Bioreactors for treatment of VOCs and odours – A review

Journal of Environmental Management 91 (2010) 1039–1054

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Review

Bioreactors for treatment of VOCs and odours – A review Sandeep Mudliar*, Balendu Giri, Kiran Padoley, Dewanand Satpute, Rashmi Dixit, Praveena Bhatt, Ram Pandey, Asha Juwarkar, Atul Vaidya Environmental Biotechnology Division, National Environmental Engineering Research Institute (NEERI), Nehru Marg, Nagpur 440 020, India

a r t i c l e i n f o

a b s t r a c t

Article history: Received 25 August 2008 Received in revised form 12 December 2009 Accepted 3 January 2010 Available online 23 February 2010

Volatile organic compounds (VOCs) and odorous compounds discharged into the environment create ecological and health hazards. In the recent past, biological waste air treatment processes using bioreactors have gained popularity in control of VOCs and odour, since they offer a cost effective and environment friendly alternative to conventional air pollution control technologies. This review provides an overview of the various bioreactors that are used in VOC and odour abatement, along with details on their configuration and design, mechanism of operation, insights into the microbial biodegradation process and future R&D needs in this area. Ó 2010 Elsevier Ltd. All rights reserved.

Keywords: VOC Odour Biological treatment Bioreactor Biofilter Membrane bioreactor

1. Introduction Volatile organic compounds (VOCs) and odorous compounds emitted from various industries pose problem to human and environmental health. With increasing population, and new residential and industrial developments, the demand for VOC and odour control systems to provide nuisance-free breathable air is increasing. Stringent environmental legislations enforced by government agencies, have led polluting industries to adopt effective air pollution treatment processes in order to comply with these regulations. As a consequence, biological treatment techniques for VOC and odor control have gained tremendous popularity in view of the several advantages they offer in comparison to traditional physical and chemical removal methods. Biological waste air treatment processes are not only cost effective as compared to conventional techniques such as incineration or adsorption but are also environment friendly (Devinny et al., 1999; Delhomenie and Heitz, 2005; Shareefdeen and Singh, 2005). Biological waste air treatment technology makes use of several types of bioreactors depending on the load and kind of pollutant to be treated. The type of bioreactor used for abatement has a direct consequence on the efficiency of the treatment process. An understanding of the bioreactors used for VOC and odour treatment, their

design and configuration, as well as necessary parameters for their operation will not only help in increasing the efficiency of the treatment process but also give insights to develop newer, better and robust treatment techniques. This review attempts to provide an overview of the various bioreactors used for the control of VOCs and odours, their merits and demerits, their important operational parameters and future R&D needs in this area. 2. Bioreactors in VOC and odour control Bioreactors play a very important role in the control of VOCs and odorous gases that are emitted by polluting industries. Although a number of different configurations exist, the main types of conventional air phase biological reactors include biofilters, biotrickling filters and bioscrubbers. Among the newly developed reactors are the membrane reactors (Shareefdeen and Singh, 2005; Kumar et al., 2008a,b), which have been used for VOC and odour abatement. Although, the basic pollutant removal mechanisms of all the reactors are more or less similar, differences exist in the use of microorganisms (may be either in suspended (in liquid) or immobilized (biofilm) form), packing media, pollutant concentration etc. 3. Biofilter

* Corresponding author. Tel.: þ91 712 2240097; fax: þ91 712 2249900/2249961. E-mail address: [email protected] (S. Mudliar). 0301-4797/$ – see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.jenvman.2010.01.006

Biofilters (BFs) are reactors in which a humid polluted air stream is passed through a porous packed bed on which a mixed culture of

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pollutant-degrading organisms is immobilized (Fig. 1). Generally, the pollutant gases flow through the porous packing media, are transported from the gaseous phase to the microbial biofilm (through liquid phase or moisture) and the biological oxidation of VOCs occurs. BFs are used to treat a wide variety of organic and inorganic pollutants in industrial and municipal exhaust streams. Although, traditionally used for treatment of odorous gases from sewage treatment plants and composting facilities, BFs now find wide application in treatment of several VOCs and odours. Among these are odourants such as ammonia, hydrogen sulphide, mercaptan, disulphides, etc., and VOCs like propane, butane, styrene, phenols, ethylene chloride, methanol, etc. Bench and pilot scale studies have shown that 60 out of 189 hazardous air pollutants (HAPs) can be successfully treated with biofiltration (Devinny et al., 1999; Shareefdeen and Singh, 2005). In Europe, more than 600 chemical processing industries use BFs for deodorization and treatment of VOCs. BFs are typically used for the treatment of large volumes of air streams containing low concentration of VOCs or odorants. The advantages and disadvantages of BFs are discussed below. Advantages: (a) (b) (c) (d)

Cost effective with (low operating and capital costs) Low pressure drop Treat large volumes of low concentration VOCs or odorants Secondary waste streams are not produced

Disadvantages: (a) (b) (c) (d)

Clogging of the medium due to particulate matter Problem of medium deterioration Less treatment efficiency at high concentrations of pollutants Difficulty in moisture and pH control

There are two BF configurations conventionally used for VOC and odour treatment viz.  The open design BFs, with ascending gas flows, installed outside the VOC/odour generating units. These reactors require large areas, and are also exposed to climate changes.

Fig. 1. Schematics of a biofilter unit.

 The close design BFs, with either ascending or descending gas flows, installed in closed rooms. These reactors require less space than the open configuration. 3.1. Biofilter operation The operation of BFs involve a series of steps beginning with the transfer of the pollutant from the air to the water phase, adsorption to the medium or absorption into the biofilm, and finally biodegradation of the VOC/odorant within the biofilm (Devinny et al., 1999). The most important physical, chemical and biological parameters influencing the biofiltration process are described below. 3.1.1. Transfer and partitioning of pollutant The first step in the biofiltration process is the transfer of contaminants from the air to the water phase. This is generally not a rate-limiting step, and so one frequently assumes that the gas and liquid are at equilibrium. At equilibrium, the partition between air and water is generally described by Henry’s law, which is given by the equation:

Cgi ¼ Hi Cli where Cgi is the concentration of pollutant i in the gas phase, Hi is Henry’s coefficient and Cli is the concentration of i in the liquid phase (Shareefdeen and Singh, 2005). Henrys’ coefficient (constant of proportionality Hi in above equation) has been described in different units in literature. Using a non-dimensional Henrys’ coefficient, substances with values over 0.01 are considered volatile, and the higher the value, the less soluble the substrate is in water. For example, the non-dimensional Henrys’ coefficient (at 25  C) for ammonia is reported to be 0.0005, while for H2S it is 0.92. Henrys’ coefficient depends on the temperature and the chemical potential in the liquid phase (Shareefdeen and Singh, 2005). In general, the elimination capacity of a BF declines with increasing Henry’s law constant since this indicates a tendency to partition away from the liquid/biofilm phase where degradation is taking place. 3.1.2. Biofilm The biofilm is a key element of the BF, which brings about the biodegradation of the pollutants viz. VOC and odorous compounds. ‘‘Biofilm’’ is the mass of organisms growing on the surface of the solid support; it carries out the catabolic activity and transforms the pollutants to harmless products. The thickness of the biofilm is influenced by several factors. These include the type of pollutant, its rate of flow through the BF, the bedding material used, and the design and configuration of the treatment system being used. Biofilm thickness usually varies from tens of micrometers to more than 1 cm, although an average of 1 mm or less is usually observed (Shareefdeen and Singh, 2005). The activity increases with the thickness of the biofilm, up to a level termed the ‘active thickness’. Above this level, the diffusion of nutrients becomes a limiting factor (Devinny et al., 1999). Various steady state and dynamic mathematical models have been reported in literature to predict the substrate, oxygen and nutrient penetration profile in the biofilm and facilitate evaluation of overall biofilm effectiveness factor (Mudliar et al., 2008a,b; Shareefdeen and Singh, 2005; Metris et al., 2001). 3.1.3. Biofilter bed The BF bed constitutes the heart of the biofiltration process because it provides the support for microbial growth. Bohn (1992) established a list of characteristics that an ideal BF bed should

