Biomethanation I (Advances in Biochemical Engineering & Biotechnology, Volume 81) [1 ed.] 3540443223, 9783540443223

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Preface

Preface

In November 1776, Alessandro Volta performed his classic experiment disturbing the sediment of a shallow lake, collecting the gas and demonstrating that this gas was flammable. The science of Biomethanation was born and, ever since, scientists and engineers have worked at understanding this complex anaerobic biological process and harvesting the valuable methane gas produced during anaerobic decomposition. Two lines of exploitation have developed mainly during the last century: the use of anaerobic digestion for stabilization of sewage sludge, and biogas production from animal manure and/or household waste. Lately, the emphasis has been on the hygienic benefit of anaerobic treatment and its effect on pathogens or other infectious elements. The importance of producing a safe effluent suitable for recirculation to agricultural land has become a task just as important as producing the maximum yield of biogas from a given type of waste. Therefore, anaerobic digestion at elevated temperatures has become the main area of interest and has been growing during the last few years. Anaerobic digestion demands the concerted action of many groups of microbes each performing their special role in the overall degradation process. Both Bacteria and Archaea are involved in the anaerobic process while the importance, if any, of eukaryotic microorganisms outside the rumen environment is still unknown. The basic understanding of the dynamics of the complex microflora was elucidated during the latter part of the last century where the concept of inter-species hydrogen transfer was introduced and tested. The isolation of syntrophic bacteria specialized in oxidation of intermediates such as volatile fatty acids gave strength to the theories. Lately the use of molecular techniques has provided tools for studying the microflora during the biomethanation process in situ. However, until now the main focus has been on probing the dynamic changes of specific groups of microorganisms in anaerobic bioreactors and less emphasis has been devoted to evaluating the specific activities of the different groups of microbes during biomethanation. In the future we can expect that the molecular techniques will be developed to allow more dynamic studies of the action of specific microbes in the over-all process. From the present studies we know that many unknown microbes are found in anaerobic bioreactors. Especially within the domain of Archaea, there are whole phyla of microbes such as the Crenarchaeota, which make up significant fractions of microbes in a reactor but without cultured representatives. Improving the techniques for the isolation of presently unculturable microbes is a major task for the future.

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Preface

Anaerobic digestion of waste has been implemented throughout the world for treatment of wastewater, manure and solid waste and most countries have scientists, engineers and companies engaged in various aspects of this technology. Even though the implementation of anaerobic digestion has moved out of the experimental phase, there is still plenty of room for improvements. The basic understanding of the granulation process, the basis for the immobilization of anaerobic microbes to each other without support material in UASB reactors, is still lacking. Like any other bioprocess, anaerobic digestion needs further control and regulation for optimization. However, until now suitable sensors for direct evaluation of the biological process have been lacking and anaerobic bioreactors have generally been controlled by indirect measurements of biogas or methane production along with measurements of pH and temperature. The newly development of an on-line monitoring system for volatile fatty acids could be a major step in the right direction and the use of infra-red monitoring systems could bring the price down to a reasonable level. A better performance of large-scale anaerobic bioreactor systems for treatment of complex mixtures of waste can be expected to be based on on-line monitoring of the process in the future along with controlling software for qualified management of these plants. Besides treatment of waste, anaerobic digestion possesses a major potential for adding value to other biomass converting processes such as gasification, bioethanol or hydrogen from ligno-cellulosic materials. Conversion of ligno-cellulosic biomass will often leave a large fraction of the raw material untouched which will be a burden for the over-all economy of the process and will demand further treatment.Anaerobic digestion will on the other hand be capable of converting the residues from the primary conversion into valuable methane, which will decrease the cost and the environmental burden of the primary production. Biomethanation is an area in which both basic and applied research is involved. Major new developments will demand that both disciplines work together closely and take advantage of each other’s field of competence. The two volumes on Biomethanation within the series of Advances in Biochemical Engineering and Biotechnology have been constructed with this basic idea in mind and, therefore, both angles have been combined to give a full picture of the area. The first volume is devoted to giving an overview of the more fundamental aspects of anaerobic digestion while the second volume concentrates on some major applications and the potential of using anaerobic processes. The two volumes will therefore be of value for both scientists and practitioners within the field of environmental microbiology, anaerobic biotechnology, and environmental engineering. The general nature of most of the chapters along with the unique combination of new basic knowledge and practical experiences should, in addition, make the books valuable for teaching purposes. The volume editor is indebted to all the authors for their excellent contributions and their devotion and cooperation in preparing these two volumes on Biomethanation. Lyngby, January 2003

Birgitte K. Ahring

CHAPTER 6

Perspectives for Anaerobic Digestion Birgitte K. Ahring University of California, Los Angeles (UCLA), School of Engineering and Applied Science, Civil and Environmental Engineering Dept., 5732 Boelter Hall, Box 951593, Los Angeles, California 90095-1593, USA Present address: Biocentrum, The Technical University of Denmark, DTU, Block 227, 2800 Lyngby, Denmark. E-mail: [email protected]

The modern society generates large amounts of waste that represent a tremendous threat to the environment and human and animal health. To prevent and control this, a range of different waste treatment and disposal methods are used. The choice of method must always be based on maximum safety, minimum environmental impact and, as far as possible, on valorization of the waste and final recycling of the end products. One of the main trends of today’s waste management policies is to reduce the stream of waste going to landfills and to recycle the organic material and the plant nutrients back to the soil.Anaerobic digestion (AD) is one way of achieving this goal and it will, furthermore, reduce energy consumption or may even be net energy producing. This chapter aims at provide a basic understanding of the world in which anaerobic digestion is operating today. The newest process developments as well as future perspectives will be discussed. Keywords. Anaerobic digestion, Carbon-flow, Microbiology, Antimization, Gas yild, Effluent

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Introduction

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Microbiology of Anaerobic Digestion . . . . . . . . . . . . . . . .

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General Scheme . . . . . . . . . . . . . . . . . . . . . . . . . . . Syntrophic Acetate Conversion . . . . . . . . . . . . . . . . . . . Microbiology of Thermophilic Digestion . . . . . . . . . . . . . Establishing a Stable Microflora in Thermophilic Reactors . . .

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Anaerobic Digestion as a Way to Add Extra Value

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Optimization of Anaerobic Digestion . . . . . . . . . . . . . . . . 15

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Increasing the Digestibility of the Waste . . . . . Optimization of Reactor Configuration . . . . . . Optimizing Process Control and Stability . . . . . Improving the Microbial Process and its Efficiency

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Advances in Biochemical Engineering/ Biotechnology, Vol. 81 Series Editor: T. Scheper © Springer-Verlag Berlin Heidelberg 2003

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Optimization of Effluent Quality

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Inactivation of Pathogens and Other Biological Hazards . . . . . . 23 Control of Chemical Pollutants . . . . . . . . . . . . . . . . . . . 25

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Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

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References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27

1 Introduction The modern society generates large amounts of waste that represent a tremendous threat to the environment and human and animal health. To prevent and control this, a range of different waste treatment and disposal methods is used. The choice of method must always be based on maximum safety, minimum environmental impact and, as far as possible, on valorization of the waste and final recycling of the end products. One of the main trends of today’s waste management policies is to reduce the stream of waste going to landfills and to recycle the organic material and the plant nutrients back to soil. Waste is increasingly becoming a problem and secure recirculation is gaining more and more attention. Anaerobic digestion (AD) is one way of achieving this goal and it will, furthermore, reduce energy consumption, or may even be energy producing, which is of major importance to the global environment. Anaerobic digestion has been implemented for years as a means for the stabilization of sewage sludge; however, during the past years anaerobic digestion technologies have been expanded to emphasize treatment and energy recovery from many other types of wastes including animal wastes, source-sorted household wastes, organic industrial wastes and industrial wastewater. Compared to incineration, anaerobic digestion creates more energy during the treatment of wastes, which normally have high water content. During incineration the nutrients are lost. Following the increasing interest in implementation of anaerobic digestion, optimization of this process is becoming increasingly more important. Despite the increased efforts spent on waste reduction, the amounts of waste are increasing throughout the world. This has led to ideas for a total removal of waste through injection into the deep underground (below 2 km) into old oil wells far below any the groundwater level [1]. The recovery of methane will, however, be of importance for the feasibility and economy of this technique and methane development at these high temperatures, pressures and salinity is now under investigation. This chapter focuses on the perspectives for optimization of anaerobic digestion after a brief introduction to the microbiology of anaerobic digestion. Optimization is a double-sided task: it involves both an increase of the biogas yield, which again implies an increased removal of the organic material in the waste, as well as ensuring an effluent with a sufficiently high quality to allow for recycling of the material as a fertilizer. A number of areas for improving the biogas yield will be discussed such as, for example, increasing the digestibility of the

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waste, optimizing process control, and improving the microbial process. With respect to effluent quality, emphasis will be on inactivation of pathogens and control of chemical pollutants.

2 Microbiology of Anaerobic Digestion 2.1 General Scheme

A major value of anaerobic digestion is linked to the production of biogas (methane and carbon dioxide) formed as the end product during degradation of organic material without oxygen. This energy is renewable and CO2 neutral and can be used for production of electricity and heat. Many different consortia of microorganisms with different roles in the overall process scheme are needed for the AD process, which occurs naturally in anaerobic ecosystems such as sediments, paddy fields, water-logged soils and in the rumen [2]. Three major groups of microorganisms have been identified with different functions in the overall degradation process [3] (Fig. 1): 1. The hydrolyzing and fermenting microorganisms are responsible for the initial attack on polymers and monomers found in the waste material and produce mainly acetate and hydrogen, but also varying amounts of volatile fatty acids (VFA) such as propionate and butyrate as well as some alcohols.

Fig. 1. The anaerobic degradation process

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Fig. 2. Carbon flow in anaerobic environments with active methanogens

Fig. 3. Carbon flow in anaerobic environments without active methanogens

2. The obligate hydrogen-producing acetogenic bacteria convert propionate and butyrate into acetate and hydrogen. 3. Two groups of methanogenic Archaea produce methane from acetate or hydrogen, respectively. The major part of the carbon flow in a well-operating anaerobic reactor occurs between the fermentative microorganisms and the methanogens. Only between 20 and 30 % of the carbon is transformed into intermediary products before these are metabolized to methane and carbon dioxide (Fig. 2) [4]. A balanced anaerobic digestion process demands that the products from the first two groups of microbes responsible for hydrolyzing and fermenting the material to hydrogen and acetate, simultaneously are used by the third group of microbes for the production of methane and carbon dioxide. The first group of microorganisms can survive without the presence of methanogens but will, under these conditions, form an increased amount of reduced products such as VFA (Fig. 3). The second group does, however, rely on the activity of the methanogens for removing hydrogen to make their metabolism thermodynamically possible as their reactions are endergonic under standard conditions and only occur when hydrogen is kept below a certain concentration. The relationship between the VFA-degrading bacteria and the hydrogen-utilizing methanogens is defined as syntrophic due to the dependent nature of this relationship and the process is

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Fig. 4. Interspecies hydrogen transfer

called interspecies hydrogen transfer (Fig. 4) [3]. The lower the hydrogen concentration the better are the thermodynamics of the VFA degradation. The distance between the VFA degrader and the hydrogen utilizer does, therefore, affect the concentration of hydrogen in the liquid phase, which again affects the thermodynamics of the process. Therefore, the conversion is improved in granules and flocks compared to a situation where the microbes are distributed freely in a liquid solution [5]. The two partners have to share a very small amount of energy and the conditions for ensuring energy for both microbes is very strict and can only be met within a narrow range of hydrogen concentrations [6]. 2.2 Syntrophic Acetate Conversion

