Anaerobe Tolerance to Oxygen and the Potentials of Anaerobic and Aerobic Cocultures for Wastewater Treatment

ANAEROBE TOLERANCE TO OXYGEN AND THE POTENTIALS OF ANAEROBIC AND AEROBIC COCULTURES FOR WASTEWATER TREATMENT

M.T. Kato ¹ , J.A. Field ² and G. Lettinga ²

¹Department of Civil Engineering, Federal University of Pernambuco,

Av. Acadêmico Hélio Ramos s/n, 50670-530 Recife PE, Brazil

Phone: [phone]/8221 - Fax: [phone] - E-mail: [e-mail]

²Department of Environmental Technology, Wageningen Agricultural University,

Bomenweg 2, 6702 HD Wageningen, The Netherlands.

(Received: June 11, 1997; Accepted: October 30, 1997)

Abstract - The anaerobic treatment processes are considered to be well-established methods for the elimination of easily biodegradable organic matter from wastewaters. Some difficulties concerning certain wastewaters are related to the possible presence of dissolved oxygen. The common belief is that anaerobes are oxygen intolerant. Therefore, the common practice is to use sequencing anaerobic and aerobic steps in separate tanks. Enhanced treatment by polishing off the residual biodegradable oxygen demand from effluents of anaerobic reactors, or the biodegradation of recalcitrant wastewater pollutants, usually requires sequenced anaerobic and aerobic bacteria activities. However, the combined activity of both bacteria can also be obtained in a single reactor. Previous experiments with either pure or mixed cultures showed that anaerobes can tolerate oxygen to a certain extent. The oxygen toxicity to methanogens in anaerobic sludges was quantified in batch experiments, as well as in anaerobic reactors. The results showed that methanogens have a high tolerance to oxygen. In practice, it was confirmed that dissolved oxygen does not constitute any detrimental effect on reactor treatment performance. This means that the coexistence of anaerobic and aerobic bacteria in one single reactor is feasible and increases the potentials of new applications in wastewater treatment.

Keywords: Oxygen toxicity; anaerobic sludge, methanogens, sequenced anaerobic-aerobic steps, cocultures, simultaneous anaerobic-aerobic activity, UASB and EGSB reactor, wastewater treatment.

INTRODUCTION

Adequate low-cost technology is essential for wastewater treatment. The anaerobic processes offer great potentials for wastewater treatment, since they have been successfully applied for a number of organic wastes (Ni and Nyns, 1993). They are suitable economical up-to-date alternatives that fulfill the following requirements: simplicity of design; the use of unsophisticated equipment or installations; low energy consumption; and high treatment efficiency (Lettinga and Hulshoff Pol, 1982a).

So far, however, the anaerobic treatment processes have only been considered well-established methods primarily for the elimination of easily biodegradable organic matter from wastewaters. Additionally, there are still some constraints which impair a broader application of their potentials. Modern high-rate anaerobic reactor systems usually operate at high hydraulic loading rates, resulting in a high intake of liquid influent. Some wastewaters may contain dissolved oxygen, which means that the potential danger of high levels of oxygen intake exists.

Oxygen is known as a potentially toxic compound during anaerobic treatment, especially for the methanogens, which are usually regarded as strict anaerobes (Hungate, 1969; Whitman et al., 1992). Consequently, reactor instability and low performance can result. This may explain why anaerobic and aerobic processes, when used in the biological treatment of wastewaters, are frequently utilized in separate sequential tanks for the removal of pollutants. A well-known example is the aerobic posttreatment commonly used to polish off the remaining biochemical oxygen demand (BOD) from effluents of anaerobic wastewater treatment. Other examples of sequenced anaerobic-aerobic processes needed are for the treatment of polychlorinated hydrocarbons (Fathepure and Vogel, 1991) and of adsorbable organic halogens in bleachery wastewaters (Jokela et al., 1993). Toxic aromatic pollutant mineralization, as well as the enhanced biodegradation of polyhydroxylated and chlorinated phenols, nitroaromatics and aromatic amines, is also provided by sequencing anaerobic and aerobic treatment steps (Field et al., 1995).

However, the combined activity of both anaerobic and aerobic bacteria can also be obtained in a single bioreactor. Anaerobic consortia were shown to be maintained in coculture with aerobic bacteria in oxygen-limited chemostats (Gottschal and Szewzyk, 1985; Gerritse et al., 1990; Gerritse and Gottschal, 1992). Also, some strains of methanogens were able to withstand long exposure to high levels of oxygen (Zehnder and Wuhrmann, 1977; Huser et al., 1982; Kiener and Leisinger, 1983). In practice, reactor upset was seldom reported when methanogenic sludges were exposed to air for short periods of time (Fields and Agardy, 1971; Bhuwapathanapun and Earle, 1975). This may be due to the rapid uptake of oxygen by aerobic and facultative bacteria, resulting in anaerobic microniches inside sludges. Thus, methanogens located inside sludges would be protected from contact with oxygen. These examples indicate that the best of anaerobic and aerobic processes can be further explored, and that the coexistence of cocultures in one single bioreactor opens up new possibilities to environmental technology. The objectives of this article are to present the effect of oxygen on anaerobic sludges, particularly on the tolerance of methanogens, and on anaerobic high-rate reactors and the potentials of anaerobic and aerobic cocultures for wastewater treatment.

ANAEROBIC TECHNOLOGY FOR WASTEWATER TREATMENT

Since the development of the high-rate and efficient anaerobic reactors, a large number of full-scale plants have been built (Iza et al., 1991; Lettinga and Hulshoff Pol, 1991; Lettinga et al., 1992). The development of the anaerobic processes and their successful application in the direct treatment of such wastewaters can mainly be attributed to the high retention of active biomass, which is one of the most important characteristics of the various high-rate systems developed (Iza et al., 1991). In these systems, the bacterial adhesion resulting in biofilms is the major mechanism which enables a high retention of dense and active biomass (Hulshoff Pol, 1989). Important examples of reactors with high retentions of biomass are the upflow anaerobic sludge bed (UASB) and the expanded granular sludge bed (EGSB). In these reactors, the biomass is immobilized as aggregates in the form of granules. A higher reactor biomass concentration of 30 to 50 g VSS/l (volatile suspended solids) can be reached or used with granular sludges (Rinzema, 1988). This means that higher maximum conversion rates can be expected in the UASB and EGSB processes. Since their earlier development, the UASB-like systems have been more widely applied in practice than other anaerobic systems (Lettinga and Hulshoff Pol, 1982b, 1991; Lettinga et al., 1992).

The UASB concept basically relies on the distinguishably high aggregation of anaerobic sludge that is typically observed to occur naturally under liquid upflow conditions (Lettinga and Vinken, 1980; Lettinga et al., 1980). The dense active sludge granules formed also have high settling characteristics and mechanical strength. Start-up of UASB reactors and granule formation have already been thoroughly investigated (Hulshoff Pol, 1989). The start-up and granulation were even shown with domestic sewage as the seed for the treatment of low-strength brewery wastewaters (Hanaoka et al., 1994). An adequate retention of a high amount of sludge is enhanced in the UASB reactor by a simple gas-solid-liquid (GSL) device installed in the upper part of the reactor. The result is usually an effective separation of the gas from the sludge and liquid.