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possess. The most important desirable characteristics of the BF bed include (a) high specific surface area for development of a microbial biofilm and gas-biofilm mass transfer, (b) high porosity to facilitate homogeneous distribution of gases, (c) a good water retention capacity to avoid bed drying, (d) presence and availability of intrinsic nutrients, and (e) presence of a dense and diverse indigenous microflora. Peat, soil, compost, and wood chips, are the most frequently employed basic materials in BF beds. These materials satisfy most of the required desirable criteria, and are widely available at low cost. Each of these materials has their own merits and demerits. The main advantage of soil is that, it offers a rich and varied microflora. It however, contains only a few intrinsic nutrients, presents low specific surface area and generates high-pressure drops (Swanson and Loehr, 1997). Peat has high amounts of organic matter, high specific surface area, and good water holding capacity and good permeability. However, peat contains neither high levels of mineral nutrients nor a dense indigenous microflora as in the case of soil or compost. Composts are materials that are most frequently employed in biofiltration for a variety of reasons. Compost offers a dense and varied microbial system, good water holding capacity, good air permeability, and contains large amounts of intrinsic nutrients. Moreover, the utilization of compost in BFs constitutes an effective way of recycling and utilizing waste residual organic matter, such as activated sludge from wastewater treatment plants, forest products (branches, leaves, barks), domestic residues, etc. (Alexander, 1999). However, composts are often less stable than soil or peat and have the tendency to break down and become compact, leading to increase in pressure drop in BF beds. This among other reasons is attributed to their high water holding capacity. Some authors have studied biofiltration using wood chips or barks as packing material (Smet et al., 1996a,b, 1999; Hong and Park, 2004). However, in general these authors have concluded that performances obtained with such filtering materials are less satisfactory than those obtained with compost or peat. This has been explained by the low pH-buffering capacity, the low specific surface areas and the low nutrient content of such materials. Despite these deficiencies, wood barks are still widely used in BFs as support materials, in association with peat or compost. Indeed, to prevent bed crushing and compaction, most authors suggest materials that provide the bed with good structure, easy maintenance and rigidity, which consequently delay the clogging phenomena which thereby increases the bed lifespan. Examples include wood chips or barks (Luo, 2001), perlite (Woertz et al., 2002), vermiculite (Pineda et al., 2000), glass beads (Zilli et al., 2000), polyurethane foam (Moe and Irvine, 2000), polystyrene (Arulneyam and Swaminathan, 2000), lava rock (Chitwood and Devinny, 2001), etc. Ibrahim et al. (2001) prepared a filter bed composed of activated sludge immobilized on gel beads. Christen et al. (2002) and Sene et al. (2002) developed a sugarcane-baggasebased bed, for the treatment of ethanol and benzene. Some bedstructuring agents also possess interesting chemical characteristics which they impart on the bed such as pH-buffering capacity (limestone), or general adsorbing capacity (activated carbon; Abumaizar et al., 1998). The efficiency of a BF material with respect to the pollutant for treatment is given by its adsorption coefficient or partition coefficient. Tang and Hwang (1997) reported partition coefficients of toluene as 1.43 mg g1 with compost, 2.00 mg g1 with diatomaceous earth, and 0.89 mg g1 with chaff. Beds containing activated carbon (granulated or powdered) provide adsorption coefficients for toluene approximately 10–20 times greater (50.6 mg g1 of granulated activated carbon) (Tang and Hwang, 1997; Acuna et al., 1999). Literature reports indicate that addition of activated carbon leads to improvement in biofilter degrading capacity (Abumaizar

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et al., 1998), elimination of hydrophobic compounds, and better control of loading variations (Mason et al., 2000). 3.1.4. Oxygen levels Oxygen levels play a very vital role in the performance of a BF. However, this is very case specific as can be seen from experiments discussed below, conducted by separate authors, on the effect of oxygen limitation on biofiltration. Experiments by Shareefdeen et al. (1997) using air enriched with oxygen improved the performance of the BF and demonstrated that oxygen was indeed a limiting factor. In another experiment, Deshusses et al. (1996) found that there was no significant improvement in the simultaneous removal of a mixture of methyl ethyl ketone (MEK) and methyl isobutyl ketone (MIBK), when the oxygen content in air was increased. In this case, the fact that significant cross-inhibition of MEK and MIBK biodegradation occurred suggested that kinetic effects were more important than diffusion effects. This was further demonstrated in transient experiments where spikes of either compound were injected into BFs, and both cross- and self-inhibitions were observed. Thus, role of oxygen in BF performance seems to be case specific. Oxygen is most likely to affect highperformance BFs or when thick biofilms exist. In general principal in most applications, BF operations seek to avoid anaerobic conditions. This is because, existence of even micro-anaerobic conditions lead to the formation of compounds, which themselves are odorous and this deviates from the overall goal of eliminating odorants and VOCs. Some studies, however suggest that, fortuitous anaerobic microenvironment conditions that exist in BFs, help in the degradation of organic pollutants (Shareefdeen et al., 1997). 3.1.5. Nutrients The pollutants introduced into the BFs, form the major carbon and energy source for microbial activity. Hydrogen and oxygen are found in the air, in the growth medium, and sometimes in the VOCs. The availability of the other macronutrients (N, P, K, and S) and micronutrients (vitamins, metals) is partially fulfilled by the filtering materials used in the BF. Materials such as composts are well known to contain various nutrients. Studies have demonstrated that irrespective of the filtering material employed, the steady addition of nutrients is necessary to sustain a satisfactory microbial degradation activity. For example, some studies have shown that long-term utilization of compostbased beds lead to progressive exhaustion of the intrinsic nutritive resources (Morgenroth et al., 1996). This progressive nutrient deficiency then becomes a limiting factor for the long-term biofiltration performance (Delhomenie et al., 2001a,b). Models of biofiltration performance as a function of nutrient supply, and of nitrogen in particular have been developed and experimentally validated (Delhomenie et al., 2001a,b; Alonso et al., 2001; Metris et al., 2001; Dorado et al., 2008). Nutrients for microbial growth are supplied either in the solid form which is directly inserted into the filter bed (Gribbins and Loehr, 1998), or as aqueous solutions, which is the most frequently used method. Wu et al. (1999) reviewed the most common nutrient solutions used in BFs. These include KH2PO4, NaxH(3x)PO4, KNO3, (NH4)2SO4, NH4Cl, NH4HCO3, CaCl2, MgSO4, MnSO4, FeSO4, Na2MoO4, and vitamins (B1, etc.). Given the wide range of elements and compounds influencing microbial behavior, the optimization of nutrient solutions for BFs is a challenging area of study. 3.1.6. pH As is the case with several biological processes, pH has an important influence on biofiltration efficiency. Above or below an optimum pH range, microbial activity is severely affected. Most of