Syntrophic relationships have also been found to be of importance for conversion of acetate when the acetate-degrading methanogens are inhibited by high concentrations of ammonia [7, 8] or sulfite (unpublished). Under these conditions the acetate-utilizing methanogens are inhibited and other groups of microbes replace them to obtain energy from the oxidation of acetate to hydrogen and carbon dioxide (Fig. 5). Due to thermodynamic constrains this reaction proceeds much better at increased temperatures and is the way of acetate transformation when the temperature is higher than 60°C, close to the upper temperature limit of thermophilic acetate-utilizing methanogens [9, 10]. In accordance with this, the population of Methanosarcina species disappeared more or less instantaneously from a biogas reactor operated on manure, when the temperature was increased from 55 to 65°C [11]. Concurrently, the acetate concentration first increased and

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Fig. 5. Modified anaerobic degradation process with syntrophic acetate conversion

then stabilized at a level somewhat higher than that found at 60°C [12]. This coincided with a significant increase in the population of hydrogen-utilizing methanogens [11] indicating that this group had become dominant in the overall conversion. Both syntrophic acetate oxidation and methanogenesis from acetate can be simultaneously active in a reactor system as indicated by several isotope studies often showing that less than 95 % of the methane produced from acetate is derived from the methyl group. Isotope experiments with biomass from thermophilic reactors have further shown that the concentration of acetate affects the competition between the two processes. When the concentration of acetate is low, syntrophic acetate conversion is the major process for acetate transformation [13, 14]. However, when the concentration of acetate is above the threshold level [15] for the specific population of acetate-utilizing methanogens in the reactor, these will be the major group active in the system. These findings further explain why the numbers of hydrogen-utilizing methanogens are high in thermophilic granules, which have exclusively been fed with acetate for a long period [16]. Furthermore, the numbers of acetate-utilizing methanogens are highest close to the surface of the granules, where the concentration of acetate is highest, while the populations of hydrogen-utilizing methanogens increased towards the center of the granules [16]. The first microbe found to perform acetate oxidation was a thermophilic bacterium belonging to the group of homo-acetogenic bacteria capable of reversing the acetate-forming reaction from hydrogen and carbon dioxide [17]. This bacterium used a very limited range of substrates all related to its homo-acetogenic nature [17]. Over time more microbes have been identified as being capable of

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carrying out this reaction. Some of these microbes have been found to use a large variety of substrates [18] and, furthermore, to be normal members of the populations of fermentative microbes in thermophilic reactors. This indicates that, at least in thermophilic reactors, syntrophilic acetate oxidation could be performed by a variety of the fermentative bacteria in the reactor when no other substrates are available. This needs, however, further verification. 2.3 Microbiology of Thermophilic Digestion

Microbes thriving at high temperatures have been known for years [19]. The reaction rate of many chemical reactions will double by an increase of 10°C according to the Arrhenius equation. The same is, however, not always the case for microbial reactions where the temperature response is specific for the particular microbe. Different groups of microbes have been identified where the ones of interest for anaerobic digestion are mesophilic strains with an optimum between 30 and 40°C, and thermophilic strains with an optimum between 50 and 60 °C [20]. The mixed microflora found in an anaerobic bioreactor generally shows an increasing rate from a temperature of 20 to 60 °C and the theoretical temperature gap between mesophilic and thermophilic strains is not apparent when viewing the process as a whole [21]. Anaerobic digestion at a temperature below 20°C, or at a temperature above 60 °C, generally shows a lower methane yield than within these limits. However, anaerobic digestion has been shown to be possible even at extreme thermophilic conditions of 70 °C and more [28 – 30]. Experiments with high temperature digestion of manure showed that major changes occurred in the microbial populations of the anaerobic reactor when the temperature was increased from 55 to 65 °C [12]. Besides a significant increase in the population of Archaea compared to Bacteria, also the populations of methanogens underwent large changes over time. The population of hydrogen-utilizing methanogens did, for example, change from a major population belonging to the genus Methanobacterium to another belonging to Methanococcus over a 3-month period [11]. Such results clearly demonstrate that reactors operated at extreme conditions can take months before a stable microflora has established. With this in mind it is difficult to guess if the methane yield actually will be lower after an extended period of many months. Within the normal temperature range the general carbon flow of thermophilic reactors was found to be very similar to that of mesophilic reactors [31]. A slightly higher amount of the carbon was channeled directly into acetate and a slightly smaller amount of carbon was turned over via the pool of VFA [4]. Many extreme thermophilic Bacteria or Archaea have been found to produce mainly acetate and hydrogen as their end products [32]. Therefore, less butyrate and propionate can be expected at these high temperatures. Different maximum temperatures were found for the different microbial groups in a thermophilic anaerobic reactor treating manure [33]. For instance, among the methanogens, the acetate-utilizing methanogens have a much lower temperature maximum (ca. 62 °C) compared to the hydrogen-utilizing methanogens (ca. 75 °C) [33]. However, the actual temperature of the reactor affects the specific populations

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which are active in the reactor. Therefore, a higher temperature optimum and maximum is found for the main metabolic groups in extreme thermophilic reactors compared to thermophilic reactors [28]. Methane production was found in microbial mat samples taken from a slightly alkaline hot spring at 80 °C [34]. This demonstrates that methanogenesis is possible even at this very high temperature. 2.4 Establishing a Stable Microflora in Thermophilic Reactors

Waste such as sewage sludge, manure or household waste contains many different populations of anaerobic or facultative anaerobic microorganisms. Most of these microbes are mesophilic and only a very small number of true thermophiles is present. The number of microbes in raw sewage sludge utilizing substrates such as acetate or cellulose at 60 °C is extremely low (ca. 100 per g) [35]. The numbers are somewhat higher at 55 °C but still much lower than the numbers at 37 °C [35]. These facts clearly show the problems of establishing stable reactors at higher temperatures. Where the microflora of mesophilic reactors can be established directly based on the raw material fed to the reactor, the microflora of the thermophilic reactor has to be propagated from small minority populations found in the raw materials [36]. Many thermophilic full-scale reactors have failed through history, especially within the area of sewage sludge treatment. The reason is basically a lack of understanding of the principles for establishing a stable thermophilic microflora in the reactor. The same also applies to the literature, which is full of experiments with unstable thermophilic laboratory reactors often performing poorly compared to mesophilic reactors. When reviewing the literature describing these experiments, Wiegant [37] concluded that process stability is lower in thermophilic reactors and that thermophilic reactors generally have higher concentrations of volatile fatty acids in the effluent compared to mesophilic reactors. During recent years where more thermophilic reactors have been implemented, it has been shown that this conclusion it not correct and that stable thermophilic reactors with a balanced thermophilic microflora perform just as well as stable mesophilic reactors [33, 38, 39]. The key to obtain a balanced thermophilic microflora is to give optimal growth conditions to the small numbers of thermophilic populations found in the raw material during start-up of the bioreactor [36]. If sufficient thermophilic seed material is available, it is possible to carry out a rapid start-up of a thermophilic reactor [33]. The seed material should be evaluated before use with respect to the destruction of volatile solids in the reactor from which the seed is obtained as well as the concentration of VFA. If possible, it will be beneficial to perform a methanogenic activity testing of the seed material to establish the potential of this seed for transforming extra loads of methanogenic substrates (acetate and hydrogen) [40, 41]. After addition of the seed to an empty reactor it should be allowed to equilibrate for 1 day before feeding is initiated at the desired thermophilic temperature. A slow and graduate change of the temperature only prolongs the start-up phase and does not select for true thermophiles

Fig. 6. Start-up using thermophilic digested seed material

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24 h

t d = 24 h

24 h

td = 8 h

Fig. 7. Unbalanced growth of methanogens (dots) and fermentative bacteria (rods)

possessing an optimum growth rate at thermophilic temperatures. Probing of methanogens in bioreactors has clearly demonstrated that mesophilic methanogens are present in thermophilic reactors and vice versa [42]. However, the major populations are those with an optimum temperature close to the reactor temperature [43]. After equilibration, feeding should then be initiated corresponding to a hydraulic retention time approx. 25 % higher than the retention time of the reactor where the seed came from. Normally the reactor is only partly full at this stage and, therefore, the reactor is operated in a fed-batch mode during this period of time. During start-up, the VFA concentration should be monitored on a daily basis. If the VFA concentrations continue to decrease after approx. 3 days of feeding or remain at a stable low level, the hydraulic retention time can be lowered. By repeating this pattern and, at the same time, keeping a tight eye on the concentration of VFA – especially the isoacids [44, 45] – it is possible to reach the desired final retention time in approx. 1 month. A schematic drawing of the expectable feeding pattern and the expectable response in VFA is shown in Fig. 6. For sewage sludge it is possible to obtain a stable process with a high reduction of volatile solids at a hydraulic retention time as low as 6 days at thermophilic conditions. If no seed is available, it is even more important to plan the start-up in a controlled condition. It is important to avoid over-loading and build-up of VFA. Therefore, a start-up material containing only small amounts of organic material should be chosen. Mesophilic digested waste material has a much lower organic content than raw waste material and has at least as many thermophilic microorganisms as found in this material [46]. Immediately upon an increase of the temperature to the thermophilic region, these thermophilic microbes will start growing. As the acid-producing microbes grow much faster than the

Fig. 8. Start-up using mesophilic digested seed material. Arrows indicate feeding of the reactor

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methanogens, the first reaction is often an increase in the concentration of VFA [47] (Fig. 7). Due to the limited amount of undigested material present, this increase is only relatively small and does not affect the over-all digestion process. Immediately after a decreasing trend is seen in the volatile fatty acids corresponding to growth of the population of VFA-degrading and methanogenic microbes, it is appropriate to start feeding.A portion corresponding to approximately the double of the desired hydraulic retention time is appropriate. Depending on the response in VFA concentration, this trend can be continued every day unless the VFA starts to increase. After 3 to 5 days of continuous feeding it is time to lower the hydraulic retention time again and this pattern can be continued until the reactor has reached the desired hydraulic retention time. This is normally reached within a period of 2 months. A schematic drawing of the expectable feeding pattern and the expectable response in VFA is shown in Fig. 8. The time needed for performing the start-up with a small amount of thermophilic seed material can further be reduced by addition of mesophilic-digested material in addition to the daily feeding with raw waste material. This was used with success for start-up of the new thermophilic sewage sludge digester at Western Lake Superior Sanitary District in Duluth, Minnesota during the summer of 2001 [48].

3 Anaerobic Digestion Plants A large number of different AD-technologies and AD plants are found today throughout the world. The largest number of AD plants in the modern society treats primary and secondary sludge (biosolid) in municipal wastewater treatment plants. These plants basically stabilize the waste material and the biogas produced is often of minor importance. For some of the large wastewater plants, the biogas produced is used for electricity production and the idea of improving the biogas yield is attracting increased interest [49, 50]. A large number of single household biogas units have further been implemented in developing countries such as China, India and Africa [51, 52]. These units will normally provide gas for cooking and lighting in the households. Another major field for anaerobic digestion is the industrial wastewater from, especially, food processing industries where the wastewater is heavily polluted with easily degradable organic carbon [53]. Treatment of municipal wastewater has further been implemented in developing countries such as India especially where the average temperature is rather constant [54]. Anaerobic treatment is basically a way to reduce BOD while the nutrients such as nitrogen and phosphor are left untouched [50]. Recent studies have, however, shown that nitrogen can be denitrified in a chemoautotrophic anaerobic process using nitrite as an electron acceptor [55, 56]. A better way to implement anaerobic treatment of municipal wastewater would be to recover all nutrients and heavy metals from the wastewater after the anaerobic treatment using membrane technology. In this way, the benefits of anaerobic process with respect to space, speed, low sludge production and cost can be fully exploited and the valuable nutrients can be reused.