The UASB reactor concept represents remarkable progress in the anaerobic treatment of wastewaters since, besides being simple to design, it can accommodate very high volumetric organic loads and provide high treatment performance (Lettinga et al., 1980; Lettinga and Hulshoff Pol, 1991). Several benefits can be derived from the UASB treatment of wastewaters compared with the other aerobic or anaerobic processes. For comparable treatment requirements, less reactor volume and space are needed for UASB plants and high-grade energy is produced from the biogas. Another advantage of the UASB is the relatively low cost of the technology requiring no sophisticated equipment. Additionally, the anaerobic granular sludge is generally well stabilized and significantly less excess sludge is produced compared, for instance, with that of the aerobic systems. Also, when stored at low temperatures, activity and settleability are maintained even after not being fed for long periods (Lettinga and Hulshoff Pol, 1982b; Lettinga et al., 1992). Nevertheless, some modifications of the UASB reactor concept have been proposed in order to improve its applicability (Lettinga and van Haandel, 1993). The conventional UASB reactor seldom utilizes effluent recirculation and a modification employing it resulted in the EGSB concept.

The characteristics of the EGSB reactor are very similar to those of the UASB reactor. However in an EGSB-type reactor, the granular sludge bed is expanded and the hydraulic mixing is intensified in order to improve the wastewater-biomass contact (de Man et al., 1988; van der Last and Lettinga, 1991). A higher superficial liquid velocity is achieved by applying effluent recirculation. In full-scale EGSB reactors, a sophisticated influent distribution system will be required. In the UASB reactor the sludge bed behaves more as a static bed because the liquid upflow velocity (V_up) is usually in the range of 0.5 to 1.5 m/h (de Man et al., 1988). In contrast, the EGSB utilizes V_up exceeding 5 to 6 m/h by using high effluent recirculation ratios combined with taller reactors. In general, a relatively high height-diameter ratio of 20 or higher is used (de Man et al., 1988; Rinzema, 1988; van der Last, 1991). Nonetheless, shallow reactors can also be used. The EGSB reactor utilizes a partially or fully expanded bed of granule sludge. In the latter case, the liquid phase is completely mixed (Rinzema, 1988).

OXYGEN-RELATED PROBLEMS IN THE ANAEROBIC TREATMENT OF WASTEWATER

Despite all the developments in anaerobic processes for the treatment of domestic sewage and a number of industrial wastewaters, one remaining question concerns the possible effect of dissolved oxygen entering the treatment system on reactor performance. A general belief is that anaerobic treatment of wastewaters can face a serious problem due to the possible presence of dissolved oxygen. Dissolved oxygen can be present in certain effluents, particularly those of low strength. Additionally, high hydraulic loading rates are usually applied in UASB and especially in EGSB reactors. This means that in some cases, high levels of oxygen intake can occur which would represent a potential danger to reactor stability. This danger is justified because oxygen is suspected to be a toxic compound during anaerobic treatment, especially for the end-of-food microorganisms, the acetogens and methanogens, which are usually regarded as strict anaerobes (Hungate, 1969; Whitman et al., 1992).

In fact, oxygen is a powerful reagent which generates potentially toxic radicals to all living cells, especially hydrogen peroxide and superoxide (Pfenning, 1978; Morris, 1979; Gottschalk and Peinemann, 1992). The toxic effect can damage the chromosomal DNA, as suggested for Roseburia cecicola which is a strict anaerobe considered to be intolerant to oxygen (Martin and Savage, 1988). Obligate anaerobic bacteria, unlike the aerobic and facultative bacteria, can be defined as those microorganisms unable to synthesize a respiratory chain with oxygen as the terminal electron acceptor and oxidize organic substrates to carbon dioxide and water. They are considered to live strictly without oxygen (Gottschalk and Peinemann, 1992). Aerobic and facultative bacteria are regarded as possessing appropriate protective mechanisms against the oxygen radicals. The main hypotheses for the protective mechanisms is the ability to produce two enzymes, superoxide dismutase (SOD) and catalase. SOD seems to be indispensable to all aerobes (Morris, 1979; Gottschalk and Peinemann, 1992), despite the claim that a few aerobes lack it (Fee, 1992). The total lack of SOD has also been suggested as the reason for the oxygen intolerance among the strict obligate anaerobes (Morris, 1979). Curiously enough, many obligate anaerobes do contain SOD (Table 1) and they can tolerate to some extent low levels of oxygen (Rolfe et al., 1978; Morris, 1979). These findings reveal that the obligate anaerobes differ in their sensitivity to oxygen, varying from those showing a strict intolerance to others possessing some intrinsic tolerance.

THE EFFECT OF OXYGEN ON ANAEROBES IN PURE CULTURES

Based on experiments with pure cultures, oxygen tension and redox potential were reported as affecting sensitivity. The effect of oxygen on 8 different species of anaerobic bacteria was shown to be related to the oxygen tension (0.2 to 3 atm), resulting in different oxygen sensitivity patterns for the microorganisms (Fredete et al., 1967). In another experiment, microorganisms which did not exhibit growth, even on reduced media, at oxygen tensions not greater than 0.5% were grouped as strict anaerobes, e.g., Clostridium haemolycum; at oxygen tensions greater than 0.5% up to as high as 2 to 8% were grouped as moderate anaerobes, e.g., Bacteroides fragilis; and in the absence of oxygen, but maximally at intermediate levels of 5 to 10%, were best described as microaerophiles, e.g., Vibrio sputorum and V. fetus (Loesche, 1969). The exclusion of molecular oxygen together with an environment of very low redox potential are postulated as being essential for the anaerobic microorganisms, especially the methanogens (Hungate, 1969, 1984; Gottschalk and Peinemann, 1992). However, some experiments showed the separate effects of oxygen and redox potential on anaerobic cultures (O’Brien and Morris, 1971; Walden and Hentges, 1975). The continuous anaerobic culture of Bacteroides fragilis, submitted to different redox conditions alone, revealed that there was no detectable change in the viable cell density. However, the introduction of oxygen at 10 to 100% atmospheric saturation during 6 h resulted in a steady decline in the viable cell density. The mechanism of oxygen-mediated inhibition was assumed to be due to the superoxide radicals formed. However, the effect was bacteriostatic rather than bactericidal. The explanation of the bacteriostatic effect was that strains of B. fragilis contained significant concentrations of superoxide dismutase, as parallel investigations showed (Onderdonk et al., 1976).

Microorganisms Enzyme levels (units/mg proteins) SOD catalase Period of O2 tolerance (h) Aerobes and facultatives Lactobacillus plantarum 49.7 32.2 >72 Escherichia coli (aerobically grown) 33.4 46.0 >72 Escherichia coli (anaerobically grown) 15.3 11.0 >72 Pseudomonas aeruginosa 10.9 112.9 >72 Micrococcus radiodurans 7.0 289.0 n.r. Saccharomyces cerevisiae 3.7 13.5 n.r. Mycobacterium sp. 2.9 2.7 n.r. Rhizobium japonicum 2.6 0.7 n.r. Halobacterium salinarium 2.1 3.4 n.r. Pseudomonas sp. 2.0 22.5 n.r. Escherichia coli 1.8 6.1 n.r. Obligate anaerobes Clostridium perfringens (clinical) 1.4 0 >72 Clostridium perfringens 0.4 0 >72 Bacteriodes fragilis (clinical) 6.8 7.1 48 Bifidobacterium adolescentis 0.3 0 48 Bacteriodes vulgatus 12.5 0 8 Propionibacterium acnes (clinical) 0.9 103.2 8 Propionibacterium acnes 0.3 136.2 8 Bacteriodes fragilis 7.0 15.2 4 Bacteriodes melaninogenicus 0 0 2 Eubacterium lentum 3.6 0 1 Fusobacterium nucleatum 0 0 1 Clostridium aminovalericum 0 0 0.75 Peptostreptococcus anaerobius 0 0 0.75 Butyribacterium rettgeri 1.6 0 n.r. Streptococcus lactis 1.4 0 n.r. Streptococcus faecalis 0.8 0 n.r. Zymobacterium oroticum 0.6 0 n.r. Streptococcus mutans 0.5 0 n.r. Streptococcus bavis 0.3 0 n.r. Streptococcus mitis 0.2 0 n.r. Veillonella alcalescens 0 0 n.r. Clostridium pasteurianum, sticklandii, lentoputrescens, cellobioparum, barkeri 0 0 n.r. Clostridium acetobutylicum 0 n.r. n.r. Clostridium sp. (strain M. C.) 0 0 n.r. Butyrivibrio fibrisolvens 0 0.1 n.r.