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the microorganisms in BFs are neutrophilic i.e. their optimum pH is 7. Lu et al. (2002), observed maximum degradation of BTEX between pH values of 7.5 and 8.0. Lee et al. (2002) also reported a pH of 7.0 to be optimal for BTEX degradation. Veiga et al. (1999), studied the effect of pH on alkyl benzene degradation (between pH 3.5 and 7.0), and found that alkyl benzene degradation increased with pH. Arnold et al. (1997) stated that styrene elimination improved in a neutral medium. VOCs that contain hetero-atoms (S, O, and N) are converted into acidic products, which tend to reduce the pH (Christen et al., 2002), affect microorganisms and cause corrosion problems in downstream conduits (Webster and Devinny, 1998). Similar observations during VOC degradation due to formation of acidic intermediates have also been reported by various authors (Shareefdeen and Singh, 2005; Maestre et al., 2007). Kennes and Thalasso (1998) reported that among the organic materials employed in BFs, soil exhibited the best intrinsic pH-buffering capacity followed by composts and wood chips. Peats are naturally acidic (pH 3.0–4.0), and have low buffering capacity. To maintain the pH (at neutral) some authors have reported insertion of buffer materials into the filter beds, for e.g. – calcium carbonate (Smet et al., 1996a,b), and dolomite (Smet et al., 1999). The pH can also be controlled by bed irrigation with nutrient solutions that contain pH buffers, for example Ca (OH)2, NaOH (Zilli et al., 2000), NaHCO3 (Tang and Hwang, 1997), and urea (Delhomenie et al., 2002), etc. 3.1.7. Moisture content The moisture content of the filter bed is a critical factor for biofilter performance because microorganisms require water to carry out their normal metabolic activity. Sub-optimal moisture levels leads to drying of the bed and development of fissures that cause channeling and short-circuiting (Shareefdeen and Singh, 2005). Deprivation of water to microorganisms causes a significant reduction in the biodegradation rate. Excess water inhibits transfer of oxygen and hydrophobic pollutants to the biofilm, thereby promoting the development of anaerobic zones within the bed and limiting the reaction rate. Too much water can also result in foul smelling emissions due to the lack of oxygen, increasing backpressure due to reduced void volume, and channeling of the gas within the bed. Optimal water levels vary with different filtering material, depending on medium, surface area, porosity and other factors. Moisture content for optimal operation of the biological filter should be within 30–60% by weight, depending on the filtering medium used. Moisture levels in a biofilter are often maintained through pre- humidification of the inlet gas stream. Also, it is often necessary to provide direct application of water to the bed through a sprinkler system at the top of the bed. More advanced controls include the use of load cells that sense the weight of filter bed and are connected to sprinkler controls. Supplemental moisture supply may be required because bio-oxidation is an exothermic reaction, and so drying can occur within the bed. Drying of the packing material can lead to localize dry spots, and can result in nonuniform gas distribution and reduction in the activity of microorganisms (Shareefdeen and Singh, 2005). Recently, it has been reported that, biofilters tend to experience drying at the air inlet port, which causes decreased pollutant removal over time (Sakuma et al., 2009). Control of moisture requires a better understanding of the drying of the support due to changes in inlet air temperature and relative humidity and from production of metabolic heat during pollutant oxidation. Various models are now available to study drying and its effect on biofilter performance. These models describe the variations in pollutant concentration, air relative humidity, temperature, and water content of the media, to predict

water evaporation from the packing material as a consequence of metabolic heat generation and variations of the relative humidity of the inlet air stream, and also the resulting decrease in biofilter performance (Devinny and Hodge, 1995; Devinny et al., 1999; Metris et al., 2001; Morales et al., 2003). 3.1.8. Microorganisms Microorganisms are the catalysts for biodegradation of VOCs and odours. For the degradation of VOCs, heterotrophic microorganisms have been extensively reported (most often bacteria or fungi). The bed inoculation depends on both the nature of the filtering material and the biodegradability level of the VOC to be treated. Many scientific workers prefer taking advantage of the ecosystems already prevailing in the beds (Delhomenie et al., 2001a,b, 2002; Mohseni and Allen, 2000). After an acclimatization period, the most resistant population to the toxic VOC is naturally selected and a microbial hierarchy is established in the bed. In other cases (such as for recalcitrant VOCs), researchers inoculate the BF beds with consortia extracted from sewage sludge, or strains derived from either commercial sources or isolated from previously operated BF. In general, in terms of biomass density, a BF contains between 106 and 1010 cfu of bacteria and actinomycetes per gram of bed (Krailas et al., 2000). Pedersen and Arvin (1995), Pedersen et al. (1997), and Delhomenie et al. (2001a), have reported that in BFs, the degrading species represents between 1 and 15% of the total population. 3.1.9. Biofilm architecture VOC elimination is the result of many, interdependent processes that simultaneously take place inside the biofilter. To date, little information exists about biofilm architecture in BFs. Previous work with scanning confocal laser microscopy has revealed the existence of cell-free channels extending from the biofilm-liquid interface to the substratum and their possible role in enhancing pollutant and oxygen mass transfer (Cox and Deshusses, 1998). A new and promising development is the use of computed axial tomography (CAT) X-ray scanning to characterize the biofilm macro- architecture (Shareefdeen et al., 1997). CAT scans of a toluene-degrading BF containing a large amount of biomass immobilized on polypropylene pall rings showed a heterogeneous distribution of biomass with large areas completely filled with biomass whereas other sections of the reactor covered by <1 mm thick biomass. Further, image analysis revealed the presence of air/water channels ranging in area from <5 to380 mm2, with smaller channels (0– 60 mm2) contributing to more than 80% of the interfacial area. In future, further application of high resolution X-ray and possibly CAT scanning techniques could contribute to a better understanding of the architecture of biofilms. Such progress could lead to a better understanding of pollutant mass transfer in BFs and ultimately to a better design of materials for stable culture support (Shareefdeen et al., 1997). 3.2. Future needs Biofiltration has clearly been shown to be a cost and energy efficient technology for treatment of a range of waste emissions containing VOCs and odours. There is a need to work on innovative strategies such as pretreatment of VOCs and odours to remove particulates and/or enhance biodegradability and improve techniques to treat more complicated polluted airstreams especially multiple pollutant mixtures. Moreover, BF technology found field application well before its fundamental principles were understood. This has resulted in several cases of unsuccessful or suboptimum operation of large-scale BFs. Today, with much better understanding of the fundamental principles underlying the

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biofiltration process, scope exists for designing better bioreactors with optimal operating conditions. A number of fundamental questions however remain unanswered These include the quantification of biomass turnover, the understanding of biodegradation kinetic relationships and factors influencing these relationships, complex ecological complexity of BFs, and the interrelationships between pollutants, oxygen and essential nutrients (Cox and Deshusses, 1998). All these factors have been shown to significantly influence both the performance and the long-term stability of BFs, and thus require further investigation. In particular, quantitative studies are necessary. This would be made easier with the expanding use of modern tools of biotechnology. Further, there is need to develop design correlations on mass/heat transfer, diffusion coefficient in biofilm, gas/liquid holds up to facilitate improved biofilter design. Furthermore, the conventional phenomenological models are best with difficulties such as requirement of the detailed knowledge of the underlying physicochemical phenomena and extensive time required for their development, testing and validation. Hence, there is also a need to develop generic mathematical models of the biofiltration process using biological inspired computing techniques for quantitative robust prediction and design optimization (Narendra et al., 2006; Omkar et al., 2008). Finally, the largest problem to overcome will be the translation of recent and future basic advances into real process improvements.