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A large-scale biogas facility for treating manure from several farms in combination with other organic wastes such as food wastes and source-sorted household wastes – the so-called co-digestion – was launched in Denmark at the end of the 1980s [57, 58]. Addition of even small amounts of organic industrial wastes increases the gas production significantly (Fig. 9). Especially fatty or oily wastes have a much higher gas potential than manure and a much higher concentration of organic material (higher dry matter content) – but also wastes rich in carbohydrates and proteins will improve the gas yield per unit of reactor volume. Digestion of sewage sludge or manure yields from about 1 – 2 cubic meter biogas per cubic meter reactor volume per day while the reactors will produce between 4 and 10 cubic meter biogas per cubic meter reactor volume with addition of ca. 20 % fatty waste. Today around 22 large-scale AD plants have been built in Denmark mainly in the regions with high manure production and all of these plants are co-digesting many types of raw materials [33, 59]. The idea of large scale centralized AD plants treating mixtures of waste have spread throughout Europe and to the rest of the world especially during the last ten years [60]. Besides common biogas plants, the numbers of farm biogas plants for large pig farms have steadily increased in many European countries [61]. Many of these plants further supplement the manure with raw materials with a higher gas potential. Recently, a large number of biogas projects are on their way in USA [62] and especially in California due to the energy crisis starting at the end of year 2000, which again

Fig. 9. Addition of 5 % fish oil will double the biogas production

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has resulted in higher prices on electricity. Biogas plants in USA are often much larger compared to the so-called “large-scale” biogas plants in Denmark and treat manure from diaries or feeding lots having up to hundred thousand cows or from major pig or chicken production. The AD technology is, therefore, suitable both for small and large-scale applications. The economics depend upon the scale, and larger plants will in general have a better economy than smaller plants. Raw sewage sludge has a very low dry matter content and, therefore, the potential for treating waste in these plants is only used to a low degree. Addition of food waste, restaurant waste or organic industrial waste could be a good way to make use of this potential. Several concepts are based on treatment of mixtures of sewage sludge in combination with household waste such as the Finish Waasa process [63]. Very good results have been obtained in Grindsted, Denmark with co-digestion of source-sorted food waste together with sewage sludge. The food waste was collected in paper bags and only a small amount was removed before the digestion process [65].

4 Anaerobic Digestion as a Way to Add Extra Value Production of biofuels from biomass such as bioethanol or gasification of biomass only makes use of a fraction of the biomass. The same is true for many other biomass-based productions of non-food products. The biomass fraction left is, however, often a good substrate for methane production (Fig. 10). In this way biogas production can add approximately 30 % more value to the production of bioethanol from biomass such as wet straw or corn stovers [66]. The AD process will further purify the process water allowing for recirculation within the system, which will further decrease the cost of ethanol production. In the future it is expected that more valuable products than methane will be sought from waste. However, these niche productions will as a rule only use a part of the waste and methane production from the final residues can add further value to the production and will decrease the pollution load of the end products before their final disposal.

Fig. 10. Simultaneous production of bioethanol and biogas

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5 Optimization of Anaerobic Digestion The economy of a biogas plant is directly linked to the amount of biogas produced per unit of raw material treated in the plant. Some costs are fixed such as the cost of transporting the material to the AD plant and back again to the end user or the end destination, while others are variable such as construction costs. Lowering the water content of the raw material and running the process with higher dry matter content can significantly decrease the cost of treatment. This is of major importance when the raw material has to be transported to a centralized biogas facility and in this case it is often beneficial to separate the manure into a solid fraction and a liquid fraction, which is left behind at the farm. The liquid fraction can be used as a nitrogen-rich fertilizer at the form. The potential for increasing the biogas yield of manure or sewage sludge is larger as only approximately half of the organic material is converted in this type of material. This is, however, not the case for most organic industrial wastes or source-sorted household waste, which have been found to be more easily degradable, and approximately 80 % of the organic material is converted to biogas [67, 68]. Manure is, however, the major raw material available for a largescale use of AD technology in most of the world and a large-scale implementation of AD will have to be based on this raw material. AD will further improve the quality of manure by making a more stable material with fewer pathogens and less odor. For wastewater sludge the interest in increasing the conversion of the organic material is further linked to the reduction in the final amount of biosolid, which has to be disposed after the treatment. Suitable end-use of digested sewage sludge or biosolids is becoming an increasing problem for many communities throughout the world. Some major ways to improve the gas yield in AD plants will be by (Fig. 11): 1. 2. 3. 4.

Increasing the digestibility of the waste, Optimizing the reactor configuration, Optimizing process control and stability, and Improving the microbial process and its efficiency

5.1 Increasing the Digestibility of the Waste

Several methods have been discovered to increase the digestibility of manure or sewage sludge ranging from mechanical, chemical to biological methods such as enzyme treatment. Chemical treatment with bases or acids or treatment with mixtures of enzymes have generally been found to increase the accessibility to microbial conversion into biogas – but these processes have all been found to be too expensive for practical implementation [69]. Decreasing the particle size was found to increase the gas production from manure and the increase in gas production exceeded by far the extra costs of implementing a macerating unit with several knives [70].

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More gas Sewage

Anaerobic digestion Animal wastes

Biogenic wastes

A. Optimization of biogas production

• Increasing the digestibility

• Optimizing reactor

Source-sorted household wastes

configuration

• Optimizing process control

• Improving microbial process

Food-processing wastes

B. Optimization of effluent quality

• Inactivation of Crops

Fertilizer

pathogens

• Control of chemical pollutants

Fig. 11. Major goals for anaerobic digestion of today

A handful of wastewater plants in the world have further implemented the use of thermal hydrolysis where concentrated sludge is treated by a combination of high temperature (133 – 180 °C) and pressure (3 –10 bar) with the aim of improving the digestibility of the sludge (The Cambi Process) [71]. However, the influence of this process on gas production per unit sludge treated is still not fully documented, and the amount of sludge ending as carbon dioxide due to the treatment is unknown. The wet oxidation process using alkaline conditions and oxygen in addition to high temperature and pressure has been found to be superior for breaking the lignin associated to hemicellulose and cellulose as lignocelluloses [72]. The products of the lignin-oxidation (carboxylic acids and phenolics) are further found to have highly convertible to methane and carbon dioxide (approx. 80 % COD removal) [73]. The pure cellulose and hemicellulose found in the hydrolysate is expected to give a methane yield corresponding to the methane potential of the mannouronic sugars. Due to the hydrolytic capability of microbes in the AD process, it is expected that enzyme addition is not needed for conversion of hydrolysates produced by wet oxidation. However, this still needs to be verified along with the optimal way of implementing wet oxidation as part of the AD process for materials such as manure containing a high fraction of lignocellulosic material. Furthermore, the economics of this extra step needs to be evaluated. Another way to enhance the digestibility of the raw material is by Pulse PowerTM technology developed by Scientific Utilization Inc. in Decatur, AL [74]. This equipment incorporates rapid-pulse high-power electric technology ori-

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ginally developed for antimissile laser and particle-beam devices, which produces disruptive shock waves in the raw material. The shocks are expected to break large molecules into shorter fractions and have been claimed to enhance destruction of volatile solids by 50 to 100 %. However, a study of the efficiency of this method in a full-scale system showed no improvements. 5.2 Optimization of Reactor Configuration

The AD process can be conducted in a single-step or multi-step process [47]. Continuous processes are generally most favorable when treating large amounts of waste and thermophilic temperatures have the largest potential due to higher reaction rates which corresponds to smaller reactor volumes. Separation of the solid phase from the liquid phase of manure or sewage sludge is a technical solution which has been well documented during the last years and which can be implemented both before and after the anaerobic reactor [75 – 77]. Separation will further allow for optimal design of the process so that the liquid fraction can be left at the farm, or treated locally in small compact plants, while the solid faction can be transported to a centralized plant for treatment. Especially pig manure contains very high concentrations of phosphorus and, therefore, large land areas are needed to use the manure as a fertilizer afterwards. If the solid fraction is removed from the manure, the farm is able to use the liquid fraction on a much smaller land area and pipes can be used for the spreading. After digestion the phosphorus-rich solid fraction is an excellent fertilizer [78]. If the separation is carried out on fresh manure approximately 70 % of the gas potential will remain in the solids [70]. Recently, we demonstrated that the conversion of organic material in manure could be increased along with an increase in the over-all biogas yield by using a two-phase system combining a short hydrolysis step performed at 70 °C followed by a methane-producing step at 55 °C, both done in continuous stirred tank reactors (CSTRs) (Fig. 12). The performance was compared to a singlephase process in a CSTR reactor with the same over-all retention time, and the first estimation showed that the extra gas produced was sufficient to justify the implementation of an additional reactor and the need for extra heating energy (unpublished). The possibility of using an immobilized reactor system after a short hydrolytic step during a two-phase conversion of waste such as manure, sewage sludge and household waste does possess a potential, which needs more attention (Fig. 11).A number of systems such as the up-flow anaerobic sludge blanket reactor are in use throughout the world for treatment of industrial wastewaters. However, only limited experience has been obtained from full-scale use of this reactor for the treatment of solids [79, 80]. Having retention times in the range of hours, the potential is apparent, as much smaller reactors are needed to treat the same amount of waste. Furthermore, the immobilized system has lower construction and running costs, no stirring system will be needed. The sludge produced in the system can be recirculated back to the hydrolysis step decreasing the final sludge production from the system, and again the phosphorus is concentrated in the sol-

Fig. 12. Different reactor configurations for anaerobic digestion of waste

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id fertilizer while the liquid fertilizer is rich in nitrogen. The over-all economy is, therefore, improved although a separation is needed between the hydrolysis reactor and the immobilized reactor to ensure that the amount of suspended solids in the influent to the immobilized system is acceptable. 5.3 Optimizing Process Control and Stability

Process stability is important for the operation and economy of any AD plant. Imbalance often affects the methanogens in the anaerobic process and leads to a VFA accumulation [44]. It is important to note that some inhibitory compounds equally affect all the major groups in the anaerobic digestion process. This is the case for long-chain fatty acids [81, 82] or for phthalate esters such as DEHP [83]. No VFA accumulation was observed when reactors were inhibited with DEHP. Inhibitory compounds in waste are, however, generally either ammonia or sulfide, which are found in high concentrations in some types of waste [84–87]. Furthermore, high concentrations of proteins in the incoming waste can lead to the development of inhibitory concentrations of ammonia and sulfide. For both of these compounds, the toxic effect is dependent on pH and temperature – the higher the temperature and pH, the higher the toxic effect [88]. Due to the high ammonia concentration, thermophilic digestion of swine manure has been found to be difficult [89]. Adaptation to an inhibitory compound is, however, possible over time and the anaerobic process can work with stable performance but with a lower gas yield as long as the concentration of the toxic compound is kept relatively constant. Process stability is, however, lost if the concentration of the inhibitor is fluctuating as seen in the large-scale biogas plants when treating many types of wastes in different ratios. In immobilized anaerobic systems, the biomass has generally been found capable of withstanding much higher concentrations of inhibitory compounds [90]. This is probably due to concentration gradients in the biofilm creating niches where the microbes are protected from toxic concentrations of the inhibitory compounds. The use of a two-phase digestion system is, therefore, expected to show superior performance by compared to a one-phase system for waste containing high concentrations of inhibitors. Increasing the biomass concentration in a biogas reactor by recirculation of the biomass was found to increase the gas yield during anaerobic digestion of swine waste [91, 92]. In accordance with this, the inhibitory effect of swine manure can be counter-acted by addition of other wastes such as lipid-containing wastes, which result in a higher biomass concentration in the reactor besides a dilution of the manure. Process problems in AD plants are generally difficult to detect before the process is severely affected and the gas production decreases. In general the plant operator has very little information about the condition of the process and no instruments inform him when the process is becoming unstable (Fig. 13). As a result, the plant is often operated with a very low organic loading to prevent problems from occurring. A good sensor would, however, make it possible to optimize the operation and to ensure maximum use of the reactor space without having any process failures (Fig. 14).