Microorganisms	Enzyme levels (units/mg proteins) SOD catalase		Period of O2 tolerance (h)
Aerobes and facultatives
Lactobacillus plantarum	49.7	32.2	>72
Escherichia coli (aerobically grown)	33.4	46.0	>72
Escherichia coli (anaerobically grown)	15.3	11.0	>72
Pseudomonas aeruginosa	10.9	112.9	>72
Micrococcus radiodurans	7.0	289.0	n.r.
Saccharomyces cerevisiae	3.7	13.5	n.r.
Mycobacterium sp.	2.9	2.7	n.r.
Rhizobium japonicum	2.6	0.7	n.r.
Halobacterium salinarium	2.1	3.4	n.r.
Pseudomonas sp.	2.0	22.5	n.r.
Escherichia coli	1.8	6.1	n.r.
Obligate anaerobes
Clostridium perfringens (clinical)	1.4	0	>72
Clostridium perfringens	0.4	0	>72
Bacteriodes fragilis (clinical)	6.8	7.1	48
Bifidobacterium adolescentis	0.3	0	48
Bacteriodes vulgatus	12.5	0	8
Propionibacterium acnes (clinical)	0.9	103.2	8
Propionibacterium acnes	0.3	136.2	8
Bacteriodes fragilis	7.0	15.2	4
Bacteriodes melaninogenicus	0	0	2
Eubacterium lentum	3.6	0	1
Fusobacterium nucleatum	0	0	1
Clostridium aminovalericum	0	0	0.75
Peptostreptococcus anaerobius	0	0	0.75
Butyribacterium rettgeri	1.6	0	n.r.
Streptococcus lactis	1.4	0	n.r.
Streptococcus faecalis	0.8	0	n.r.
Zymobacterium oroticum	0.6	0	n.r.
Streptococcus mutans	0.5	0	n.r.
Streptococcus bavis	0.3	0	n.r.
Streptococcus mitis	0.2	0	n.r.
Veillonella alcalescens	0	0	n.r.
Clostridium pasteurianum, sticklandii, lentoputrescens, cellobioparum, barkeri	0	0	n.r.
Clostridium acetobutylicum	0	n.r.	n.r.
Clostridium sp. (strain M. C.)	0	0	n.r.
Butyrivibrio fibrisolvens	0	0.1	n.r.

Table 1: Oxygen sensitivity of microorganisms related to the levels of the enzymes superoxide dismutase (SOD) and catalase (Rolfe et al., 1978; Morris, 1979)

n.r. = not reported

The most noteworthy group of obligate anaerobic bacteria for the wastewater treatment processes are the methanogens. Since the work on the isolation of pure cultures of Methanosarcina barkeri and Methanobacterium formicicum, the widespread view is that the methanogens are fastidious microorganisms requiring strict anaerobic conditions for growth and methane production. Afterwards, the isolation of Methanobacterium ruminantium added significantly to this belief due to the Hungate technique which is still employed today with some modifications (Zehnder et al., 1981). Possible traces of oxygen were eliminated by the addition of reducing agents such as cysteine or sulphide to poise the redox potential within the low range (Hungate, 1984). This is exemplified by Methanococcus voltae and Mc. vannielii which are considered highly sensitive to oxygen due to their lack of SOD (Kiener and Leisinger, 1983). Also, two thermophilic methanogens isolated from a 55?C anaerobic kelp digester were shown to immediately cease growth and methane production after being exposed to only 0.1% oxygen in the head space (Hungate, 1969). Methanobacterium ruminantium, M. mobile and Methanobacterium strain AZ were found to be highly sensitive to oxygen since their growth and methane production were completely prevented at 0.01 ppm dissolved oxygen (Zehnder and Wuhrmann, 1977). However, it was also shown that the sensitivity of methanogens to oxygen does not necessarily mean that the effect is bactericidal. The resistance to oxygen contact is demonstrated in the case of Methanobacterium strain AZ. The presence of 7 ppm dissolved oxygen during four days did not result in a die-off after the removal of all oxygen and restoration of the reducing conditions (Zehnder and Wuhrmann, 1977).

In fact, further investigations also demonstrated that at least some methanogens do have an intrinsic tolerance to oxygen. Only a slow decline was observed in the methane production of Methanobacterium strain M.o.H. when exposed to oxygen (Roberton and Wolfe, 1970). Also, the oxygen tolerance was observed in several bacteria belonging to the orders Methanomicrobiales and Methanobacterium bryantii. This was attributed to the SOD, which was detected and correlated with the defense against oxygen toxicity (Kirby et al., 1981; Kiener and Leisinger, 1983).

Ecosystems such as sludge digesters can be periodically subjected to stress by accidental entrance of air. This fact has been used as an argument in favour of the development of oxygen-tolerant methanogens isolated in pure or enriched cultures from such ecosystems (Huser et al., 1982; Kiener and Leisinger, 1983). Examples are Methanobacterium thermoautotrophicum, Methanobrevibacter arboriphilus and Methanosarcina barkeri which showed an ability to survive for hours in the presence of air without a decrease in the number of colony-forming units. These strains were originally isolated from sludge digesters. Methanobacterium strain AZ was also isolated in pure culture from sewage-digested sludge (Zehnder and Wuhrmann, 1977). In contrast, Methanococcus voltae and Mc. vannielii, which were killed without any lag phase upon contact with air, were originally isolated from sea and lake sediments where they were not exposed to oxygen. Methanothrix soehngenii could be enriched not only from several sludge digesters but also from aerobic samples of pretreated raw sewage. No lysed cell and similar rates of methane production compared with the controls were observed in experiments with pure oxygen, up to 48 h exposure (Huser et al., 1982). Methanobacterium and Methanosarcina are also commonly isolated from dry and oxic paddy soil demonstrating that they can survive under aerobic conditions between flooding periods. That methanogenesis starts within a few days after flooding was demonstrated in research with the isolated pure cultures of Methanosarcina barkeri strain Fusaro, Methanosarcina strain MVF4 and Methanobacterium strain HVF5 (Fetzer et al., 1993).