4. Biotrickling filter The schematic description of a typical biotrickling filter (BTF) is provided in Fig. 2 (Delhomenie and Heitz, 2005). In such a filter, the gas is carried through a packed bed, which is continuously irrigated with an aqueous solution containing essential nutrients required by the biological system. Several studies have shown that the choice of a co- or counter-current configuration for liquid and gaseous phases does not influence the biodegradation performance (Cox and Deshusses, 1999). Microorganisms grow on the packing material of the biofilter as biofilm. The pollutant to be treated is initially absorbed by the aqueous film that surrounds the biofilm, and then the biodegradation takes place within the biofilm.

Fig. 2. Schematics of a biotrickling filter unit.

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The filtering material used in a BTF has to facilitate the gas and liquid flows through the bed, favor the development of the microflora, and should resist crushing and compaction. BTF packing that best meet these specifications are made from inert materials such as resins, ceramics, celite, polyurethane, foam (Centinkaya et al., 2000) etc. However, most of these materials present limited specific surface areas between 100 and 300 m1 (Roh, 2000) with exceptions in some cases where >1000 m1 for polyurethanebased beds have been reported. As they are made from inert or synthetic material, BTFs need to be inoculated with suitable microbial culture. The use of activated sludge as initial microbial inoculum has been extensively reported (Lu et al., 2002; Oh and Bartha, 1997). The advantages and limitations of BTFs include the following. Advantages: (a) Less operating and capital costs (b) Low pressure drop (c) Capability to treat acid degradation products of VOCs Disadvantages: (a) Accumulation of excess biomass in the filter bed (b) Complexity in construction and operation (c) Secondary waste streams Due to the permanent trickling mechanism, biofiltration processes are more adapted for the elimination of water soluble VOCs. Nevertheless, as the contact between microorganisms and the pollutants occur simultaneously (Cox and Deshusses, 1999), the solubility specifications are less stringent than for bioscrubbers (Henry coefficient <0.1; Van Groenestijn and Hesselink, 1993). In a typical BTF, VOC inlet concentrations are generally less than 0.5 g m3. The continuous distribution of the nutrient solution facilitates the control of the biological operating parameters (viz. pH etc.). As the contact between the microorganisms and the pollutants occurs after the VOC diffusion in the liquid film, the liquid flow rate and the recycling rate are recognized to be critical parameters for BTF operation. Research has suggested that an increase in the liquid flow rate should result in proportional increase in the active exchange surface for gas–liquid mass transfer, and then improve the degradation rate (Alonso et al., 2000). Some researchers have shown that maintaining minimum water and nutrient supply is sufficient to achieve good performance (Lu et al., 2002; Thalasso et al., 1996). In addition, as the distribution and the recycling of nutrient solutions add to energy costs, other studies suggest that the optimum recycling and distribution flow rates have to be found experimentally and on a case-by-case basis (Dolfing et al., 1993). The major drawback of BTFs is the accumulation of excess biomass in the filter bed. Some researchers have demonstrated that, in the course of the degradation process, the biofilm thickness can be several millimeters (Janni et al., 2001; Cohen, 2001) which can cause clogging, an increase in pressure drop, bed channeling, creation of anaerobic zones and can ultimately lead to performance loss (Alonso et al., 2001). Several studies have been attempted to develop solutions to the clogging problem in BTF. The control strategies suggested are of three types: mechanical, chemical or biological. Mechanical treatment includes bed stirring (Wubker et al., 1997; Laurenzis et al., 1998) or bed back washings with water (counter-current washings), which permit the draining of excess accumulated biomass (Smith et al., 1996). Chemical treatments include dissociation of the chemical binding between the biomass and the bed particle surface either by damage to the biomass by creating nutrient or water deficiency, or utilization of disinfecting

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agents (Diks et al., 1994; Schonduve et al., 1996; Cox and Deshusses, 1999; Armon et al., 2000; Chen and Stewart, 2000). Biological methods utilize biomass predators, such as protozoa (Cox and Deshusses, 1999). Amongst all of these methods, back washings with water are the most efficient and certainly cause less disruption to the biotrickling filter ecosystem and performance (Cai et al., 2004). 4.1. Operation and applications of biotrickling filters The important parameters of bioreactors in general (nutrients, pH, microorganisms, oxygen levels, etc.) have been described in detail for BFs in the preceding section. They will therefore not be dealt with individually for all reactors. BTFs find wide application in VOC and odour treatment. As compared to conventional compost or soil bed BFs which are generally limited to the elimination of odorous compounds and nonchlorinated volatile organic compounds, a wider range of pollutants can potentially be treated in BTFs. This is because, environmental conditions can be better controlled in the BTFs and potentially toxic dead-end metabolites can be purged out of the system. Also, laboratory BTFs offer the opportunity to work with monocultures, especially using genetically engineered microorganisms. Oh and Bartha (1997) first reported the elimination of nitrobenzene vapors in a laboratory scale BTF. They used a stable microbial consortium enriched from sewage sludge and immobilized it on perlite. During the start up period of four weeks, the inlet nitrobenzene concentration was kept relatively low (<80 mg m3) to avoid poisoning of the culture, after which high and sustained nitrobenzene elimination was observed with 80–90% degradation for inlet concentrations ranging from 100 to 300 mg m3 and an empty bed gas contact time of 21 seconds. This corresponds to an elimination capacity of 50 g m3 h1, a high value that could lead to an economically viable process. A nitrogen balance showed that 98% of the nitrobenzene nitrogen was converted into ammonia while a small amount of nitrite was produced. Two other compounds of interest, namely diethyl ether (Bauerle and Fischer, 1987) and gasoline additive methyl tert-butyl ether (MTBE) (Fortin and Deshusses, 1999) were reported to be biodegraded in laboratory BTFs. Fortin and Deshusses (1999) achieved 75% removal efficiency for an inlet MTBE concentration of 0.8 g m3 with empty bed residence time less than a minute in BTF. This corresponds to an elimination capacity of 50 g m3h1, an extremely high value for a compound for which biodegradation in situ still remains a challenge. The reactor studied by Fortin and Deshusses (1999) was originally inoculated with various samples of aquifer material and soil contaminated with MTBE. Interestingly, MTBE removal was significant only after addition of traces of a peat humic substance (PHS) extract to the recycle liquid. As biomass accumulated in the reactors, the benefits of the PHS were no longer significant. While several reports exist on bio-stimulation using PHS in wastewater treatment, the exact mechanisms involved in bio-stimulation using PHS are yet to be elucidated. Also noteworthy, is a study by Sun and Wood (1997), who immobilized a pure culture of Burkholderia cepacia PR123 (TOM23C) constitutively expressing toluene ortho-monooxygenase to cometabolize the biodegradation of trichloroethylene (TCE) vapors in a BTF. Aerobic biodegradation of TCE only occurs through cometabolism, and addition of a growth substrate (usually toluene, methane, propane, phenol, or ammonia), which is required to induce the expression of the appropriate TCE-degrading enzyme. Bacterium B. cepacia PR123, however, expresses toluene orthomonooxygenase constitutively, which circumvents the problem of competitive inhibition of TCE oxidation by the usual inducers during the growth phase. The authors Sun and Wood (1997) used