Fig. 13. Operation of today’s biogas plant

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Fig. 14. Future operation

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Recently, a sensor has been developed that can measure VFA on-line in biogas reactors [93]. This development allows a continuous monitoring of the anaerobic process and with the development of logic control systems it should be possible to improve the economy of AD plants through a more stable and optimized operation in the future. A number of kinetic models have been developed for anaerobic waste reactors but so far no control algorithm has been developed and tested based on VFA data in addition to the normal data available at the AD plant (flow rates, amount of raw materials, gas production, temperature etc.). A combined sensor and control system can be expected in the future. 5.4 Improving the Microbial Process and its Efficiency

Improving bioprocesses by implementation of microbial populations with improved degradation abilities (bioaugmentation) has been known for years. However, only few studies have been done on bioaugmentation and the results are inconsistent. To obtain a clear picture of the potential to use specific microbes for improvement of the process, it is necessary to follow the fate of the microbes added to the reactor system over time. Only microbes with the ability to thrive and proliferate in the reactor will be of importance in a longer-term prospective. Molecular techniques are available today for studies of populations in reactor systems and using such techniques we demonstrated that a specific cellulolytic bacterium, present in manure inhabited the reactor [94]. The same technique could be used to test for specific added microbes. Compared to controls without any pretreatment a more than 20 % increase in the methane yield was found by incubating separated fibers from cow manure with specific extreme thermophilic xylanolytic microbes for 2 days before the material was resuspended in the liquid and digested [69]. This finding seems promising in the context of the two-phase system described above and deserves further examination. Isolation and characterization of the acetate-utilizing methanogens from thermophilic manure plants in Denmark showed important differences between the different isolates of Methanosarcina species with respect to temperature optimum and growth rates [95]. The strain derived from the best performing thermophilic biogas plant was the acetate-utilizing methanogen with the highest growth rate and highest temperature optimum.When using this methanogen to seed reactors where the organic loading was increased by a sudden addition of lipids to the feed of manure, the seeded reactor was found to be superior to overcome the changes compared to the unseeded reactor, which was inhibited severely and accumulated VFA. No major accumulation of VFA was found in the seeded reactor compared to the unseeded reactor, and biomass from this reactor had a much higher specific methanogenic activity on acetate than for hydrogen and formate, which was almost the same in both reactors. The fact that seeding had an effect even after several retention times indicates that the added methanogen grew in the reactor as further demonstrated using a probe specific to this strain (unpublished). These findings have practical implications and

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show that better performance can be obtained when lipid-containing waste is introduced into a biogas reactor operated on manure if the reactor is seeded with a robust acetate-utiling methanogenic strain with a higher growth rate than the native strain in the system.

6 Optimization of Effluent Quality Besides production of biogas the AD plant produces a residue or effluent with a potential market value. Until now, no applicable standards for these products have been available and recycling of AD-residues has generally been poorly regulated in most countries. The main issues related to quality management when recycling AD-residues is 1. to break the chain of disease transmission by inactivation of pathogens and other biological hazards and 2. to control chemical pollutants (organic and inorganic). Inactivation of pathogens is increased with increasing temperatures and, therefore, thermophilic digestion has a much high sanitary effect than mesophilic digestion. 6.1 Inactivation of Pathogens and Other Biological Hazards

Sewage sludge and segregated household wastes are both high-risk wastes that can be heavily contaminated with pathogens [96]. Several reviews on pathogens in livestock waste, factors influencing microbial movement and methods for inactivating pathogens have been published [97 – 99] New regulations with respect to concentrations of pathogens and organic pollutants could potentially be threatening to land-disposal of digested material. The regulations made both by the US EPA and the European Union demand specific treatment processes before the use of sewage sludge on agricultural land [1, 100]. However, for unrestricted use of digested sewage sludge a further reduction of pathogens will be required. Several studies found anaerobic digestion to be superior to aerobic digestion in reducing the density level of pathogens [101]. Conventional mesophilic digestion was found to be insufficient for meeting the new requirements for unrestricted land-use [101, 102]. An increased elimination of pathogens can be achieved by using different treatment processes including composting of sewage sludge at high temperatures, exposing the material to radiation or specific temperatures for a defined period. Thermophilic digestion is still not included as a mean for producing a clean effluent. Biosolid A demands that the number of coliforms is less than 1000 per g dry solid, the number of Salmonella is less than 3 per 4 g dry solid, enteric virus and helmic ova should not be detectable. These restrictions are based on logarithmic reduction experiments performed in buffer with the respective microbes. However, experience obtained under anaerobic digestion showed that other parameters than time and temperature are of importance for reduction of

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pathogens. The anaerobic environment seems to have an additional effect, which is not accounted for in the biosolid classification [102]. It was demonstrated that the high level of VFA [103], ammonia and sulfide [104] and alkaline pH enhance inactivation of pathogens [105]. Several reports on thermophilic AD of sewage sludge showed that thermophilic digestion is more efficient in reducing the pathogens and pathogen indicators than mesophilic digestion [102, 106 –108]. Further improvement of the process by extension of one digester into a series of several reactors operating at 55 °C [110], application of a two-stage process with an acidogenic reactor operating at 55 °C and 60 °C coupled to a 37 °C methanogenic reactor [111], or combinations of thermophilic pretreatment or posttreatment at 62 °C with conventional mesophilic digestion [109] all showed that it is possible to make a stabilized sludge fulfilling the requirements for Biosolid A. Recent experiments done at Terminal Island Treatment Plant in Los Angeles clearly showed that Class A biosolid can be produced by switching the process from mesophilic digestion at around 35 °C to thermophilic digestion temperatures at 55 °C [112]. However, for unrestricted use of the effluent a required holding period of one day is needed at 55 °C, which is difficult to meet in a conventional sewage sludge treatment system. By testing the effluent quality and demonstrating that the solid meets all requirements it is possible to obtain a Class A biosolid classification. The testing program, however, has to be repeated on a regular basis to maintain the classification. Pathogens are further of major interest when manure from several farms is treated in centralized large-scale biogas plants. When the Danish Action Program for Large Scale Biogas Plants was implemented more than 10 years ago, hygienic aspects were central as a consequence of transperation of manure from several farms. Therefore, a veterinary program was initiated and this led to implementation of a number of control functions, which have gained major interest and respect throughout the world [96, 113]. In this program the fecal Streptococci (enterococci) were found to be excellent indicator organisms instead of coliforms (the FS method) during digestion at temperatures up to 60 °C. These microbes are present in manure or other materials of intestinal origin as the coliforms, but in contrast to coliforms, they are much more resistant to high temperatures and the anaerobic environments and, therefore, most pathogenic bacteria, viruses and parasitic eggs will be inactivated long before these microbes. An FS log10 reduction of around 4 and 5 was needed to give an acceptable effluent quality, which basically implies that the AD process is operated at thermophilic conditions or that a high-temperature step is added to a mesophilic reactor [96]. Pathogenic viruses have been identified in sewage sludge, segregated household waste and manure [96]. The absence of enteric viruses showed no correlation with porcine parvovirus in a previous study of thermophilic anaerobic digestion of manure [102]. This indicates that this virus could be a poor indicator for human pathogenic viruses. The effect of the AD process on viruses further depends on the way the virus is found the environment. For instance, it was found that a virus in tissue was less sensitive than a free-living virus [114]. Much more work is, however, needed to understand the fate of a virus during anaero-

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bic digestion including the possibility to remove a resistant virus by pretreatment such as thermal hydrolysis or wet oxidation. Besides the potential for pathogens, material of animal origin such as waste from slaughterhouses and milled bones from cows and sheep can contain infectious elements resulting in transmittable spongiform encephalopathy (TSE). To reduce the risk of spreading of these diseases the European Commission has very recently defined special methods for slaughtering of animals to ensure that all risk material is removed from the feed-chain [115]. This regulation defines how the different types of wastes (animal by-products) not used for human consumption have to be treated. All waste of animal origin is divided in three categories with different demands. The new regulation demands an approval of all plants treating animal waste including quality control by the plant and society. The banning of meat and bone meal as fodder for animals intended for human consumption following the increased number of European cases of bovine spongiform encephalopathy (BSE) has led to investigations of possible and safe disposal methods of the meal. During the discussion of disposal methods, anaerobic digestion followed by utilization of the fertilizer value by spreading the digested sludge on arable land was suggested. The idea was, however, abandoned, and instead, a large part of the Danish meat and bone meal is utilized in cement production and is thereby lost from cycling of nutrients. BSE and the variant of Creutzfeldt-Jakob disease (vCJD) are generally considered to be transmitted by the ingestion of proteinaceous agents (prions), which accumulate in the brain and spinal cord of infected animals and humans. The disease-causing protein (PrPSc) is an abnormal isomer of a host-encoded protein (PrP) that has the ability to change the conformation of normal PrP to PrPSc. The infectious PrPSc is, however, considered to be extremely resistant to enzymatic degradation, heat, and chemical treatment. Proteases are ineffective in inactivating PrPSc, and bioassays have shown that protein remained infectious after autoclaving at temperatures up to 138 °C for 60 min [116]. Among the different chemical inactivation methods tested, alkaline treatments have so far shown most promise, although they are not completely effective. Complete inactivation might, however, be achieved by combination of methods. Based upon one study, in which scrapie-infected hamster brain homogenate remained infectious after 3 years incubation in soil [117], we assume that only a minor reduction of prions will occur during the AD process and that sufficient pretreatment will be necessary to eliminate prions before the anaerobic reactor. 6.2 Control of Chemical Pollutants

Among the chemical pollutants, heavy metals are mainly problematic in wastes of industrial origin and are found in high concentrations in some organic waste and in sewage sludge from wastewater treatment plants with certain industrial influents. Through source reduction and elimination of specific types of wastes, it is generally possible to meet the standards regarding heavy metals for use of residues produced by AD.

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Agricultural waste can contain persistent organic contaminants such as pesticides, antibiotics, and other medicine residues. Industrial wastes, sewage sludge, and household wastes can contain aromatics, aliphatic and halogenated hydrocarbons, organochlorine pesticides, PCBs, PAHs, phthalates, linear alkyl benzenesulfonates (LAS), nonylphenol and nonylphenol ethoxylates. During AD most of the water-soluble organic contaminants are degraded to various degrees. However, hydrophobic compounds such as high molecular phthalates, PAH, and LAS are tightly bound to the particulate phase and are partly unavailable for biological conversion [118]. The potential to remove organic pollutants by pretreatment of sewage sludge by wet oxidation was studied very recently. Unfortunately, these results showed that the conditions suitable for keeping a biogas potential in the waste material resulted in production of high amounts of organic pollutants with a smaller molecular weight than the initial pollutants industries (unpublished). Effectively, a complete decontamination demands incubation at very high temperatures (more than 250 °C) and pressure, which implies that the final gas potential is marginal and that the costs are very high. When comparing full-scale mesophilic and thermophilic AD-reactors operated on the same sewage sludge, it was found that the thermophilic process delivered an effluent with significant lower concentrations of organic pollutants than the effluent from the mesophilic reactor [119]. A higher bioavailability due to a higher solubility of the hydrophobic elements could explain the differences observed. Recent experiments indicated that extreme thermophilic processes improve this reaction further but this needs to be further investigated before any conclusions can be drawn.

7 Conclusions Aaerobic digestion is an important way of handling waste in society. While the emphasis previously was focused on stabilization of sewage sludge, emphasis today is focusing on creation of an effluent, which safely can be used as a fertilizer on farmland. Production of biogas is furthermore gaining more attention, especially for treatment of manure from large-scale animal production. In this picture other types of organic waste such as wastes from food processing or from households will be interesting as a mean of boosting the gas production and, thereby, the economy of the AD plant. Anaerobic digestion can further add value during use of waste or other biomasses for the production of chemicals and energy and this synergy is expected to be further exploited in the future. Anaerobic digestion is a mature technology today. However, as demonstrated in this chapter, there is plenty of room for optimization and improvements. The standardized CSTR reactor has its limitations and implementation of more efficient reactor types such as the immobilized reactor systems has a major potential for treatment of solid waste. Process control on the current AD plant is still relying on in- and output data and no information is available on-line for checking the state of the process and its performance. The microbiology in AD plants is normally regarded as a big black box and very few attempts have been made to control the actual microflora in bioreactors treating waste. Recent research

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has demonstrated that AD plants within close distance of each other can possess different microfloras with different characteristics. Some microbial strains will add superior characteristics to the reactor system and this has major implications for the future of AD plants. Most waste is only partly degraded in the AD plant. Improving the digestibility of waste by using physical or chemical pretreatment methods, which will make the waste more accessible for anaerobic degradation, is another area with major perspectives. One of the most promising areas for the future is the use of extreme thermophilic digestion within the AD plant. The high temperature process will allow for better hydrolysis of the solids, for better sanitation and for better removal of xenobiotics during the treatment process. Acknowledgement. I would like to thank Zuzana Mladenovska, Hinnerk Hartmann, Thomas Ishøy and Peter Westermann for valuable input to this chapter.