It was also suggested that besides the intrinsic tolerance due to the SOD that some methanogenic species possess, other factors could be responsible for protection against the oxygen. Methanosarcina barkeri strain Fusaro was shown to have a number of redox carriers which decreased the redox potential when chemical oxidant agents were used. However, this reducing capacity was not enough to avoid inhibition by oxygen at concentrations higher than 0.5% in the gas phase. Nonetheless, the capacity to adjust the redox potential in its own redox environment to a certain extent could also be one of the reasons for its good survival in dry and oxic soil (Fetzer and Conrad, 1993). The existence of cells in aggregates can also affect the oxygen tolerance. This seems to be the case for Methanosarcina which showed higher oxygen tolerance in cell aggregates than in dispersed cells. The arrangement in cell aggregates is postulated to provide protection of the cells against oxygen, since individual cells showed higher sensitivity to oxygen (Kiener and Leisinger, 1983). Methanosarcina mazei S-6 is normally considered to grow as large aggregates, allowing the sedimentation and permanence in a completely oxygen-free environment of natural sediments and in anaerobic reactors (Xun et al., 1988).

THE EFFECT OF OXYGEN ON ANAEROBES IN MIXED CULTURES

In the case of natural mixed cultures of anaerobic sludges or sediments, oxygen can incidentally come into contact with methanogens. Oxygen was detected in the gas of a number of anaerobic digesters (Scott, Williams, Whitmore and Lloyd, 1983). By measurements in situ, oxygen was found even in the rumen liquor and gas evolved from cattle, sheep and goats, and it is known that rumen microorganisms in pure cultures are killed by traces of oxygen (Scott, Yarlett, Hillman, Williams, Williams and Lloyd, 1983). Addition of oxygen at increased pressure in a sample of bovine rumen liquor showed that, in fact, methane production decreased. However, some minutes later after each addition of oxygen, methane-producing activity was partially reversible. Dissolved oxygen could not be detected in the samples exposed to oxygen. This was attributed to the presence of facultative bacteria in the rumen. The facultative bacteria can rapidly consume oxygen and create anaerobic microenvironments, which explains the survival of the methanogens in the bulk liquid phase (Scott, Williams, Whitmore and Lloyd, 1983; Scott, Yarlett, Hillman, Williams, Williams and Lloyd, 1983). Due to the role of the facultative bacteria, it is expected that each unit of dissolved oxygen introduced into the anaerobic mixed culture would be reflected in a unit of oxidized substrate. Since the oxygen dissolved in liquid is very low and the available substrate is relatively high, it would be unusual to detect oxygen in the gas phase of digesters, as reported by Scott, Williams, Whitmore and Lloyd (1983).

In fact, natural mixed cultures represent a complex consortium of microorganisms, where the consumption of oxygen in a single reactor can very likely occur due to the presence of facultative bacteria in natural anaerobic sludges. Regardless of the intrinsic tolerance that has been reported in some strict obligate anaerobes like the methanogens, the aerobic or facultative bacteria are probably the most important factor for the protection of methanogens against O₂ exposure in anaerobic sludges. In the presence of aerobic or facultative substrates, the oxygen can undoubtedly be rapidly consumed. This has been demonstrated during ethanol production by a coimmobilized mixed culture of the aerobic fungus Aspergillus awamori and the anaerobic bacterium Zymomonas mobilis, under highly aerobic conditions (Tanaka et al., 1986). Defined mixed cultures of an obligate aerobic Pseudomonas testosteroni strain and an anaerobic Veillonella alcalescens strain also grown in a microaerophilic chemostat, were shown to coexist and compete for common substrates (Gerritse et al., 1990). The cocultivation of the facultative anaerobes Vibrio sp. and the strictly anaerobic sulphate-reducing bacteria Desulfovibrio HL21 was demonstrated in glucose chemostats under oxygen-limiting conditions (Gottschal and Szewzyk, 1985).

Marine sediments are generally reducing environments covered only by thin, highly oxidized surface layers (Jørgensen, 1977a, 1977b; Battersby et al., 1985). The obligate anaerobe Desulfovibrio spp. and the sulphide oxidizing bacteria Beggiatoa spp were found present near the surface layers. The occurrence of pyrite (FeS₂) in the oxidized layer can only be explained on the one hand by the rapid oxidation of Fe⁺², and on the other hand by the formation of H₂S from the sulphate within reducing environments (Jørgensen, 1977a, 1977b).

Oxygen limitation within natural biofilms may also occur due to the biofilm itself. The biofilm’s thickness and architecture represent a physical barrier for the diffusion of oxygen, enhancing the creation of segregated zones of lowered oxygen concentrations towards the centre. Since very complex systems can exist in separate layers or microniches, growth and competition occur between different groups of microorganisms. In this case, the oxygen limitation within biofilms depends on the available substrate and oxygen concentration in the bulk liquid phase. Nitrification and denitrification are examples of aerobic and anaerobic layer formation, respectively. Several investigators have used microsensors to determine the oxygen profiles within the biofilms, usually resulting in gradients from the oxic outer layers to the anoxic layers in the centre (Wittler et al., 1986; Peters et al., 1987; Revsbech et al., 1988, 1989; de Beer, 1990; Hooijmans, 1990; Wijffels et al., 1991). A general finding is that oxygen rarely penetrates more than a few hundred mm due to the oxygen uptake and diffusion limitation. Consequently, anaerobic microorganisms located inside the biofilms would be adequately protected against contact with oxygen, allowing the occurrence of denitrification, sulphate reduction or methanogenesis (Wijkieks et al., 1991). The occurrence of microniches of strict anaerobes in aerobic environments is best illustrated by aerobic sludge samples or by the well-aerated flocs in activated sludge. The fact that methanogens are present in aerobic sludge was demonstrated when Methanothrix soehngenii could be enriched from such samples (Huser et al., 1982). Activated sludge was used as seed for the start-up of two UASB reactors (Wu et al., 1987). Granule formation in the UASB reactors was very interesting. Methanobacterium, Methanococcus and Methanosarcina were observed in both the original aerobic activated sludge flocs and the granular sludges formed. According to fluorescent microscopic examination, several anaerobic nuclei may have existed deep inside the flocs. Granulation also occurred in a UASB reactor seeded with raw wastes from an activated sludge tank (Noyola and Moreno, 1994). The start-up of an anaerobic filter with activated sludge has also been described (Campos, 1990).

THE EFFECT OF OXYGEN ON METHANOGENS IN GRANULAR SLUDGES

The previous studies, in either pure or mixed cultures, show that at least some methanogens have some tolerance to oxygen exposure. Start-up of UASB reactors was successfully conducted with sludges obtained from aerobic activated sludge plants, indicating the presence of anaerobic bacteria in aerobic environments.

To quantify the toxicity of oxygen to methanogens in granular sludges, five sludges with distinct properties collected from UASB and EGSB reactors were used (Kato et al., 1993a, 1993b). The results of batch experiments reveal that methanogens in granular sludge have a high tolerance to oxygen. The concentration of oxygen causing 50% inhibition of methanogenic activity was between 7% and 41% of the oxygen added to the head space of flasks which corresponded to 0.05 and 6 mg/l of the dissolved oxygen prevailing in the media. In order to determine the mechanisms of O₂ tolerance, the roles of granule size, respiration rate of the facultative bacteria, shaking regime and substrate were investigated. The respiration rate of the facultative bacteria present in the granular sludges was the most important mechanism of oxygen tolerance since it was well correlated with highly tolerant sludges. The absence of facultative substrate for respiring O₂ drastically decreased the oxygen tolerance. There was no correlation of O₂ tolerance with granule size and shaking only affected the least tolerant sludges.