glucose as a carbon and energy source and observed TCE eliminations up to 200 times higher than previously reported. As observed in other bioreactors for TCE aerobic co-metabolism, rapid inactivation of the TCE-degrading enzyme by TCE breakdown products (e.g. TCE epoxide) remained a problem. Biotrickling filtration is a maturing technology, and the number of full-scale BTFs is rapidly increasing. In the past few 2–3 years, several successful conversions of full-scale chemical scrubbers to biotrickling filters have been demonstrated (Kraakman, 2001; Kraakman, 2003; Gabriel and Deshusses, 2003). Until recently, it was thought that successful bio-treatment in BTFs required a gas contact time ranging from 10 to 30 s (Wu et al., 2001). However, some researchers have shown that in BTFs the residence time can be reduced to 5 s (Shareefdeen and Singh, 2005). Recently, an alkaline biotrickling filter was shown to be very effective for treatment of H2S odours (Sanchez et al., 2008). 4.2. Future needs Recent research in the field of biotrickling filtration for air pollution control have focused on various aspects pertaining to the microbiology of pollutant-degrading microorganisms, kinetics of pollutant uptake, and means to control biomass accumulation. Nevertheless, additional information on the fundamental principles underlying biotrickling filtration is needed. Key questions to be addressed are mainly concerned with the complex ecology of biofilms. In particular, studies are needed to understand the overall role of secondary processes (i.e. those processes not directly associated with the elimination of the primary pollutant) and how these can be controlled in practice. In the future, the ability to control the ecology of biofilms in BTFs may enable optimal balancing of the net growth of biomass, so that reactor stability can be ensured over a very long period. Additional research is needed for better understanding of the kinetic relationships for pollutant biodegradation. Particularly, understanding of the biodegradation of mixtures of pollutants, role and impact of oxygen and ancillary nutrients on the rate of biodegradation and on the biomass yield, and to determine the influences of various stresses, such as changing conditions and mass transfer limitations, is important. These studies are extremely relevant for future implementation of BTFs in actual field conditions (Cox and Deshusses, 1999). In a nutshell, review of the recent research work emphasizes on the need of fundamental understanding of the degradation process through in situ analysis and extended application of modern tools in biotechnology. This is essential in order to establish baseline information (presently not available) for rational reactor design and optimum process operation. This, together with number of pilot scale application and demonstration of techno-economic viability, would transfer this technology from lab to the field. 5. Bioscrubber A bioscrubber unit (Fig. 3, Delhomenie and Heitz, 2005) consists of two subunits namely (1) an absorption unit and (2) a bioreactor unit. In the absorption unit, input gaseous contaminants are transferred to the liquid phase. Gas and liquid phases flow countercurrently within the column, which may contain the packing material. Nevertheless, the addition of inert packing provides increased transfer surface between the VOC and the aqueous phase (Van Groenestijn and Hesselink, 1993). The washed gaseous phase is released at the top of the column whereas the separated contaminated liquid phase is pumped to an agitated, aerated bioreactor. This reactor unit contains the appropriate microbial strains suspended in the aqueous phase in nutrient solution (media) essential for their growth and maintenance.

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column and any biodegradation in the absorber lowers the pollutant concentration in the liquid, which increases the concentration gradient and gas transfer. Biodegradation in the bioreactor unit continues to drive the pollutant concentration to low levels. The carbon is regenerated as the VOCs desorb and degrade, and is then recycled to the absorber. PAC also provides a surface for the growth of biofilms, and may assist in treating peak loads and in decreasing inhibition by toxic compounds. Addition of 2–5% PAC to biomass slurry is recommended to improve bioscrubber performance (Hammervold et al., 2000). 5.1.2. Anoxic bioscrubber Aerobic bioreactors are used to transform inorganic pollutants and degrade organic pollutants in conventional bioscrubbers, whereas systems combining scrubbers and up flow anaerobic sludge blanket (UASB) bioreactors have been used for the degradation of perchloroethylene and also to treat waste gases containing NOX and SOX (Centinkaya et al., 2000; Janssen et al., 2000; Shareefdeen and Singh, 2005). Fig. 3. Schematics of a bioscrubber unit.

Most of the bioscrubbers being operated presently use activated sludge derived from wastewater treatment plants as inoculum (Ottengraf, 1987). In some cases, bioreactors are directly inoculated with specific degrading strains. The residence time for such bioreactors range between 20 and 40 days and these are operated practically as activated sludge processes including recycle of sludge. Part of the treated solution is recycled for absorption of VOCs to the absorption unit. The advantages of bioscrubbers are as follows. (a) Operational stability and better control of operating parameters (pH, nutrients); (b) Relatively lower pressure drops (c) Relatively smaller space requirement. Disadvantages of bioscrubbers include(a) Bioscrubbers are adapted to treat readily soluble VOCs (alcohols, ketones), with low Henry coefficients (<0.01), and at concentrations less than 5 g m3 in the gaseous phase (b) Provides low specific surface area for gas–liquid mass transfer (generally <300 m1) (c) Excess sludge generation (d) Generation of liquid waste Some studies have shown that the addition of emulsifying agents (silicon oil, phthalate) in the aqueous solution can significantly improve the elimination of less soluble compounds, because they favor the VOC mass transfer from gas to the liquid phase (Mortgat, 2001). 5.1. Variations in bioscrubber design Substantial modifications in bioscrubber design have been done in the recent past to enhance their performance for VOC and odour treatment. Some modified bioscrubber units are listed below. 5.1.1. Sorptive-slurry bioscrubber The sorptive-slurry bioscrubber consists of a suspended growth bioscrubber with powdered activated carbon (PAC) added to the biomass slurry (Kok, 1992). Gaseous VOCs partition into the slurry in the absorber unit and adsorb onto the carbon. Adsorption on the

5.1.3. Two-liquid phase bioscrubber The concept of two-liquid phase bioscrubber unit originates from the fact that application of conventional bioscrubbers becomes limited for the treatment of pollutants that are readily soluble in water. Consequently, the addition of an organic solvent to the water phase can enhance biodegradation of more hydrophobic compounds (Deziel et al., 1999), and facilitate the elimination of a range of hydrophilic and hydrophobic compounds. Besides improving bioavailability, solvents can reduce the toxicity of contaminants, and can act as a buffer system for fluctuating loads of pollutants. The addition of 10–30% water immiscible high boilingpoint solvent to the liquid phase facilitates the absorption of hydrophobic compounds from the gas phase in the absorber. Twoliquid phase bioreactor systems have been tested for the removal of alkanes, benzene, styrene, phenol, naphthalene, and pentachlorophenol. Solvents such as silicon oil, paraffin oil, dibutyl phthalate, di-n-octyl phthalate, di-n-nonylphthalate, and pristine are good candidates for this application, and among them silicon oil has been found to be the best for two-liquid phase systems (Yeom and Daugulis, 2001). 5.1.4. Airlift bioscrubber Edwards and Nirmalakhandan (1999), have described an airlift bioscrubber having a combined absorption/biodegradation reactor configuration, for the removal of air phase benzene, toluene and xylene (BTEX) compounds. The reactor comprises of two concentric tubes, with the inner tube shorter than the outer tube. As in some conventional airlift bioreactors, the inner tube serves as the down comer, and the annular space between the two tubes works as a riser (Ward, 1989). Air is introduced into the reactor through the sparger, which is located near the bottom of the riser. The combined medium of mixture of air and water in the riser has lower density than the water in the down comer, resulting in a fluid circulation within the reactor. A mathematical model developed for this process indicated that removal rates of >99% can be achieved for benzene and toluene in the air stream with concentrations <1000 ppmv, and that the airlift bioscrubber should be operated at biomass concentrations of 2 g 11 or greater for operational stability. Recently, Jianping et al. (2005) reported the simultaneous removal of ethyl acetate and ethanol in air streams using a gas– liquid-solid three-phase flow airlift bioreactor. 5.1.5. Spray column bioscrubber Bioscrubbers have been suitably modified as per the demand and nature of the odorants to be treated. A modified bioscrubber for