8 References 1. Iranpour R, Oh S, Kim H, Shao YJ, Hagekhalil A, Schafer P, Stenstrom MK, Ahring BK (2002) Wat Environ Federation in press 2. Zinder SH (1993) Physiological ecology of methanogens. In: Ferry JG (ed), Methanogenesis. Ecology, physiology, biochemistry & genetics. Chapman & Hall, New York, p 128 3. Schink B (1992) Syntrophism among prokaryotes. In: Balows A, Trüper HG, Dworkin M, Harder W, Schleifer K-H (eds), The prokaryotes. Springer Verlag, Berlin Heidelberg New York p 276 4. Mackie RI, Bryant MP (1981) Appl Environ Microbiol 41:1363 5. Thiele JH, Zeikus JG (1988) Interactions between hydrogen- and formate-producing bacteria and methanogens during anaerobic digestion. In: Erickson LE, Fung DY (eds), Handbook on anaerobic fermentation. Marcel Dekker, New York, p 537 6. Westermann P (1996) World J Microbiol Biotechnol 12:497 7. Schnürer A, Zellner G, Svensson BH (1999) FEMS Microbiol Ecol 29: 249 8. Schnürer A, Houwen FP, Svensson BH (1994) Arch Microbiol 162:70 9. Zinder SH, Koch M (1984) Arch Microbiol 138:263 10. Lee MJ, Zinder SH (1988) Appl Environ Microbiol 54:1457 11. Ahring BK, Ibrahim A, Mladenovska Z (2002) FEMS Microbiology Ecology submitted 12. Ahring BK, Ibrahim AA, Mladenovska Z (2001) Wat Res 35:2446 13. Petersen SP, Ahring BK (1991) FEMS Microbiol Ecol 86:149 14. Ahring BK (1995) Antonie van Leeuwenhoek 67:91 15. Westermann P, Ahring BK, Mah RA (1989) Appl Environ Microbiol 55:514 16. Ahring BK, Schmidt JE, Winther-Nielsen M, Macario AJL, Conway de Macario E (1993) Appl Environ Microbiol 59:2538 17. Lee MJ, Zinder SH (1988) Appl Environ Microbiol 54:124 18. Schnürer A, Schink B, Svensson BH (1996) Int J Syst Bacteriol 46:1145 19. Wiegel J (1990) FEMS Microbiol Rev 75:155 20. Lettinga G (1995) Antonie van Leeuwenhoek 67:3 21. van Lier JB, ten Brummeler E, Lettinga G (1993) J Ferment Bioeng 76:140 22. Kettunen RH, Rintala JA (1997) Appl Microbiol Biotechnol 48:570 23. Kotsyurbenko OR, Nozhevnikova AN, Kalyuzhnyi SV, Zavarzin GA (1993) Microbiology 62:462 24. Zeikus JG, Winfrey MR (1976) Appl Environ Microbiol 31:99 25. Westermann P (1994) FEMS Microbiol Ecol 13:295 26. Varel VH, Hashimoto AG, Chen YR (1980) Appl Environ Microbiol 40:217

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B. K. Ahring Uemura S, Tseng I-C, Harada H (1995) Environ Technol 16:987 Rintala J, Lepistö S, Ahring B (1993) Appl Environ Microbiol 59:1742 Lepistö R, Rintala J (1996) Biores Technol 56:221 van Lier JB (1996) Antonie van Leeuwenhoek 69:1 Zinder SH (1990) FEMS Microbiol Rev 75:125 Wiegel J (1992) The obligatory anaerobic thermophilic bacteria. In: Kristjansson JK (ed), Thermophilic bacteria. Boca Raton, CRC Press, p 105 Ahring BK (1994) Wat Sci Tech 30:241 Ahring BK (1992) Proceedings from International conference Thermophiles. Science and Technology. Reykjavik, Iceland, p 130 Chen M (1983) Appl Environ Microbiol 45:1271 Ahring BK, Mladenovska Z, Iranpour R, Westermann P (2002) Wat Sci Tech 45:293 Wiegant WM (1986) PhD Thesis, Landbouwhogeschool Wageningen, Netherlands Dinsdale RM, Hawkes FR, Hawkes DL (1997) Wat Res 31:163 Mackie RI, Bryant MP (1995) Appl Microbiol Biotechnol 43:346 Sørensen AH, Ahring BK (1993) Appl Microbiol Biotechnol 40:427 Nopharatana A, Clarke WP, Pullammanappallil PC, Silvey P, Chynoweth DP (1998) Biores Technol 64:169 Zheng D, Raskin L (2000) Microb Ecol 39:246 van Lier JB, Grolle KCF, Stams AJM, Conway de Macario E, Lettinga G (1992) Appl Microbiol Biotechnol 37:130 Ahring BK, Sandberg M, Angelidaki I (1995) Appl Microbiol Biotechnol 43:559 Angelidaki I, Ahring BK (1995) Antonie van Leeuwenhoek 68:285 Griffin ME, McMahon KD, Mackie RI, Raskin L (1998) Biotechnol Bioeng 57:342 Pohland FG, Ghosh S (1971) Environ Lett 1:255 Krugel S, Hamel K, Ahring BK (2002) WEF’s 16th Annual Residuals and Biosolids Management Conference. Austin, Texas, USA, March 3–6, 2002 Verstraete W, de Beer D, Pena M, Lettinga G, Lens P (1996) World J Microbiol Biotechnol 12:221 Verstraete W, Vandevivere P (1999) Critical Reviews Environ Sci Technol 28:151 Qureshi MA, Kharbanda VP (1983) J Scient Ind Res 42:597 Day DL, Chen TH, Anderson JC, Steinberg MP (1990) Biomass 21:83 Driessen W, Yspeert P (1999) Wat Sci Tech 40:221 Seghezzo L, Zeeman G, van Lier JB, Hamelers HVM, Lettinga G (1998) Biores Technol 65:175 Jetten MSM, Wagner M, Fuerst J, van Loosdrecht M, Kuenen G, Strous M (2001) Current Opinion in Biotechnology 12:283 Van Loosdrecht MCM, Jetten MSM (1998) Wat Sci Tech 38:1 Ahring BK, Angelidaki I, Johansen K (1992) Wat Sci Tech 25:311 Tafdrup, S (1992) Proceedings from Seventh International Symposium on Anaerobic digestion, 23–27 January, 1994, Cape Town, South Africa, p 460 Mæng H, Lund H, Hvelplund F (1999) Appl Energy 64:195 Dagnall S (1995) Biores Technol 52:275 Hammond G (1993) Biorecovery 2:141 Lusk P (1999) BioCycle 40:52 Rintala JA, Järvinen KT (1996) Waste Management Research 14:163 Rosenwinkel K-H, Meyer H (1999) Wat Sci Tech 40:101 Hartmann H, Møller HB, Ahring BK (2002) Proceedings from VII Latin American Workshop and Symposium on Anaerobic Digestion. Yucatán, México, October 2002 Clausen A (2001) PhD Thesis, Technical University of Denmark, Lyngby, Denmark Hartmann H, Angelidaki I, Ahring BK (2001) Proceedings from 9th World Congress of Anaerobic Digestion, Antwerpen – Belgium, Sept 2–6, 2001, p 301 Scherer PA, Vollmer G-R, Fakhouri T, Martensen S (2000) Wat Sci Tech 41:83 Angelidaki I, Ahring BK (2000) Wat Sci Tech 41:189 Hartmann H, Angelidaki I, Ahring BK (2000) Wat Sci Tech 41:145

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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. 98. 99. 100. 101. 102. 103. 104. 105. 106. 107. 108. 109. 110. 111. 112. 113. 114.

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Kepp U, Machenbach I, Weisz N, Solheim OE (2000) Wat Sci Tech 42:89 Ahring BK, Jensen K, Nielsen P, Bjerre AB, Schmidt AS (1996) Biores Technol 58:107 Torry-Smith M, Sommer P, Ahring BK (2003) Biotechnol – Bio Wat Res submitted Greene H (1995) Anaerobic digester process enhancement by pulse power treatment. Scientific Utilization, Inc., Decatur, AL Møller HB, Sommer SG, Ahring BK (2002) Biores Technol submitted Kalyuzhnyi S, Fedorovich V, Nozhevnikova A (1998) Biores Technol 65:221 Kalyuzhnyi S, Sklyar V, Fedorovich V, Kovalev A, Nozhevnikova A, Klapwijk A (1999) Wat Sci Tech 40:223 Worley JW, Das KC (2000) Appl Eng Agri 16:555 Aoki N, Kawase M (1991) Wat Sci Tech 23:1147 Kübler H, Schertler C (1994) Wat Sci Tech 30:367 Angelidaki I, Ahring BK (1992) Appl Microbiol Biotechnol 37:808 Koster IW, Cramer A (1987) Appl Environ Microbiol 53:403 Alatriste-Mondragon F, Iranpour R, Ahring BK (2002) Wat Res in press Sandberg M, Ahring BK (1992) Appl Microbiol Biotechnol 36:800 Omil F, Mendez R, Lema JM (1996) Water SA 22:173 Athanassopoulos N, Kouinis J, Papadimitriou A, Koutinas AA (1989) Biological Wastes 30:53 Daoming S, Forster CF (1994) Environ Technol 15:287 Koster IW, Koomen E (1988) Appl Microbiol Biotechnol 28:500 Hansen KH, Angelidaki I, Ahring BK (1998) Wat Res 32:5 Speece RE. (1996) Anaerobic biotechnology for industrial wastewaters. Archae Press, Nashville, Tennessee, USA Hansen KH, Angelidaki I, Ahring BK (1999) Wat Res 33:1805 Boopathy R (1998) Biores Technol 64:1 Pind PF, Angelidaki I, Ahring BK (2002) Biotechnol Bioengineering submitted Mladenovska Z, Ishøy T, Mandiralioglu A, Westermann P, Ahring BK (2001) Proceedings from International conference Anaerobic Digestion 2001, 2–6 September 2001, Antwerpen, Belgium, p 183–188 Mladenovska Z, Ahring BK (2000) FEMS Microbiol Ecol 31:225 Bendixen HJ (1999) IEA Bioenergy Workshop. Hygienic and environmental aspects of anaerobic digestion: Legislation and experience in Europe, Stuttgart 29–31 March 1999, p 27 Mawdsley JL, Bardgett RD, Merry RJ, Pain BF, Theodorou MK (1995) Appl Soil Ecol 2:1 Pell AN (1997) J Dairy Sci 80:2673 Turner C, Burton CH (1997) Biores Technol 61:9 Aitken MD, Mullennix RW (1992) Wat Environ Res 64:915 Ponugoti PR, Dahab MF, Surampalli R (1997) Wat Environ Res 69:1195 Lund B, Jensen VF, Have P, Ahring B (1996) Antonie van Leeuwenhoek 69:25 Kunte DP, Yeole TY, Chiplonkar SA, Ranade DR (1998) J Appl Microbiol 84:138 Arridge H, Oragui JI, Pearson HW, Mara DD, Silva SA (1995) Wat Sci Tech 31:249 Carrington EG, Pike EB, Auty D, Morris R (1991) Wat Sci Tech 24:377 Watanabe H, Kitamura T, Ochi S, Ozaki M (1997) Wat Sci Tech 36:25 Nielsen B, Petersen G (2000) Wat Sci Tech 42:65 Duarte EA, Mendes B, Oliveira JS (1992) Wat Sci Tech 26:2169 Cheunbarn TP, Krishna R (2000) J Environ Engineering 126:796 Krugel S, Nemeth L, Peddie C (1998) Wat Sci Tech 38:409 Huyard A, Ferran B, Audic J-M (2000) Wat Sci Tech 42:41 Iranpour R, Shao YJ, Stenstrom M, Ahring BK (2002) Wat Environ Res in press Bendixen HJ (1995) Wat Sci Tech 30:171 Ahring BK, Lund B, Jungersen G, Have P, Frøkjær Jensen V (1995) Modelstudier vedrørende overlevelse af virus i gyllebaseret biomasse under udrådning i laboratorieskala biogasanlæg. Smitstofreduktion i biomasse. Danish Veterinary Service, Frederiksberg. Vol II: Rep. no. 10