A hypothesis was formulated to describe the oxygen tolerance of methanogens in granular sludges. Facultative bacteria rapidly consume oxygen, creating anaerobic microniches inside the granules where the methanogens are well protected against contact with oxygen. This hypothesis can be justified by the poor penetration of oxygen into biofilms. In the literature, O₂ penetration into actively respiring aerobic biofilms has been reported to reach a depth of only 100 to 300 mm. (de Beer, 1990; Hooijmans, 1990). The diameter of the granules, in great excess of that required for creating anaerobic zones, would explain the lack of correlation found between oxygen tolerance and sludge granule size. When substrate was not supplied, facultative respiration was less active. Thus, oxygen could penetrate deeper into the granules and come into contact with the methanogens, which would explain the enhanced toxicity. Some tolerance was still evident even in the absence of substrate, indicating that methanogens do have a limited intrinsic tolerance to oxygen. As reported previously, several methanogens are reported to contain the enzyme SOD which neutralizes toxic oxygen radicals (Rolfe et al., 1978; Morris, 1979). Moreover, the existence of redox carriers and of growth in aggregates were also reported to be protective factors for strict anaerobes in pure cultures (Kiener and Leisinger, 1983; Fetzer and Conrad, 1993).

These results have important implications for the anaerobic treatment of wastewaters. Oxygen toxicity is not expected to occur because wastewaters usually contain a large excess of BOD, required by facultative bacteria to consume the dissolved oxygen present in the wastewaters. Under normal conditions, only dissolved oxygen will enter into the system with concentrations below 10 mg/l, while even low-strength wastewaters usually contain at least several hundred milligrams per litre of BOD. In practice, the non-detrimental effect of dissolved oxygen was also confirmed in UASB and EGSB studies, since chemical oxygen demand (COD) removal efficiencies and values of effluent redox were very similar in parallel ethanol reactors, one of which was fed with and another without dissolved oxygen in the influent (Kato, Field, Kleerebezem and Lettinga, 1994; Kato, Field, Versteeg and Lettinga, 1994).

Since even higher levels of oxygen can be tolerated, a second implication is that oxygen can be added to anaerobic reactors to enhance the degradation of recalcitrant pollutants. Nonetheless, despite the non-toxicity of oxygen to natural mixed cultures in granular sludges used in wastewater treatment, there is still a concern about the substrate competition between methanogens and facultative bacteria. If too much oxygen were fed into the reactor, then all the substrate would be consumed by the facultative bacteria since they have much higher specific activities and growth rates than those of the methanogens. Methanogens without access to substrate will be outcompeted. Thus, the competition can perhaps be regulated by oxygen supply.

In a study conducted to investigate the competition for substrate in anaerobic-aerobic cocultures, it was shown that methane production and oxygen consumption occurred simultaneously (Kato, Field, Versteeg and Lettinga, 1994). First of all, a distinction was made between O₂ intolerant and O₂ tolerant sludges. In the intolerant sludges, each unit of O₂ was reflected in a unit of BOD stolen from methanogens. This would signify that no methane could be expected to be produced if the oxygen supply were in excess of the BOD. Thus, it could be predicted that substrate competition would easily be won by the facultative bacteria in unlimited O₂ conditions. In the tolerant sludges, methane production could be demonstrated even when oxygen was not limiting, suggesting that substrate competition occurs between methanogens and facultative bacteria, even in well-aerated environments. At dissolved oxygen concentrations of 23, 2 and 1 mg O₂/l, 2%, 15% and 25% of the substrate were converted to methane, respectively. The occurrence of methanogenesis under unlimited O₂ conditions is hypothesized to occur if there is a better transport of substrate to positions deep inside the granule than of oxygen. Mass transfer by diffusion depends on the concentration gradient. Since in practice oxygen has a maximum solubility of less than 10 mg/l, this indicates that at most only weak gradients can be developed. Substrate, on the other hand, has unlimited solubility and was usually applied at 1000 mg COD/l or higher. Thus, much higher diffusion gradients could be established for the substrate.

Additionally, the results showed that the competition for substrate during prolonged periods of O₂ exposure (18 days) eventually resulted in the development of complex mixed cultures where at least three trophic groups could be identified: the methanogens, the facultative heterotrophic bacteria and the methanotrophic bacteria. Nevertheless, it was confirmed that since competition for substrate can be regulated, oxygen can be added to anaerobic reactors. In the literature, examples are reported which give evidence of the ability of anaerobic-aerobic cocultures to degrade recalcitrant pollutants in one bioreactor, instead of using separate anaerobic and aerobic phases (Beunink and Rehm, 1988, 1990; Gerritse and Gottschal, 1992).

THE POTENTIALS OF THE PRACTICAL APPLICATION OF COCULTURES OF ANAEROBIC AND AEROBIC BACTERIA

Previously, the occurrence of methanogenesis in oxygen-limiting chemostats was also demonstrated in cocultures. The combined activity of facultative bacteria and methanogens competing for the same substrate was evidenced by simultaneous oxygen consumption and methane production. A gradual increase in oxygen supply did not significantly interfere with methanogenesis. It is very interesting that at a low oxygen supply, methane production was stimulated by 20% (Gerritse et al., 1990). The increased methanogenesis in the presence of traces of oxygen was also reported for a reactor fed with domestic sewage or algal biomass, and containing sludges from anaerobic reactors (Pirt and Lee, 1983; Scott, Williams, Whitmore and Lloyd, 1983). This was explained by the possible stimulation of acetate production for the methanogens by the facultative bacteria. Further experiments with several mixed cultures of methanogenic and aerobic bacteria clearly indicated that they can share the same habitat under oxygen-limited conditions. Cocultures of Methanobacterium formicicum or Methanosarcina barkeri with the aerobic heterotrophic Comamonas testosteroni were evident in continuous chemostats fed with small amounts of oxygen (Gerritse and Gottschal, 1993).

From the previous studies, it turns out that the coexistence of anaerobic and aerobic or facultative bacteria in a single reactor, competing for the same substrates, can be regulated through a balanced oxygen supply. Separate anaerobic and aerobic phases have commonly been used for removal of pollutants. However, new possibilities are the use of a single-phase anaerobic-aerobic wastewater treatment system. In practice, these possibilities may include polishing off the residual BOD and the biodegradation of polychlorinated hydrocarbons and adsorbable organic halogens (AOX).

The aerobic bacteria are usually used in a posttreatment step to polish off the residual BOD after the anaerobic treatment of wastewaters (Lettinga and Hulshoff Pol, 1991). Since the occurrence of methanogenesis has been demonstrated under oxygen-limited conditions (Pirt and Lee, 1983; Scott, Williams, Whitmore and Lloyd, 1983; Gerritse et al., 1990; Gerritse and Gottschal, 1993), the residual BOD can be eliminated in the same reactor where anaerobic treatment is occurring. Also, in some cases, the residual BOD is caused by toxic micropollutants which are not substrates for anaerobic bacteria. Examples are resin acids in the forest industry wastewaters (Sierra-Alvarez, 1990; Field, 1989; Lettinga et al., 1991; Sierra-Alvarez et al., 1993). Thus, the addition of small amounts of oxygen to an anaerobic reactor can eliminate them too. Polychlorinated hydrocarbons are not known to be degraded by aerobic bacteria or mineralized by anaerobic bacteria. These pollutants have only been eliminated in sequenced anaerobic-aerobic reactors since reductive dechlorination is first necessary to yield hydrocarbons, which are then mineralized under aerobic conditions (Fathepure and Vogel, 1991). The use of sequenced anaerobic-aerobic reactors was also shown to be more efficient to remove AOX from chlorolignins in bleachery wastewaters, compared with the use of aerobic or anaerobic treatment alone (Jokela et al., 1993). Also, the application of simultaneous anaerobic and aerobic degradation was demonstrated for the elimination of recalcitrant pollutants under limited oxygen conditions. Many recalcitrant pollutants cannot be fully mineralized under aerobic or anaerobic conditions, but require the sequenced activity of anaerobic and aerobic bacteria. Immobilized cocultures of Enterobacter cloacae and Alcaligenes sp. were able to degrade 1,1,1-trichloro-2,2-bis-(4-chlorophenyl)-ethane (DDT) or 4-chloro-2-nitrophenol (CNP) (Beunink and Rehm, 1988, 1990).