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treatment of large volumes of waste gas and industrial emissions with low concentration of odorants with poor solubility was investigated by Cesario et al. (1992). The modified system consists of two bioreactor units, which include (a) a spray column bioreactor with liquid impelled loop reactor (LILR) and (b) a spray column bioreactor connected to an airlift loop reactor (AIR). In both the systems, gaseous odorants/liquid contactor is a spray column for transfer of the odorants from gas to the water phase. This is interconnected to either liquid impelled loop reactor or airlift loop reactor. Both the attached systems (LILR and ALR) contain culture medium and the biomass. The spray column and LILR/ALR are interconnected having a provision for recycling of the contents of the reactor (Cesario et al., 1992). The spray column reactor with LILR contains nutrient medium. It is necessary that the nutrient medium in the reactor must possess characteristics such as immiscibility with water, non-biodegradability, should be non-toxic to the biocatalyst, should have low vapor pressure, relatively low viscosity, and a density different from the density of water. The water immiscible liquid is recycled between the absorber and the bioreactor. However, the operation (recycling) depends upon (a) the volume of industrial emission to be treated, (b) physicochemical characteristics and (c) the concentration of the odorants in the gaseous emission and also (d) the targeted removal efficiency. The novel reactor of a spray column system with LILR/ALR is still on laboratory scale and requires extensive investigations to ascertain the odorant transfer from gas to the liquid absorber (ALR). 5.2. Future needs Bioscrubber are generally considered to be useful for the treatment of waste gases containing water soluble pollutants (Henry’s law coefficient, H < 0.01), but depending on the concentration and type of pollutant, bioscrubbers may also efficiently be used for odorous gases with other characteristics. Bioscrubbers, provide substantial advantages for waste gas treatment because of their smaller space requirements, high loading rates, reliable operation, process control, low risk of clogging, and low operating cost. When high concentrations of contaminants are to be treated, bioscrubbers offer more advantages than conventional BFs, BTFs, and chemical scrubbers. Combination of bioscrubbers and some polishing steps may improve the treatment efficiency for gases with mixture of hydrophobic and hydrophilic compounds. Bioscrubbers offer higher elimination efficiencies for gases generated from liquid wastes, viz. H2S, NH3, and organic sulfur compounds. However, due to the acidifying nature of these substances, substantial oxidation and sulfuric acid production in the scrubber may cause the pH to drop and decrease mass transfer efficiency. The capacity of the adsorption section for handling higher H2S concentration needs to be further improved. This could be achieved by increasing the buffering capacity of the scrubbing medium and also through pH control. The bioscrubber application can be made more attractive than the use of other conventional BF or BTF at relatively high pollutant concentration (>0.5 g m3) with necessary modifications. There is also a need to develop integrated/hybrid reactor configurations to achieve optimal and efficient bio-treatment of VOCs/odours, individually and in mixtures.

Application of membrane bioreactors for waste gas treatment has been reviewed before (Reij et al., 1998; Kumar et al., 2008a,b). The concentration difference between the gas phase and the biofilm phase provides the driving force for diffusion across the membrane. A pressure difference is not applied. The driving force depends strongly on the air–water partition coefficient of the diffusing volatile component. For components with a high partition coefficient the driving force for mass transfer is small. The concentration in the liquid, which depends on the biodegrading activity of the microbial population, also affects the driving force. The surface of the membrane forms the contact area (Reij et al., 1998). In a BF bioreactor unit, waste gas is blown through a bed of compost or soil, where microorganisms consume or degrade the gaseous organic pollutants. No separate water phase is present. An advantage of the membrane bioreactor over the BF is the presence of a discrete water phase allowing optimal humidification of the biomass and removal of the degradation products, thus avoiding inactivation of the biomass. In a membrane bioreactor, the membrane serves as the interface between the gas phase and the liquid phase (Fig. 4, Reij et al., 1998). The gas–liquid interface thus created (e.g. in hollow fibre reactors) is larger than in other types of gas–liquid contactors (Yang and Cussler, 1986). Moreover, in biotrickling filter and the bioscrubber, a packed bed of inert material is present on which water is continuously sprayed. The pollutants in these reactors have to diffuse through the water phase before it is consumed by the microorganisms. For pollutants with poor watersolubility such a layer of water causes a substantial additional resistance for mass transfer (De Heijder et al., 1994). In the membrane bioreactor, on the contrary, the liquid phase is situated at the opposite side of the biofilm and hardly forms a barrier for mass transfer of the poorly water soluble pollutants (Fig. 4). As mentioned before, large gas–liquid interfaces of 1000– 10,000 m2m3 can be created in hollow fibre reactors (Rautenbach and Albrecht, 1989), allowing high mass transfer rates. The pressure drop in the gas phase is much lower than observed in BFs, where pressure drop may become significant. Advantages of membrane bioreactors include. (a) No moving parts (b) Process easy to scale up (c) Flow of gas and liquid can be varied independently, without the problems of flooding, loading, or foaming Disadvantages of the bioreactor are(a) High construction costs (b) Long-term operational stability (needs investigation)

6. Membrane bioreactors Membrane bioreactors were designed as an alternative to conventional bioreactors for waste gas treatment. The membrane bioreactor allows the selective permeation of the pollutant, which is not allowed in any of the reactors discussed previously.

Fig. 4. Schematics of a membrane bioreactor containing microporous hydrophobic membrane, a biofilm and suspended cells.

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(c) Possible clogging of the liquid channels due the formation of excess biomass

6.1. Membrane materials Two types of membrane materials have been used to prevent mixing of the gas and liquid phases and simultaneous transfer of volatile components. These types are hydrophobic micro porous membrane and dense membrane. 6.1.1. Micro porous membranes A micro porous membrane has a highly porous structure, with a typical commercial membrane containing 30–85% pore space (Hartmans et al., 1992), including open pores in the surface of the membrane. The surface pores are generally sub micrometer in size, which prevents organisms from passing through the pores. If the pore size distribution is not sufficiently controlled or pore size is too large, then intrusion by organisms and organics will occur, significantly reducing mass transfer and potentially plugging the gas phase (Hartmans et al., 1992; Attaway et al., 2002; Fitch et al., 2003). Hydrophobic micro porous membranes consist of a polymer matrix of polypropylene or Teflon and contain pores with a diameter in the range of 0.01–1.0 mm. Since the membrane material is hydrophobic, the pores are filled with gas. Water does not enter the pores, unless a certain critical pressure at the liquid side is exceeded. If the transferred component disappears by chemical reaction, its mass transfer rate increases (Prasad and Sirkar, 1992). Active microorganisms present in the liquid will thus enhance it significantly. If the microorganisms are present as a biofilm on the membrane, liquid flow close to the membrane is absent and as a consequence the mass transfer coefficient does not apply at all. In this case (biofilm), simultaneous diffusion and reaction in a stagnant layer need to be calculated (Reij et al., 1995). 6.1.2. Dense membranes A dense-phase membrane has no pores for removal to occur. The contaminant must dissolve in to the membrane and diffuse through the membrane. In case of transport through a dense membrane, the diffusing volatile component is absorbed in the membrane material and diffusion takes place in the dense polymer (Reij et al., 1998). The mass transfer coefficient inside a dense membrane depends on both the solubility and the diffusivity of the volatile component in the dense matrix (Crank and Park, 1968). Interaction between the various gaseous components is assumed to be absent, i.e. independent diffusion. For each volatile component, the solubility and diffusivity are different and the mass transfer resistances of dense membranes for various gases may differ considerably due to specific interactions between the components in the gas phase and the membrane material. As a consequence, components can be selectively extracted from or retained in the gas phase by a proper choice of the membrane material (Reij et al., 1998). 6.2. Gas–liquid contactors for membrane bioreactors Both micro porous and dense membranes have been used for a variety of processes that involve gas–liquid contact. Micro porous material is generally applied in hollow fibres, although spiralwound and plate-and-frame modules have also been used (Sirkar, 1992; Wickramasinghe et al., 1992). Micro porous membranes can be applied as gas–liquid contactors when selective action of the membrane is not required. Volatile components diffuse through this material depending on their diffusion coefficient in air and their vapor pressure, while the membrane serves as gas–liquid contact area.