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115. Regulation of the European Parliament and of the Council laying down health rules concerning animal by-products not intended for human consumption, 2001 116. Taylor DM (1998) Journal of Food Safety 18:265 117. Brown P, Gajdusek DC (1991) Lancet 337:269 118. Ejlertsson J, Alnervik M, Jonsson S, Svensson BH (1997) Environ Sci Tech 31:2761 119. Alatriste-Mondragon F, Ahring BK (2002) Wat Res submitted

Received: March 2002

CHAPTER 6

Metabolic Interactions Between Methanogenic Consortia and Anaerobic Respiring Bacteria A.J.M. Stams 1 · S.J.W.H. Oude Elferink 2 · P. Westermann 3 1

2 3

Wageningen University and Research Centre, Laboratory of Microbiology, Hesselink van Suchtelenweg 4, 6703 CT Wageningen, The Netherlands. E-mail: [email protected] ID TNO Animal Nutrition, Edelhertweg 15, P.O. Box 65, 8200 AB, The Netherlands. E-mail: [email protected] Department of Environmental Microbiology and Biotechnology, The Technical University of Denmark, Building 227, 2800 Lyngby, Denmark. E-mail: [email protected]

Most types of anaerobic respiration are able to outcompete methanogenic consortia for common substrates if the respective electron acceptors are present in sufficient amounts. Furthermore, several products or intermediate compounds formed by anaerobic respiring bacteria are toxic to methanogenic consortia. Despite the potentially adverse effects, only few inorganic electron acceptors potentially utilizable for anaerobic respiration have been investigated with respect to negative interactions in anaerobic digesters. In this chapter we review competitive and inhibitory interactions between anaerobic respiring populations and methanogenic consortia in bioreactors. Due to the few studies in anaerobic digesters, many of our discussions are based upon studies of defined cultures or natural ecosystems. Keywords. Competition, Sulfate reduction, Denitrification, Acetogenesis, Inhibition

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Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

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Metabolic Interactions in Methanogenic Bioreactors

2.1 2.1.1 2.1.2 2.2

Competitive Interactions . . . Kinetic Competition . . . . . Thermodynamic Competition Inhibitory Interactions . . . .

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Competition in the Presence of Oxygen . . . . . . . . . . . . . . Competition Between Nitrogen Reducers and Methanogenic Consortia . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.3 Competition Between Manganese and Iron Reducers and Methanogenic Consortia . . . . . . . . . . . . . . . . . . . . . . 3.4 Competition Between Sulfate-Reducing and Acetogenic Bacteria and Methanogenic Consortia . . . . . . . . . . . . . . . . . . . 3.4.1 Competition for Hydrogen . . . . . . . . . . . . . . . . . . . . . 3.4.2 Competition for Acetate . . . . . . . . . . . . . . . . . . . . . . 3.4.3 Competition for Methanol . . . . . . . . . . . . . . . . . . . . .

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3.4.4 Competition for Organic Acids and Ethanol . . . . . . . . . . . . . 46 3.4.5 Competition for Sulfate . . . . . . . . . . . . . . . . . . . . . . . . 48 3.5 Competition Between Sulfate-Reducers and Acetogens in the Absence of Sulfate . . . . . . . . . . . . . . . . . . . . . . . . 49 4

Inhibition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 50

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Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

6

References

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1 Introduction Very few environments exist in which only one population of microorganisms thrives or where populations of microorganisms do not affect each other either positively or negatively. As discussed in Chap. 1, anaerobic ecosystems such as methanogenic bioreactors are characteristic by their complex food-chains and the close symbiotic relationship between the different links in the chain, and are often exemplified as classical symbiotic ecosystems in which organisms consume the products of the preceding link in the chain, rather than consuming each other. The symbiosis between hydrogen-producing and hydrogenconsuming microorganisms is confined to a narrow range of hydrogen partial pressures outside which the reactions become thermodynamically unfavorable for one or the other part of the relationship. This can be caused by overloading with easily degradable compounds or by unintentional influence of inhibitory compounds. Compounds inhibiting methane production in a digester might exert their action either by direct inhibition of microbes in the anaerobic degradation chain or by stimulating microorganisms present in the digester to compete with methanogens or preceding links leading to reduced methane production and other unfavorable effects such as corrosion [1]. In this chapter we will discuss various types of direct and indirect competitive interactions between methanogenic consortia and anaerobic respiring bacteria in anaerobic bioreactors.

2 Metabolic Interactions in Methanogenic Bioreactors 2.1 Competitive Interactions

Competition between two or more populations of microorganisms is a negative relationship in which the different populations often are adversely affected with respect to their survival and growth. Also competition is considered the most important interaction among organisms, and is one of the major responsible causes of the selection pressure leading to the evolution of species.

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Fig. 1. Model of kinetic and thermodynamic competition among sulfate-reducing bacteria and methanogenic Archaea

The competitive interactions among anaerobic microorganisms can be roughly divided into kinetic competition and thermodynamic competition (Fig. 1). Kinetic competition refers to the determination of competitive capabilities by kinetic measurements of microbial growth, although the underlying mechanism for the observed effects might be thermodynamic. Thermodynamic competition means that one organism is capable of growing at and maintaining a substrate concentration below the minimum concentration for uptake (threshold concentration) of other organisms due to a higher energy yield in the conversion of the compound. In anaerobic fermentation of organic compounds, numerous pathways and combinations of pathways are used leading to different energy yields. However, since anaerobic fermentation is internally optimized in the cells to gain a maximum energy yield and an optimal redox balance [2, 3] the energetic outcome is often the same. This has the consequence that fermentative competitive interactions are mainly of kinetic character. Most of the studies which have examined competition between anaerobic fermenting bacteria have focused on gastrointestinal systems [4] and very little is known on this type of competitive interaction in anaerobic digestion processes. Apart from interactions between fermenting sulfate-reducing bacteria and acetogenic bacteria, we will not discuss this topic in this chapter.

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Table 1. The respiration hierarchy Acceptor

Product

E0¢ (V)

Oxygen O2 Manganic ion Mn4+ Ferric ion Fe3+ Nitrate NO3– Selenate SeO42– Arsenate AsO43– Sulfate SO42– Carbon dioxide CO2 Carbon dioxide CO2

Water H2O Manganous ion Mn2+ Ferrous ion Fe2+ Nitrogen N2 Selenite SeO32– Arsenite AsO33– Sulfide HS – Methane CH4 Acetate CH3COO –

+0.82 +0,80 +0.77 +0.76 +0.48 +0.14 –0.22 –0.24 –0.29

In contrast to aerobic conditions where most heterotrophic microorganisms utilize oxygen as a terminal electron acceptor and in most cases follow the same metabolic pathway ending in complete mineralization of the organic compounds into CO2 and H2O, the biochemical diversity of anaerobic microbial communities is huge. A large number of electron acceptors can be used by different anaerobic organisms in anaerobic respiration processes (Table 1). The most important inorganic electron acceptors are Mn4+, Fe3+, NO3– , SO42– and CO2 . The respiration processes where these acceptors are used are normally separated either in space or time. This is due to a different energy outcome of the processes according to the Gibbs equation: DG0¢ = –n · F · DE0¢ in which DG0¢ is the Gibbs free energy at pH = 7; n is the number of electrons transferred in the oxidation-reduction reaction; F is Faraday’s constant (96.490 kJ/V) and DE0¢ is the redox potential (E0¢) of the electron-accepting reaction minus the redox potential of the electron-donating reaction. From this equation it is obvious that the larger the difference is between the redox potentials of the half-reactions, the larger is the amount of energy available to the organism performing the reaction. The consequence is a hierarchy, which often resembles the order seen in Table 1. In most environments, some of the respiration processes do not occur, or only occur to a minor extent, due to the lack or exhaustion of available electron acceptors. The energy available to a respiring organism is not only dependent upon the difference in redox potential between electron donor and acceptor. Also concentrations of the reactants and temperatures deviating from standard conditions affect the energy outcome according to the Nernst equation DG = DG0 + RT · ln [B]/[A] in which DG0 is the change in Gibbs free energy under standard conditions, R is the gas constant, T is temperature and [B] and [A] are the concentrations of the two components of the reaction A ¤ B. According to the respiration hierarchy, sulfate reduction excludes methanogenic utilization of common substrates, which is verified in high-sulfate environments such as marine sediments [5]. However in, e.g., freshwater sediments, the two processes can coexist or even be dominated by methanogenesis due to equilibrium displacements caused by low sulfate concentrations making sulfate reduction thermodynamically less favorable than methane production [6].

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2.1.1 Kinetic Competition

This is the classical competitive interaction, the theory of which has been established in studies of defined cultures in chemostats [7, 8]. According to kinetically-based competition models, the outcome of interactions between two microorganisms competing for the same growth-limiting substrate can be predicted from the relationship between substrate concentration and the specific growth rate (µ) according to the Monod equation: µ = µmax ¥ S/Ks + S. Two typical relationships can be observed in studies of competitive interactions (Fig. 2 a, b). In Fig. 2a, organism I will grow faster than organism II at any substrate concentration, while the outcome in Fig. 2b is dependent upon the substrate concentration. The pattern seen in Fig. 2a is typical of organisms utilizing different electron acceptors with different energy yields for the oxidation of a common substrate, since the energy yield is higher for the electron acceptor with the highest redox potential at all electron donor concentrations. The pattern seen in Fig. 2b is typical for organisms utilizing the same metabolism but having different ecological strategies. In natural ecosystems, such as sediments, the concentration of nutrients needed to support growth is often very low. Among the organisms using the same type of metabolism under these conditions, type II in Fig. 2b having a high substrate affinity (low Ks) and a relatively low maximal growth rate (µmax) will normally dominate. This group is assigned to “K selec-

Growth rate

a

b

Substrate concentration Fig. 2. Growth rate as a function of substrate concentration in two different scenarios (a and b). a represents two organisms with different energy metabolism, I having the highest energy yield. b represents two organisms with the same energy metabolism, but with different ecological strategies. I is assigned to “r” selection while II is assigned to “K” selection

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tion” which refers to organisms that can most effectively utilize the resources available [9]. In gastrointestinal environments and anaerobic bioreactors, opportunistic types of organisms (type I) will normally dominate, since type II has a longer doubling time than the retention time of the system. This group is assigned to “r selection” referring to a high potential r value (rate of population growth/individual) [9]. 2.1.2 Thermodynamic Competition