Mineralization of the herbicide 2,3,6-trichlorobenzoic acid was also demonstrated in a single microaerophilic chemostat. An anaerobic enrichment resulted due to the aerobic activity of Pseudomonas aeruginosa JB2 in oxygen consumption (Gerritse and Gottschal, 1992). These previous possibilities may, therefore, represent promising applications for the environmental technology. A biofilm type worth considering for use in such single reactors is the granular sludge. Anaerobic microenvironments would be better protected and maintained in granular sludge because oxygen would hardly penetrate deep into the biofilm, contrary to what occurs in dispersed sludge. A model for the creation of anaerobic zones in granular sludges inside an aerated reactor is shown in Figure 1.

Figure 1: Model for the degradation of simple substrates by cocultures of anaerobic and facultative bacteria in granular sludge inside an aerated reactor.

CONCLUSIONS

It was shown that, contrary to the commonly held belief anaerobes, particularly the methanogens in sludges from wastewater treatment plants, have a high tolerance to oxygen. The activity of aerobic or facultative bacteria was shown to be the most important mechanism of oxygen tolerance. Some methanogens even present intrinsic tolerance. In practice, oxygen toxicity would never occur in anaerobic bioreactors under normal conditions where only dissolved oxygen would enter the system. The simultaneous coexistence of methanogens and aerobic bacteria, shown by the concurrent uptake of oxygen and methane production, indicates the evidence that anaerobic and aerobic cocultures can be maintained in a single reactor. This possibility adds to the arsenal of biodegradable capacities and potentially opens up new applications in environmental technology. Air or oxygen can be added directly to anaerobic reactors like UASB or EGSB to enhance the treatment of several wastewaters containing simple or recalcitrant pollutants.

