differ in their affinity for phosphate. The low-affinity (high phosphorus concentration) transporter (high phosphorus concentration) is energetically efficient, while the high-affinity transporter (low phosphorus concentration) is less efficient. This finding is highly significant for the use of EBPR for achieving very low dissolved phosphate concentrations (Jenkins, 2007). It is necessary to reduce the dissolved phosphate concentration to low values as early as possible in the aerobic zone, so that the high-affinity transporter can operate over as much of the aerobic zone as possible. Narayanan et al. (2006) found that it was possible to consistently achieve dissolved orthophosphate concentrations of less than 0.1 mg/L, by staging the aerobic zone and by making conditions favorable for rapid dissolved orthophosphate uptake (adequate dissolved oxygen) in the first stage. The Future of Biological Process Population Dynamics Research

We are on the cusp of developing a far better understanding of the mechanisms and capabilities of biological processes through the intelligent combination of MBTs and process engineering. Here, in conclusion, are some areas that may prove fruitful.

* Functional structure/and organization of floes. What types of communication and interaction take place between the microorganisms that make up floes, and can we influence these to our advantage? Do microorganisms in floes behave differently from dispersed microorganisms, and, if so, how? Are there important differences in microbial composition and properties between floes and biofilms?

* Mechanism of water retention by floes. What are the detailed chemical/physical nature and properties of biopolymers in floes, especially with respect to their role in water retention? Can these biopolymers find uses in other processes and in other fields?

* Bioaugmentation. Can the use of bioaugmentation be expanded to areas such as the treatment of toxic, poorly biodegradable wastes? Can microorganisms for bioaugmentation benefit from encapsulation? Can nitrite lock be overcome by bioaugmentation?

* Mechanism of microbial end-product formation. What is the nature of the residual dissolved organic matter in secondary effluents? Does its quantity and composition change with operating conditions and process modifications?

* "New" processes (i.e., MBRs, hybrid reactors, and the Cannibal process [US Filter, Warrendale, Pennsylvania]). Because of the different solids separation methods, MBRs tend to contain more dispersed microorganisms and smaller floes than with gravity clarifiers. Does this have any effect on the types and activities of the microorganisms present? Because the membrane in MBRs can retain some of the effluent "dissolved" organics found in traditional activated sludge effluents, do MBRs provide additional degradation of this material, and, if so, by what mechanisms? How do the populations of the fixed media and suspended growth differ and influence each other in hybrid systems?

This list is by no means complete. It is just a sampling of the things that I would be thinking about doing if I were starting my research career now, rather than 50 years ago. I will leave some of these things to those who are doing that now. I wish them (after an appropriate amount of toil) much success in their endeavors. I only hope that they enjoy themselves as much as I did.

Credits

I would like to acknowledge the contributions of the students and colleagues who did the research at the University of California at Berkeley that is mentioned in this paper. They are Jonathon Palm, Oliver Hao, Tony Lau, Peter Strom, Michael Richard, Denny Parker, Gregory Shimizu, Andre van Niekerk, Sang-Eun Lee, Y-J Shao, Ben Koopman, Don Gabb, George Ekama, Kay Johnson, J. B. Neethling, Bernardo Vega-Rodriguez, Paul Pitt, Dan Cha, Linda Blackall, Valter Tandoi, Krishna Pagilla, Daniel Marnais, Abe Fainsod, B. Narayanan, clay Radke, Trina McMahon, Jay Keasling, and the late Clark Weddle and Gerrit Marais.

Glen Daigger, Alex Ekster, Trina MacMahon, B. Narayanan, J. B. Neethling, Bob Okey, Mike Stenstrom, Chi-Chung Tang, and Al Waromski are thanked for critically reviewing a draft of this paper. Sarah Muren prepared the manuscript, and Joan Jenkins verified the citations.

Submitted for publication February 27, 2008; accepted for publication March 4, 2008

The deadline to submit Discussions of this paper is November 15, 2008.

Lawrence E. Peirano Professor Emeritus, Civil and Environmental Engineering, University of California at Berkeley; e-mail: flocdoc@pacbell.net.

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Krist

Source: Water Environment Research

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