In natural environments, the substrate concentration for most organisms is normally well below Ks . For all organisms, there is a specific minimum concentration of substrate necessary to gain conservable energy. This minimal “quantum” of energy, which can be conserved, corresponds to the energy needed for translocation of 1 proton. The phosphorylation of ATP to ADP has a DG¢ of +49 kJ/mol corresponding to 60–70 kJ/mol when compensating for energy conservation efficiency [10]. Since 3 protons are needed in the phosphorylation of ADP to ATP, we can assume that the smallest amount of energy which can be conserved is 1/3 of the phosphorylation energy, corresponding to a minimum DG¢ of –20 kJ/mol. Inserting this value and DG¢0 for different respiration processes in the Nernst equation, the substrate concentration yielding the minimum amount of energy (the threshold concentration) can be calculated for each process under the prevailing conditions of the specific ecosystem. Several authors have shown that organisms utilizing electron acceptors with higher redox potentials can maintain electron donor concentrations below the threshold for uptake of organisms utilizing electron acceptors with lower redox potentials [11–13]. Other studies have shown that significant differences in threshold values for common substrates also can be found among species utilizing the same type of metabolism [14]. 2.2 Inhibitory Interactions

Several compounds, which serve as electron donors to respiring bacteria, might inhibit members of the methanogenic consortia. Also some products from anaerobic respiration might affect the activity of these consortia. The modes of action can be indirect by increasing the redox potential to levels that interfere with the biochemistry of the anaerobic microorganisms, or direct by chemical reaction with proteins or other cell constituents. It has been assumed that many anaerobic microorganisms have specific demands for low redox potentials in their environment to make their energy metabolism thermodynamically possible [15]. This conception has since been moderated and several reports have shown that the parameters controlling growth of most anaerobes is the oxygen concentration and only to a lesser degree the redox potential of the environment. This has been demonstrated in studies of fermentative rumen bacteria, but also in studies of microorganisms considered extremely sensitive to aerobic conditions [16]. Fetzer and Conrad

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37

[17] have, for instance, demonstrated that methane production in axenic cultures of Methanosarcina barkeri proceeded at normal rates in oxygen-free media where the redox potential was elevated to +420 mV. The direct inhibition of methanogenic consortia by electron acceptors is mediated by several mechanisms. Oxygen is toxic to all obligatory anaerobic microorganisms. Many anaerobes are rich in flavine enzymes, and may also contain quinones and iron-sulfur proteins, which can react spontaneously with oxygen to yield hydrogen peroxide, superoxide and hydroxyl radicals. Since most anaerobes lack peroxidase, catalase and superoxide dismutases, which destroy the reactive oxygen species, damage of essential cell components can occur upon oxygen exposure. Superoxide dismutase has, however, been demonstrated in some anaerobic microorganisms. Kirby et al. [18] have, for instance, characterized a superoxide dismutase from the obligatory anaerobe Methanobacterium bryantii. Other electron acceptors, such as oxidized nitrogen and sulfur species, have also been shown inhibitory to anaerobic microorganisms. Although the metabolism of these electron acceptors is competitive to anaerobes utilizing electron acceptors with a more negative redox potential, the reduction of the inhibitory compounds might lead to the production of less inhibitory compounds and, hence, relieve the inhibition. In some cases, however, the products of anaerobic respiration are more toxic than the parent compounds. This will be discussed in details in the next chapters.

3 Competition 3.1 Competition in the Presence of Oxygen

Although oxygen is the naturally occurring electron acceptor yielding the highest amount of energy leading to effective outcompetition of anaerobic microorganisms, oxygen respiration and anaerobic metabolism are mutually exclusive processes mainly due to the toxicity of oxygen, which can be observed in all aerobic environments. Most facultatively aerobic microorganisms capable of anaerobic respiration suppress these processes in favor of oxygen respiration when oxygen is present. Only environments in which rapid changes between oxic and anoxic conditions occur, such as alternating sludge treatment basins, favor constitutively anaerobic respiring bacteria [19]. In true oxic environments, anaerobic processes are normally only occurring in organic-rich micro- and macro-niches, where oxygen is depleted at a higher rate than it diffuses into the niche. Oxygen is normally excluded in anaerobic digestion processes, and only small amounts might enter the reactors together with, e.g., strongly aerated substrates [20]. Due to the low solubility of oxygen, this does normally not pose a problem to the anaerobic microorganisms in the digester and is rapidly scavenged by facultative bacteria. Kato et al. [21] demonstrated a high oxygen tolerance of methanogens in granular sludge due to mainly oxygen consumption by facultatively anaerobic bacteria metabolizing easily degradable substrates.

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3.2 Competition Between Nitrogen Reducers and Methanogenic Consortia

From an immediate evaluation of redox potentials of methanogenesis and nitrate reduction, it is obvious that nitrate reducers should outcompete methanogens due to the much higher energy yield of nitrate respiration. This has been verified in a few natural environments [22]. Under most circumstances, however, the effects of nitrogen oxides to anaerobic digestion are ambiguous and very complex, and to our knowledge no certain verification of competition in anaerobic digesters in which inhibition has been excluded has been published so far. Denitrification and methanogenesis are performed by microbial populations each requiring their distinct environmental conditions. Most true denitrifiers are facultatively anaerobic bacteria utilizing either oxygen respiration or denitrification as sole energy source. If none of these metabolisms are possible due to the lack of appropriate electron acceptors, the bacteria will probably not thrive in the digester. Instead, fermentative bacteria reduce oxidized nitrogen species for dissimilatory electron dissipation. The product is either nitrite or ammonia, and only the reduction of nitrate to nitrite is energy yielding. The further reduction to ammonia is considered non-energy yielding and hence without competitive value. Several authors have shown that high carbon to nitrogen ratios which are normally found in anaerobic digesters favor dissimilatory nitrate reduction to ammonia [23], while others [24] found that a high COD/NO3– did not favor dissimilatory reduction of nitrate to ammonia. The nature of the carbon source has also been shown to influence whether nitrate is reduced to ammonia or dinitrogen [25]. When glucose or glycerol was added as carbon source, 50% of the nitrate was reduced to ammonium, while 100% was denitrified completely in the presence of acetate or lactate. Several authors have demonstrated that denitrification and methanogenesis can proceed in the same reactor as long as the two processes are spatially separated. Hendriksen and Ahring [26] found that denitrification took place in the bottom of an upflow anaerobic sludge blanket reactor utilizing all available nitrate. Methanogenensis occurred in the uppermost part of the reactor, which was depleted from nitrogen oxides. In a mixed culture system of denitrifying and methanogenic sludge in a digester enriched with methanol, Chen and Lin [27] observed no competitive interactions between the two communities. Methanogenesis was, however, inhibited as long as nitrate or nitrite was present in the reactor. Percheron et al. [24] studied methanogenesis and nitrate reduction in an anaerobic digester fed with sulfate-rich wastewater. Sulfate reduction was inhibited by the presence of nitrate while methanogenesis proceeded until the onset of denitrification and production of nitrite after which it also was inhibited. Sulfide served as electron donor for some of the denitrifying bacteria. When sulfide was precipitated by ferrous iron, only dissimilatory nitrate reduction occurred with no nitrite production. This led to a stimulation of methanogenesis compared to a control digester, probably due to extensive acetate production by the dissimilatory nitrate reducers. No specific competitive interactions between methanogens and denitrifiers were verified in this study either. Clarens et al. [28] studied the effects of nitrogen oxides and denitrification on a

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39

pure culture of an aceticlastic methanogen (Methanosarcina mazei) and mixed cultures of M. mazei and a denitrifying bacterium (Pseudomonas stutzeri). It was demonstrated that the observed cessation of methanogenesis upon nitrate addition was a consequence of inhibition by denitrification products (NO2– , N2O) rather than competition by the denitrifying bacterium. The authors found that 50 mM NO3– inhibited methanogenesis from acetate by 65% while 0.18 mM NO2– and 0.32 mM N2O were almost completely inhibitory. In a study of the effects of nitrogen oxides on methanogenesis and other metabolic activities in an anoxic rice-field soil, Klüber and Conrad [22] tried to resolve inhibition and competition among nitrogen-respiring bacteria and methanogens. The addition of nitrate, nitrite, nitrous oxide and nitric oxide all resulted in an immediate arrest of methanogenesis until the nitrogen oxides were consumed. Methanogenesis then resumed at a similar or lower rate. None of the nitrogen oxides affected acetate concentrations negatively while nitrate, nitrite and nitrous oxide additions temporarily reduced hydrogen partial pressures to low exergonic or even endergonic values for methanogenesis. Since more than 70% of the methane produced was derived from acetate, Klüber and Conrad’s results indicate that toxicity rather than competition is responsible for the inhibition observed. When nitrate, nitrite or nitrous oxide were added, sulfate or/and ferric iron concentrations increased, probably as respiration products of sulfide and ferrous iron oxidation coupled to denitrification. The maintenance of low hydrogen partial pressures might, therefore, be due to the activity of iron or sulfate reducing bacteria rather than denitrifying bacteria. The mechanisms of nitrogen oxide inhibition of methanogenesis in anaerobic digesters can be considered far from solved, and is probably a complex mechanism composed of toxicity, competition, and indirect stimulation of other respiring bacteria by oxidation of reduced electron acceptors such as ferrous iron and sulfide. 3.3 Competition Between Manganese and Iron Reducers and Methanogenic Consortia

Both manganic ions [Mn(IV)] and ferric ions [Fe(III)] can act as potent electron acceptors in anaerobic respiratory processes carried out by a variety of microorganisms coupled to the oxidation of organic and inorganic compounds. Mn(IV) and Fe(III) can also be reduced in non-enzymatic chemical reactions under anaerobic conditions, and much of the effort in earlier studies of respiration of the two compounds was devoted to the separation of non-biological from biological reductions [29]. The isolation of numerous bacteria capable of Fe(III) and Mn(IV) reduction and properly designed experiments with environmental samples have, however, unambiguously verified this type of bacterial respiration. Several authors have shown that methanogenesis and other terminal anaerobic processes can be outcompeted by ferric- and manganic-reducing bacteria due to their maintenance of acetate concentrations and H2 partial pressures below the threshold of methanogenic Archaea and sulfate-reducing bacteria [12]. Although respiration with Mn(IV) or Fe(III) is thermodynamically more favorable than sulfate reduction or methanogenesis, several authors have shown

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that Mn(IV) and Fe(III) respiration is less efficient with crystalline than with amorphous forms of the two electron acceptors [30, 31]. Lovley and Phillips [12] showed that sulfate reduction and methanogenesis are only inhibited by Fe(III)reducing bacteria when Fe(III) is in an amorphous form. In a study of anaerobic respiration processes in flooded soils, Peters and Conrad [32] found that Mn(IV), Fe(III), and sulfate reduction proceeded simultaneously, possibly due to the crystalline structures of the Mn and Fe minerals in the soils. Since oxidation of short-chain fatty acids and H2 are the main electrondonating processes of both Fe(III) and Mn(IV) reduction, one could expect that these two electron acceptors could play a significant role in anaerobic digestion when present in high concentrations. Besides direct competitive interactions, Mn(IV) has been shown to act as an electron acceptor in the oxidation of elemental sulfur (S0) to sulfate catalyzed by sulfate-reducing bacteria [33]. This could lead to the stimulation of sulfate reduction upon exhaustion of Mn(IV). Very few investigations have, however, been carried out regarding the effects of Fe(III) and Mn(IV) on anaerobic digestion. One major reason for this could be the very low contents of iron and manganese normally found in wastewater. In average wastewater with a BOD of 290 g O2/m3, the typical iron and manganese concentrations have been estimated to 3.5 mg/g BOD and 0.35 mg/g BOD, respectively [34]. In a study of the effect of ferric chloride addition to anaerobic sludge digesters to precipitate struvite (MgNH4PO4 · 6 H2O), Mamais et al. [35] added FeCl3 at doses ranging from 0 to 20.5 mM Fe/L. A slight increase in gas production was observed upon FeCl3 addition, but no other effects were found. Further investigations are needed with respect to these two electron acceptors to clarify their actual and potential effects on different anaerobic digestion processes. 3.4 Competition Between Sulfate-Reducing and Acetogenic Bacteria and Methanogenic Consortia