• Battersby, N. S.; Malcolm, S. J.; Brown, C. M. and Stanley, S. O., Sulphate Reduction in Oxic and Sub-oxic North-East Atlantic Sediments. FEMS Microbiol. Ecol. 31: 225-228 (1985).
• Beunink, J. and Rehm, H.-J., Synchronous Anaerobic and Aerobic Degradation of DDT by an Immobilized Mixed Culture System. Appl. Microbiol. Biotechnol. 29: 72-80 (1988).
• Beunink, J. and Rehm, H.-J., Coupled Reductive and Oxidative Degradation of 4-chloro-2-nitrophenol by a Co-immobilized Mixed Culture System. Appl. Microbiol. Biotechnol. 34: 108-115 (1990).
• Bhuwapathanapun, S. and Earle, R.L., Some Effects of Aeration on Anaerobic Digestion, pp. 1-14. In Proc. of the Biotechnology Conference. Massey University, Palmerston, New Zealand (1975).
• Campos, J.R., Projeto e Operaçăo de Filtros Anaeróbios para Tratamento de Efluentes Líquidos Industriais, pp. 133-177. In Proc. of the Regional Workshop and Conference on Anaerobic Treatment of Wastewater in Latin America. Mexico City, Mexico (1990).
• de Beer, D. Microelectrode Studies in Biofilms and Sediments. Ph.D. diss., Universiteit van Amsterdam. Amsterdam, The Netherlands (1990).
• de Man, A.W.A.; van der Last, A.R.M. and Lettinga, G., The Use of EGSB and UASB Anaerobic Systems for Low Strength Soluble and Complex Wastewaters at Temperatures Ranging from 8 to 30°C, pp. 197-209. In Hall, E. R. and Hobson, P. N. (eds.), Proc. of the Fifth International Symposium on Anaerobic Digestion, Bologna, Italy. Pergamon, Oxford, England (1988).
• Fathepure, B. Z. and Vogel, T. M., Complete Degradation of Polychlorinated Hydrocarbons by a Two-stage Biofilm Reactor. Appl. Environ. Microbiol. 57: 3418-3422 (1991).
• Fee, J.A., Is Superoxide Important in Oxygen Poisoning? Trends Biochem. Sci. 7: 84-86 (1992).
• Fetzer, S. and Conrad, R., Effect of Redox Potential on Methanogenesis by Methanosarcina barkeri Arch. Microbiol. 160: 108-113 (1993).
• Fetzer, S.; Bak, F. and Conrad, R., Sensitivity of Methanogenic Bacteria from Paddy Soil to Oxygen and Desiccation. FEMS Microbiol. Ecol. 12: 107-115 (1993).
• Field, J.A., The Effect of Tannic Compounds on Anaerobic Wastewater Treatment. Ph.D. diss., Landbouwuniversiteit. Wageningen, The Netherlands (1989).
• Field, J.A.; Stams, A.J.M.; Kato, M. and Schraa, G., Enhanced Biodegradation of Aromatic Pollutants in Cocultures of Anaerobic and Aerobic Bacterial Consortia. Antonie van Leeuwenhoek 67: 47-77 (1995).
• Fields, M. and Agardy, F. J., Oxygen Toxicity in Digesters, pp. 284-293. In Proc. of the 26th Industrial Wastes Conference. Purdue University, West Lafayette, U.S.A. (1971).
• Fredete, V.; Plante, C. and Roy, A., Numerical Data Concerning the Sensitivity of Anaerobic Bacteria to Oxygen. J. Bacteriol. 94: 2012-2017 (1967).
• Gerritse, J. and Gottschal, J.C., Mineralization of the Herbicide 2,3,6-trichlorobenzoic Acid by a Co-culture of Anaerobic and Aerobic Bacteria. FEMS Microbiol. Ecol. 101: 89-98 (1992).
• Gerritse, J. and Gottschal, J.C., Two-membered Mixed Cultures of Methanogenic and Aerobic Bacteria in O₂-limited Chemostats. J. Gen. Microbiol. 139: 1853-1860 (1993).
• Gerritse, J.; Schut, F. and Gottschal, J.C., Mixed Chemostat Cultures of Obligately Aerobic and Fermentative or Methanogenic Bacteria Grown Under Oxygen-limiting Conditions. FEMS Microbiol. Lett. 66: 87-94 (1990).
• Gottschal, J.C. and Szewzyk, R., Growth of a Facultative Anaerobe Under Oxygen-limiting Conditions in Pure Culture and in Co-culture with a Sulfate-reducing Bacterium. FEMS Microbiol. Ecol. 31: 159-170 (1985).
• Gottschalk, G. and Peinemann, S., The Anaerobic Way of Life, pp. 300-311. In Balows, A., Tramper, H. G., Dworkin, M., Harder, W. and Schleifer, K.-H. (eds.), The Prokaryotes. A Handbook on the Biology of Bacteria: Ecophysiology, Isolation, Identification, Applications, vol. I. 2nd ed. Springer-Verlag, New York (1992).
• Hanaoka, T.; Nakahata, S.; Shouji, T.; Sasaki, M., Maaskant, W. and Wildschut, L., Granular Sludge Formation from Digested Sewage Sludge in a UASB Reactor Treating Low Strength Brewery Wastewater, pp. 518-527. In Proc. of the Seventh International Symposium on Anaerobic Digestion. Cape Town, South Africa (1994).
• Hooijmans, C.M., Diffusion Coupled with Bioconversion in Immobilized Systems. Ph.D. diss., Technische Universiteit. Delft, The Netherlands (1990).
• Hulshoff Pol, L.W., The Phenomenon of Granulation of Anaerobic Sludge. Ph.D. diss., Landbouwuniversiteit. Wageningen, The Netherlands (1989).
• Hungate, R.E., A Roll Tube Method for Cultivation of Strict Anaerobes, pp. 117-132. In Norris, J. B. and Ribbons, D. W. (eds.), Methods in Microbiology, vol. 3B. Academic Press, New York (1969).
• Hungate, R. E., Biochemistry and Microbiology of Anaerobic Digestion. Revista de Microbiologia, Săo Paulo 15: 278-292 (1984).
• Huser, B.A.; Wuhrmann, K. and Zehnder, A.J.B., Methanothrix soehngenii gen. nov. sp. nov., A New Acetotrophic Non Hydrogen Oxidizing Methane Bacterium. Arch. Microbiol. 132: 1-9 (1982).
• Iza, J.; Colleran, E.; Paris, J. M. and Wu, W.-M., International Workshop on Anaerobic Treatment Technology for Municipal and Industrial Wastewaters: Summary Paper. Wat. Sci. Technol. 24: 1-16 (1991).
• Jokela, J.K.; Laine, M.; Ek, M. and Salkinoja-Salonen, M., Effect of Biological Treatment on Halogenated Organics in Bleached Kraft Pulp Mill Effluents Studied by Weight Distribution Analysis. Environ. Sci. Technol. 27: 547-557 (1993).
• Jorgensen, B.B., Bacterial Sulfate Reduction Within Reduced Microniches of Oxidized Marine Sediments. Mar. Biol. 41: 7-17 (1977a).
• Jorgensen, B.B., Distribution of Colorless Sulfur Bacteria (Beggiatoa spp) in a Coastal Marine Sediment. Mar. Biol. 41: 19-28 (1977b).
• Kato, M.T.; Field, J.A. and Lettinga, G., High Tolerance of Methanogens in Granular Sludge to Oxygen. Biotechnol. Bioeng. 42: 1360-1366 (1993a).
• Kato, M.T.; Field, J.A. and Lettinga, G., Methanogenesis in Granular Sludge Exposed to Oxygen. FEMS Microbiol. Lett. 114: 317-324 (1993b).
• Kato, M.T.; Field, J. A.; Kleerebezem, R. and Lettinga, G., Treatment of Low Strength Soluble Wastewaters in UASB Reactors. J. Ferment.Bioeng. 77 (6): 679-686 (1994).
• Kato, M.T.; Field, J.A.; Versteeg, P. and Lettinga. G., Feasibility of Expanded Granular Sludge Bed Reactors for the Anaerobic Treatment of Low-Strength Soluble Wastewaters. Biotechnol. Bioeng. 44: 469-479 (1994).
• Kiener, A. and Leisinger, T., Oxygen Sensitivity of Methanogenic Bacteria. Syst. Appl. Microbiol. 4: 305-312 (1983).
• Kirby, T.W.; Lancaster Jr., J.R. and Fridovich, I., Isolation and Characterization of the Iron-containing Superoxide Dismutase of Methanobacterium bryantii Arch. Biochem. Biophysics 210: 140-148 (1981).
• Lettinga, G. and Hulshoff Pol, L.W., Anaerobic Wastewater Treatment Technology with Emphasis to Upflow Anaerobic Sludge Bed (UASB) Reactor Systems, pp. 21-39. In Anaerobic Reactor Technology, Wageningen Agricultural University. Wageningen, The Netherlands (1982a).
• Lettinga, G. and Hulshoff Pol, L.W., Wastewater Treatment and Resource Recovery Systems, the Place of Anaerobic Treatment and Digestion Processes, pp. 1-6. In Anaerobic Reactor Technology, Wageningen Agricultural University. Wageningen, The Netherlands (1982b).
• Lettinga, G. and Hulshoff Pol, L.W., UASB-process Design for Various Types of Wastewaters. Wat. Sci. Technol. 24: 87-107 (1991).
• Lettinga, G. and van Haandel, A.C., Anaerobic Digestion for Energy Production and Environmental Protection, pp. 817-839. In Johansson, T. B., Kelly, H., Reddy, A. K. N. and Williams, R. H. (eds.), Renewable Energy. Sources for Fuel and Electricity. Island Press, Washington, D.C. (1993).