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Dense material is available as tubes (usually silicone tubing), with a wall-thickness of at least several hundred micrometers, and as composite membranes. Composite membranes consist of a thin, selective top layer (<1–30 mm) of dense material, supported by a highly porous support layer of, for, e.g. non-woven polyester or a microfiltration membrane. Composite membranes can be applied in spiral-wound, plate-and-frame and in hollow fibre modules. Examples of the application of both micro porous and dense membranes as gas–liquid contactors are given in the following section (Reij et al., 1998). 6.3. Applications of membrane bioreactors in waste gas treatment In addition to the relatively new process of membrane-based gas absorption (Sirkar, 1992), membrane contactors have recently been tested for biological treatment of gas streams. In such a process, the pollutants diffuse through the membrane and are degraded by the microbial population present in the liquid phase. In this context, various publications were critically reviewed and the summarized results are presented in Table 1. In general, the biomass is supplied with carbon and oxygen from the gas phase, while water and mineral nutrients are supplied through the liquid phase. Microorganisms grow as a biofilm on the membrane, but may also be suspended in the liquid phase. Most studies on membrane bioreactors concern the removal of hydrophobic pollutants from air. Hydrophobic pollutants, like xylene, toluene, hexane, and propene, have a high air–water partition coefficient. The driving force for the transfer of these pollutants to the water phase is very small and as a consequence, mass transfer limits the biodegradation and therefore the design of the bioreactor becomes critical (Dingemansa et al., 2008a,b; Witte et al., 2009). The large gas–liquid interface and excellent mass transfer properties of membrane reactors (Yang and Cussler, 1986; Karoor and Sirkar, 1993) have inspired several workers to test membrane bioreactors for the removal of less water soluble pollutants from air (Reij et al., 1995; Bauerle et al., 1986, 1987). A dual tube dense-phase silicone membrane bioreactor was investigated for control of cyclohexane-contaminated air as part of a jet propulsion fuel remediation investigation strategy (Roberts, 2006). Mass transfer characteristics for VOC permeation through flat sheet porous and composite membranes showed that the contribution of the porous ‘‘backing’’ layer for mechanical support can be substantial in comparison to the porous layer in contact with the dense layer (Dingemans et al., 2008). Removal in the bioreactor ranged from 29.4 to 596.6 mg m2 m1 in and measured elimination capacities ranged from 46.7 to 947.9 g m3 h1. The membrane materials used in several studies were chosen such that they were impermeable to microorganisms (Hartmans et al., 1992; Freitas dos Santos et al., 1995). As a consequence, these organisms could not contaminate the gas phase. This precaution was considered to be important in case the membrane bioreactor was applied for the treatment of indoor air or manned space cabin (Binot et al., 1994). Freitas dos Santos et al. (1995) tested a reactor with silicone tubes to remove 1,2-dichloroethane from air. For the destruction of trichloroethene (TCE), Parvatiyar et al. (1996a,b) designed a new membrane bioreactor in which both an aerobic and an anaerobic region were present (Fig. 5). In the anaerobic zone, TCE is partially dechlorinated and the products are supposedly degraded further in the aerobic zone of the biofilm. The silicone membranes due to their selectivity for hydrophobic components, retains acid vapors (SO2) that hamper biodegradation of 1,2-dichloroethane (Freitas dos Santos et al., 1995). Dense membranes may also serve as a buffer, in case the supply of pollutants is variable. It should be noted, however, that due to

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Table 1 Application of membrane bioreactor for VOC and odour treatment. Contaminant (g m3)

Membrane

Load (g m3 h1)

EC (g m3 h1)

Operation (days)

Gas HRT (s)

Flux (mg m2 h1)

References

Xylol (0.1–0.6) Butanol (0.1–0.5) DCM (0.2–0.4) Toluene (0.38)

HF silicone rubber

1.8–7.9 1.3–5.4 1–4 33–1500

n.r.a

2.5-14

w90

0.9–6.3

67–302 48–207 37–151 13–180

Bauerle et al. (1986) Weckhuysen et al. (1993)

HF polyporous pp

1.8-8.1 1.8–7.3 2.4-11 2-31

Toluene (0.7–3.4) Nitrogen oxide (0.1) DCM (0.65)

HF polyporous PP HF polyporous PP Spiral-wound silicone rubber

23-121 1.3 14–35

16-42 0.1–0.9 12–28

> 20 105 11

0.9-1.8 1.9 n.r.a

15–35 3–8 25–55

Methane, TCE in liquid (264) Dimethyl sulfide (0.036–1.81) Toluene (0.75–1.5)

Silicone rubber tubing

n.a.b

12

w120

n.a.b

130

Flat composite silicone rubber on polysulfone HF polyporous polysulfone, two in series Composite with silicone rubber

5–270

5–200

79

8–24

Methanol (0.01–2.6) Toluene (0.03–4.2) Hexane (0.03–2.4) Xylol (0.1–0.6) BBTEX (7.7–15.4) BTEX (2.2–9.8) DCM (0.16) Toluene (0.075) Toluene (0.004–3.18) Toluene (0–0.97)

Composite with silicone rubber HF polyporous PP Flat polyporous PP Flat composite, Silicone rubber on PVDF Flat composite, silicone rubber on PVDF Flat composite, silicone rubber On PVDF

n.r.

a

n.r.

>140

Attaway et al. (2002) Hartmans et al. (1992) De Bo et al. (2002)

339 20

2–24 24

312 320 117 790 n.r.a n.r.a

De Bo et al. (2002)

20–144

4–24

>8 105 21 100 40 40 270

1.8–7.2 3.6–7.2 1.6 and 2.9 96–300 0.17–1.4 0.55–1.4 – 24

6–60 0–5 13–326 0.06–0.2 4.2–20.4 670–2700 29–597 128

165 46.7

2–24 40

– 1764–35,760

Dolasa and Ergas (2000) Dolasa and Ergas (2000) Fitch and England (2002) Pressmann et al. (2000) Fitch et al. (2003) Fitch and England (2002) Robert et al. (2006) Munkhtsetseg et al. (2008) Kumar et al. (2008a,b) Robert et al. (2006)

105 80 29 30–395 0–170

6 5–170

0–6 5–130 32–72 0–6 3.8–13 0.01–0.06 4.8–58 2.5–18 47–947 54 0.6 947

Composite porous PAN Dual tube silicone rubber

0.72 46.7

80

8–16 1.6–9.6

118 180 84 32–470 0–170

Toluene Cyclohexane

a

20 <1

n.r. – – 52

HF polyporous polysulfone HF polyporous PP HF polyporous PP Latex rubber tubing Dual tube silicone rubber Thermophilic membrane