In environments where sulfate is present, sulfate-reducing bacteria will compete with methanogenic consortia for common substrates. Direct competition will occur for substrates like hydrogen, acetate and methanol. Compared with methanogens, sulfate-reducing bacteria are much more versatile than methanogens. Compounds like propionate and butyrate, which require syntrophic consortia in methanogenic environments, are degraded directly by single species of sulfatereducing bacteria. The physiology of sulfate-reducing bacteria has been reviewed before by Widdel [36],Widdel and Hansen [37] and Colleran et al. [38], while the physiology of methanogenic consortia was reviewed by Stams [3], Schink [39] and Verstraete et al. [40]. Some key reactions in anaerobic environments are listed in Table 2. Kinetic properties of sulfate-reducers, methanogens, and acetogens can be used to predict the outcome of the competition for these common substrates [6, 41–44]. For bacteria growing in suspension, Monod kinetic parameters such as the half-saturation constant (Ks) and the specific growth rate (µmax) can be used. When bacterial growth is negligible, as is often the case in reactors with a dense

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Table 2. Acetogenic and methanogenic reactions, and sulfate-reducing reactions involved in the degradation of organic matter in methanogenic bioreactors, and sulfate-reducing bioreactors, respectively DG0¢ a [kJ/reaction]

Reaction Syntrophic Acetogenic reactions Æ Propionate– + 3 H2O Butyrate– + 2 H2O Æ Lactate– + 2 H2O Æ Ethanol + H2O Æ Methanol + 2 H2O Æ

Acetate– + HCO3– + H+ + 3 H2 2 Acetate– + H+ + 2 H2 Acetate– + HCO3– + H+ + 2 H2 Acetate– + H+ + 2 H2 HCO3– + H+ + 3 H2

+76.1 +48.3 –4.2 +9.6 +23.5

Methanogenic reactions 4 H2 + HCO3– + H+ Acetate– + H2O Methanol

Æ CH4 + 3 H2O Æ CH4 + HCO3– Æ 3/4 CH4 + 1/4 HCO3– + 1/4 H+ + 1/4 H2O

Sulfate-reducing reactions 4 H2 + SO42– + H+ Acetate– + SO42– Propionate– + 3/4 SO42– Butyrate– + 1/2 SO42– Lactate– + 1/2 SO42– Ethanol + 1/2 SO42– Methanol + 3/4 SO42– + 1/4 H+

Æ Æ Æ Æ Æ Æ Æ

HS– + 4 H2O 2 HCO3– + HS– Acetate– + HCO3– + 3/4 HS– + 1/4 H+ 2 Acetate– + 1/2 HS– + 1/2 H+ Acetate– + HCO3– + 1/2 HS– + 1/2 H+ Acetate– + 1/2 HS– + 1/2 H+ + H2O HCO3– + 3/4 HS–

–151.9 –47.6 –37.7 –27.8 –80.0 –66.4 –90.4

Homoacetogenic reactions Lactate– Ethanol + HCO3– Methanol + 1/2 HCO3– 4 H2 + 2 HCO3– + H+

Æ Æ Æ Æ

11/2 Acetate– + 1/2 H+ 11/2 Acetate– + H2O + 1/2 H+ 3/ Acetate– + H O 4 2 Acetate– + 4 H2O

–56.6 –42.6 –55.0 –104.6

a

–135.6 –31.0 –78.2

DG0¢-values are taken from Thauer et al. (1977) [2].

biomass concentration, Michaelis-Menten kinetics may be used to predict which type of organism has the most appropriate enzyme systems to degrade substrates. Therefore, both the Vmax/Km and the µmax/Ks ratio gives an indication of the outcome of competition at low substrate concentrations [42]. 3.4.1 Competition for Hydrogen

In anaerobic environments methanogens, homoacetogens and sulfate-reducers will compete for hydrogen. Thermodynamically, homoacetogenesis is less favorable than methanogenesis and sulfate reduction. Homoacetogens are very poor hydrogen-utilizing organisms [13]. When grown on organic substrates like ethanol and lactate in the presence of hydrogenotrophic methanogens, they even produce hydrogen. In the absence of methanogens 1.5 acetate is produced per lactate or ethanol that is degraded. However, in the presence of methanogens only 1 acetate per lactate or ethanol is produced, while reducing equivalents are disposed of as hydrogen.

42

A.J.M. Stams et al.

Table 3. Selected growth kinetic data of hydrogenotrophic sulfate-reducing bacteria and methanogens. For references see Oude Elferink [81] and Oude Elferink et al. [149] Microorganism Sulfate reducers Desulfovibrio desulfuricans b vulgaris b Desulfovibrio G11 Desulfobacter hydrogenophilus Desulfobacterium autotrophicum Desulfobulbus propionicus b Desufomicrobium escambium Methanogens Methanobacterium bryantii formicicum b ivanovii Methanobrevibacter arboriphilus b smithii Methanococcus vannielii Methanospirillum hungatei strain BD Methanosarcina barkeri b mazei a b

Ks (µM)

µmax (1/day)

1.6–4.3 0.7–5.5 2.4–4.2 1.2–1.6 1.0 0.7–1.1 0.2–1.7 1.4

0.3–1.9 1.2–3.1 0.8–1.7 0.7–3.4 4.1 4.1 5.8–7.3 1.2–1.8 2.4–2.8 1.4–1.8 1.4–1.7

Yield a Km (g/mol H2) (µM)

1.9 0.6–3.1 1.4–2.0

0.6 0.9 1.1

1.8–4.0 1.3–4.0 1.1

88 30 65

2 14

0.6–1.3

6.6

0.3–0.5

5.0

1.6–2.2

Vmax (µmol/min · g)

13

70 110

The yield is given in gram cell dry weight per mol. Several strains.

Studies with sediments and sludge from bioreactors have indicated that at an excess of sulfate hydrogen is mainly consumed by sulfate reducers [6, 45–49]. In reactors with immobilized biomass the activity of hydrogenotrophic methanogens is completely suppressed within a few weeks when sulfate is added [50]. As hydrogenotrophic methanogens are still present in high numbers in such reactors, this effect cannot simply be explained by Michaelis-Menten or Monod kinetic data (Table 3). In methanogenic environments the hydrogen partial pressure is low. However, by addition of sulfate the hydrogen partial pressure may even become lower. The hydrogen partial pressure becomes so low that thermodynamically hydrogenotrophic methanogenesis is not possible any more (Fig. 1). In freshwater sediments a threshold hydrogen concentration of 1.1 Pa has been measured; this value was lowered to 0.2 Pa by the addition of sulfate [6]. An additional effect of the addition of sulfate is that hydrogen formation becomes less important. In the absence of sulfate, hydrogen has to be formed by acetogenic bacteria in the oxidation of compounds like lactate, alcohols, propionate and butyrate. However, in the presence of sulfate, all these compounds can be oxidized directly by sulfate-reducers without the intermediate formation

Metabolic Interactions Between Methanogenic Consortia and Anaerobic Respiring Bacteria

43

of hydrogen. However, this explanation cannot be the only one because fermentative glucose- and amino acid-degrading bacteria will always form some hydrogen. Methanogens, which grow on H2/CO2 , are autotrophic [51].Among the hydrogen-utilizing sulfate-reducing bacteria both autotrophic and heterotrophic species have been isolated [37]. The classical Desulfovibrio species require acetate and carbon dioxide or another organic carbon source for growth whereas, e.g., Desulfobacterium sp. can use CO2 as the sole source of carbon [37, 52, 53]. An interesting observation has been made by Brysch et al. [54]. Enrichments in media with H2 and sulfate as energy substrates and carbon dioxide as the sole carbon substrate resulted in stable cultures of Desulfovibrio and Acetobacterium, in a cell ratio of about 20 to 1. The Desulfovibrio species required acetate for growth, which was provided by the homoacetogenic Acetobacterium species. Sulfate-reducing bacteria have a higher affinity for hydrogen than homoacetogens, but apparently the sulfate-reducers are dependent on the homoacetogens for synthesis of their carbon source acetate. It can be speculated that under these conditions the kinetic properties of homoacetogens determine the kinetic properties of the sulfate-reducers. In that case, methanogens would win the competition for hydrogen from the sulfate-reducers even at an excess of sulfate. Unfortunately, an experiment which could demonstrate this has never been performed. Van Houten et al. [55, 56] started up bioreactors at high hydrogen partial pressures with solely bicarbonate as carbon source. This led to the coexistence of sulfate-reducers and homoacetogens. 3.4.2 Competition for Acetate

It has been shown that in marine and freshwater sediments acetate is mainly consumed by sulfate-reducers when sufficient sulfate is present [45, 46, 49, 57]. However, for anaerobic digesters it is less clear how acetate is degraded. A complete conversion of acetate by methanogens, even at an excess of sulfate, has been reported [46–48, 50, 58–61]. However, in some studies a predominance of acetate-degrading sulfate-reducers was found [62–64]. Some factors which may affect the competition between sulfate-reducers and methanogens are discussed below. The work of Schönheit et al. [43] has indicated that the predominance of Desulfobacter postgatei in marine sediments could be explained by its higher affinity for acetate than Methanosarcina barkeri. The Km values were 0.2 and 3.0 mM, respectively (Table 4). However, in bioreactors Methanosarcina sp. are only present in high numbers when the reactors are operated at a high acetate concentration or operated at a low pH [65]. Generally, Methanosaeta (former Methanothrix, [66]) sp. are the most important aceticlastic methanogens in anaerobic bioreactors [65, 67–69]. Also in freshwater sediments Methanosaeta seems to be the most numerous acetoclastic methanogen [70]. Methanosaeta sp. have a higher affinity for acetate than Methanosarcina sp.; their Ks is about 0.4 mM [71]. In addition, D. postgatei and other Desulfobacter species are typical marine bacteria, which have not yet been isolated in freshwater media [72].

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Table 4. Selected growth kinetic data of acetotrophic sulfate-reducing bacteria and methanogenic bacteria. For references see Oude Elferink [81] and Oude Elferink et al. [149] Microorganism

Ks µmax (µM) (1/day)

Sulfate reducers Desulfobacter curvatus hydrogenophilus latus postgatei b Desulfotomaculum acetoxidans Desulforhabdus amnigenus Desulfobacca acetoxidans Methanogens Methanosarcina barkeri b mazei b Methanosaeta soehngenii b concilii a b

Yield a Km (g/mol ac.) (mM)

0.79 0.92 0.79 0.72–1.11 4.3–4.8 0.65–1.39 5.6 0.14–0.20 0.31–0.41

Vmax (µmol/min · g)

0.07–0.23

53

0.6 0.6

28 43

5.0

0.46–0.69 1.6–3.4 0.49–0.53 1.9

3.0

0.5

0.08–0.29 1.1–1.4 0.21–0.69 1.1–1.2

0.39–0.7 0.84–1.2

38 16

The yield is given in gram cell dry weight per mol. Several strains.

The aceticlastic sulfate-reducers that prefer freshwater conditions, such as Desulfoarculus baarsii [73], Desulfobacterium catecholicum [74], and Desulfococcus biacutus [75], show very poor growth with acetate. Only Desulfobacterium strain AcKo and Desulfotomaculum acetoxidans show good growth with acetate under mesophilic conditions (see Table 4). Unfortunately no Ks or Km values are available for these bacteria. Two abundant acetate-degrading sulfate-reducers, Desulforhabdus amnigenus and Desulfobacca acetoxidans, were isolated from sulfate-reducing bioreactors [76, 77]. The Michaelis-Menten parameters for D. amnigenus (KM = 0.2–1 mM, Vmax = 21–35 µmol · min–1 · g protein–1) and D. acetoxidans (KM = 0.1–1 mM, Vmax = 29–57 µmol · min–1 · g protein–1) were in the same range as or slightly better than those of most Methanosaeta species (KM = 0.4–1.2 mM, Vmax = 32–170 µmol · min–1 · g protein–1). This was also the case for the specific growth rate and the threshold value for acetate, which were 0.14–0.20 day –1 and