• Lettinga, G. and Vinken, J.N., Feasibility of the Upflow Anaerobic Sludge Blanket (UASB) Process for the Treatment of Low Strength Wastes, pp. 625-634. In Proc. of the 35th Industrial Waste Conference. Purdue University, West Lafayette, U.S.A. (1980).
• Lettinga, G.; Field, J.A.; Sierra-Alvarez, R.; van Lier, J.B. and Rintala, J., Future Perspectives for the Anaerobic Treatment of Forest Industry Wastewaters. Wat. Sci. Technol. 24: 91-102 (1991).
• Lettinga, G.; Rinzema, A. and Hulshoff Pol, L.W., Possibilities and Potentials of Anaerobic Wastewater Treatment with Emphasis on the UASB-system, pp. 73-90. In Baader, W. (ed.), Proc. of the Third Workshop of the Working Group on Biogas Production Technologies. CNREE Network on Biogas Production and Conversion for Energy. Braunschweig, Germany (1992).
• Lettinga, G.; van Velsen, A.F.M.; Hobma, S.W.; de Zeeuw, W. and Klapwijk, A., Use of the Upflow Sludge Blanket (USB) Reactor Concept for Biological Wastewater Treatment, Especially for Anaerobic Treatment. Biotechnol. Bioeng. 22: 699-734 (1980).
• Loesche, W.J., Oxygen Sensitivity of Various Anaerobic Bacteria. Appl. Microbiol. 18: 723-727 (1969).
• Martin, J.H. and Savage, D.C., Degradation of DNA in Cells and Extracts of the Obligately Anaerobic Bacterium Roseburia cecicola Upon Exposure to Air. Appl. Environ. Microbiol. 54: 1619-1621 (1988).
• Morris, J.G., Nature of Oxygen Toxicity in Anaerobic Microorganisms, pp. 149-162. In Shilo, M. (ed.), Strategies of Microbial Life in Extreme Environments. Verlag Chemie, Weinheim, Germany (1979).
• Ni, J.-Q. and Nyns, E.-J., Biomethanation - a Developing Technology in Latin America. Catholic University of Louvain and Bremen Overseas Research and Development Association. Druckerei Verwohlt, Bremen, Germany (1993).
• Noyola, A. and Moreno, G., Granule Production from Raw Waste Activated Sludge, pp. 765-774. In Proc. of the Seventh International Symposium on Anaerobic Digestion. Cape Town, South Africa (1994).
• O'Brien, R.W. and Morris, J.G., Oxygen and the Growth and Metabolism of Clostridium acetobutylicum J. Gen. Microbiol. 68: 307-318 (1971).
• Onderdonk, A.B.; Johnston, J.; Mayhew, J. W. and Gorbach, S.L., Effect of Dissolved Oxygen and E_h on Bacteroides fragilis During Continuous Culture. Appl. Environ. Microbiol. 31: 168-172 (1976).
• Peters, A.C.; Wimpenny, J.W.T. and Coombs, J.P., Oxygen Profiles in, and in the Agar Beneath, Colonies of Bacillus cereus, Staphylococcus albus and Escherichia coli J. Gen. Microbiol. 133: 1257-1263 (1987).
• Pfenning, N., Formation of Oxygen and Microbial Processes Establishing and Maintaining Anaerobic Environment, pp. 137-148. In: Shilo, M. (ed.), Strategies of Microbial Life in Extreme Environments. Verlag Chemie, Weinheim, Germany (1978).
• Pirt, S.J. and Lee, Y.K., Enhancement of Methanogenesis by Traces of Oxygen in Bacterial Digestion of Biomass. FEMS Microbiol. Lett. 18: 61-63 (1983).
• Revsbech, N.P.; Christensen, P.B.; Nielsen, L.P. and Sřrensen, J., Denitrification in a Trickling Filter Biofilm Studied by a Microsensor for Oxygen and Nitrous Oxide. Wat. Res. 23: 867-871 (1989).
• Revsbech, N.P.; Nielsen, L.P.; Christiensen, P.B. and Sřrensen, J., Combined Oxygen and Nitrous Oxide Microsensor for Denitrification Studies. Appl. Environ. Microbiol. 54: 2245-2249 (1988).
• Rinzema, A., Anaerobic Treatment of Wastewater with High Concentrations of Lipids or Sulfate. Ph.D. diss., Landbouwuniversiteit. Wageningen, The Netherlands (1988).
• Roberton, A.M. and Wolfe, R.S., Adenosine Triphosphate Pools in Methanobacterium J. Bacteriol. 102: 43-51 (1970).
• Rolfe, R.D.; Hentges, D.J.; Campbell, B.J. and Barret, J.T., Factors Related to the Oxygen Tolerance of Anaerobic Bacteria. Appl. Environ. Microbiol. 36: 306-313 (1978).
• Scott, R.I.; Williams, T.N.; Whitmore, T.N. and Lloyd, D., Direct Measurement of Methanogenesis in Anaerobic Digestors by Membrane Inlet Mass Spectrometry. Eur. J. Appl. Microbiol. Biotechnol. 18: 236-241 (1983).
• Scott, R.I.; Yarlett, N.; Hillman, K.; Williams, T.N.; Williams, A.G. and Lloyd, D., The Presence of Oxygen in Rumen Liquor and its Effects on Methanogenesis. J. Appl. Bacteriol. 55: 143-149 (1983).
• Sierra-Alvarez, R., The Role of Natural Wood Constituents on The Anaerobic Treatability of Forest Industry Wastewaters. Ph.D. diss., Landbouwuniversiteit. Wageningen, The Netherlands (1990).
• Sierra-Alvarez, R.; Field, J.A.; Kortekaas, S. and Lettinga, G., Overview of the Anaerobic Toxicity Caused by Organic Forest Industry Wastewater Pollutants. Paper presented at the 4th IAWQ Symposium on Forest Industry Wastewaters. Tampere, Finland (1993).
• Tanaka, H.; Kurosawa, H. and Murakami, H., Ethanol Production from Starch by a Coimmobilized Mixed Culture System of Aspergillus awamori and Zymomonas mobilis Biotechnol. Bioeng. 28: 1761-1768 (1986).
• van der Last, A.R.M., Anaerobe Behandeling van Voorbezonken Rioolwater met het Geexpandeerde Korrelslibbed (EGSB)- en Fluidised Bed (FB)-proces. Samenvattend rapport. Vakgroep Milieutechnologie, Landbouwuniversiteit Wageningen. Wageningen, The Netherlands (1991).
• van der Last, A.R.M. and Lettinga, G., Anaerobic Treatment of Domestic Sewage Under Moderate Climatic (Dutch) Conditions Using Upflow Reactors at Increased Superficial Velocities, pp. 153-154. In Proc. of the Sixth International Symposium on Anaerobic Digestion. Săo Paulo, Brazil (1991).
• Walden, W.C. and Hentges, D.J., Differential Effects of Oxygen and Oxidation-Reduction Potential on the Multiplication of Three Species of Anaerobic Intestinal Bacteria. Appl. Microbiol. 30: 781-785 (1975).
• Whitman, W.B.; Bowen, T.L. and Boone, D.R., The Methanogenic Bacteria, pp. 719-767. In Balows, A., Tramper, H. G., Dworkin, M., Harder, W. and Schleifer, K.-H. (eds.), The Prokaryotes. A Handbook on the Biology of Bacteria: Ecophysiology, Isolation, Identification, Applications, vol. I. 2nd ed. Springer-Verlag, New York (1992).
• Wijffels, R.H.; Eekhof, M.R.; Tramper, J.; de Beer, D. and van den Heuvel, J.C., Growth and Substrate Consumption by Immobilized Nitrobacter agilis: Validation of a Dynamic Model, pp. 697-699. In Proc. of the International Symposium of Environmental Biotechnology, vol. II. Royal Flemish Society of Engineers, Oostende, Belgium (1991).
• Wijkieks, A.M.; van Loodsdrecht, M.C. and Heijnen, J.J., Determination of Oxygen Profiles in Biofilms on Small Suspended Particles from an Air Lift Reactor, pp. 695-696. In Proc. of the International Symposium of Environmental Biotechnology, vol. II. Royal Flemish Society of Engineers, Oostende, Belgium (1991).
• Wittler, R.; Baumgartl, H.; Lübbers, D.W. and Schügerl, K., Investigations of Oxygen Transfer Into Penicillium chrysogenum Pellets by Microprobe Measurements. Biotechnol. Bioeng. 28: 1024-1036 (1986).
• Wu, W.; Hu, J.; Gu, X.; Zhao, Y.; Zhang, H. and Gu, G., Cultivation of Anaerobic Granular Sludge in UASB Reactors with Aerobic Activated Sludge as Seed. Wat. Res. 21: 789-799 (1987).
• Xun, L.; Boone, D.R. and Mah, R.A., Control of the Life Cycle of Methanosarcina mazei S-6 by Manipulation of Growth Conditions. Appl. Environ. Microbiol. 52: 2064-2068 (1988).
• Zehnder, A.J.B. and Wuhrmann, K., Physiology of a Methanobacterium Strain AZ. Arch. Microbiol. 111: 199-205 (1977).
• Zehnder, A.J.B.; Ingvorsen, K. and Marti, T., Microbiology of Methane Bacteria, pp. 45-68. In Hughes, D. E., Stafford, D. A., Wheatley, B. I., Lettinga, G., Nyns, E. J., Verstraete, W. and Wentworth, R. L. (eds.), Proc. of the Second International Symposium on Anaerobic Digestion. Elsevier Biomedical, Amsterdam, The Netherlands (1981).

Publication Dates

• Publication in this collection
  06 Oct 1998
• Date of issue
  Dec 1997

History

• Accepted
  30 Oct 1997
• Received
  11 June 1997