16 or 32

420 670 400 170

25 40 24 360

16–96 3–9 13–26 0.03–0.1 7–60 7-28 395–2189 64

a

De Bo et al. (2002) (De Bo, 2003) Parvatiyar et al. (1996a,b) Resier et al. (1994)

n.r. – – 4.3–15

36 58 33 600

TCE (0.04) Dimethy1 sulfide (0.011–1.63) Toluene (0.5–0.75) TCE (0.04–0.2) Butanol (0.6–2.3) TCE (0.13–0.21) Benzene (0.1) Benzene (0.1) Cyclohexane DMS

HF polyporous PP

a

Ergas and McGrath (1997) Ergas et al. (1999) Min et al. (2002) Freitas dos Santos et al. (1995) Clapp et al. (1999)

Attaway et al. (2001)

De Bo et al. (2002)

Configurations: HF: hollow fibre (i.d. < 0.5 mm); C: capillary (0.5 mm < i.d. < 10 mm); PP: polypropylene; PDMS: polydimethylsiloxane; PVDF: polyvinylidenefluoride; Zrf: zirfon. Compounds: MeOH: methanol; BuOH: 1-butanol; NH3: ammonia; BENZ: benzene; TCE: trichloroethylene; TOL: toluene; PROP: propylene; NO: nitric oxide; HEX: hexane; DMS: dimethylsulfide; BTEX: mixture of benzene, toluene, ethylbenzene and xylenes; DMS: dimethylsulfide; DCM: dichloromethane; DCE: dichloroethane. a n.r., Not reported or not sufficient data to calculate. b n.a., Not applicable (e.g., static gas phase).

simple thermodynamics, the equilibrium concentration in the water phase would never change upon the insertion of any type of membrane between the gas phase and the water phase. Irrespective of the membrane resistance, the driving force for mass transfer depends on the concentration to which the pollutant is reduced in the liquid phase. Therefore, the removal rate in a membrane bioreactor depends largely on the activity of the microbial population. In most of the reported studies, biofilm formation was observed to be an essential part of reactor operation (Ergas et al., 1999). Both mixed cultures and pure cultures were used for biofilm generation. The hydrophobic nature of both micro porous and silicone membranes facilitates microbial adhesion. The microorganisms located close to the membrane are exposed to higher substrate concentrations than suspended cells, making it more likely that most cell growth occurs close to the membrane. Biofilm growth may cause serious problems if excess biomass is not sloughed off. Freitas dos Santos et al. (1995) attributed the decreasing reactor performance and the increasing pressure drop over the liquid phase to extensive biofilm formation in the spiralwound membrane module they studied. Clogging of hollow fibres with a biofilm of propene-degrading Xanthobacter cells could be prevented by applying a very high liquid velocity, but still the

reactor performance decreased over a period of 3–6 months as the biofilm matured. These results suggest that, even if clogging is prevented, biofilms are prone to aging (Reij and Hartmans, 1996). During the degradation of dichloromethane (Bauerle et al., 1986) a good biofilm did not develop on the reactor membrane. The biofilm growing on dichloromethane sheared off the membrane after 94 days, causing a drop in reactor performance (Reij et al., 1995). Recently investigation on DMS removal in thermophilic membrane bioreactor indicated an elimination capacity of 54 g m3 h1 with the removal efficiency 84% at gas retention time (GRT) 24 s (Munkhtsetseg et al., 2008). The reason suggested for this was that hydrochloric acid was produced during the degradation of dichloromethane, which accumulated in the biofilm to toxic levels and destabilized the biofilm. Aziz et al. (1995) purposely repressed biofilm formation in a membrane reactor for wastewater treatment by the addition of a sequestering agent. This membrane reactor was part of a twostage bioreactor, in which methanotrophs were circulated. In the membrane reactor, the methanotrophs degraded trichloroethylene (TCE) and in a separate reactor, growth substrate was supplied. Separate reactors were required for the study, since TCE itself did not support microbial growth and could be degraded only co-

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Fig. 5. Aerobic and anaerobic biodegradation of trichloroethene (TCE) in membrane bioreactor.

metabolically when a primary substrate was supplied. Such a twostage process may not only be used for wastewater treatment but also for the removal of TCE from waste gas. A completely new strategy for the removal of TCE from air was designed by Parvatiyar et al. (1996a,b). In their newly designed membrane bioreactor, acetate was added to the liquid phase as carbon source and as electron donor to lower the oxygen tension in the biofilm. Under anaerobic conditions created in this way, TCE was partially dechlorinated. Subsequently, the products of the anaerobic dechlorination were degraded further in the aerobic zone of the biofilm (Fig. 5, Cesario et al., 1992). Their work, however, did not contain experimental evidence that both the aerobic and the anaerobic zone are present, but it is the first report on the continuous removal of TCE from air in the absence of volatile growth substrates. While reviewing literature (Table 1), it was found that the experiments reported are very varied and the approach in each case is different. For example-in most studies oxygen was made to diffuse through the membrane along with the compound to be removed from air, but Freitas dos Santos et al. (1995) supplied oxygen in the water phase. Parvatiyar et al. (1996a,b), on the other hand, maintained the liquid phase anaerobic, to allow anaerobic degradation. Secondly, the time period in each experiment varied, while some lasted for more than a year, some were completed within a few days. 6.4. Future needs All studies carried out on membrane reactors are laboratory scale experiments. To the best of our knowledge, no reports are available on pilot-plant investigations or full-scale applications of membrane reactors in biological waste gas treatment. Membrane modules appear relatively easy to scale up given their modular nature (Karoor and Sirkar, 1993); however an extensive long-term performance testing is necessary before they can be applied on full-scale. The effect of biomass on the membrane material, in the long run has not sufficiently been tested. During prolonged operation microbial polysaccharides might get absorbed to the membrane material, decreasing the critical pressure, and allowing liquid to penetrate the pores of hydrophobic micro porous membrane. This wetting of the membrane may significantly increase membrane resistance (De Heijder et al., 1994) and will be a bottleneck for longterm operation of reactors with micro porous membranes. No experimental evidence on this subject has been reported so far. Wetting of the membrane could be prevented by a thin coating of dense material applied on the liquid side of a porous membrane. Such composite membranes have been used in blood oxygenation to suppress blood-trauma and prevent the pores from filling with liquid and cell debris (Sirkar, 1992).

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In addition to the durability of the membrane material, the stability of the biomass is essential as well. The formation of thick biofilms (Freitas dos Santos et al., 1995) and clogging of the liquid channels (Kreulen et al., 1993) were shown to deteriorate reactor performance. Even when clogging was prevented by a very fast liquid flow, the performance of hollow fibre modules decreased with time (Kreulen et al., 1993). Therefore, strategies have to be developed to monitor the biofilm, to stabilize its activity, and to remove excess biomass from the membrane modules. The removal of poorly water soluble pollutants from air can be considered to be the most promising application for membrane bioreactors. The mass transfer resistance of membranes for this group of pollutants is negligible. Recently, Kumar et al. (2009), have reported enhance performance of a composite membrane bioreactor for treating toluene vapors. The composite membrane consisting of a porous poly acrylonitrile (PAN) support layer coated with a very thin (0.3 microm) dense polydimethylsiloxane (PDMS) top layer indicated maximum elimination capacity of 609 g m3 h1 along with the flux 1.2 g m3 h1. Moreover, the large gas–liquid interface of membrane modules enables efficient removal of these pollutants that in general are difficult to remove from air. Other niches for the application of membrane bioreactors are indoor applications and the removal of pollutants that require a specific microbial population (like TCE and nitrogen monoxide). Recently, the membrane bioreactor with both an anaerobic and an aerobic zone was proposed (De Heijder et al., 1994). Such a bioreactor might enable the biodegradation of pollutants, such as highly chlorinated hydrocarbons, that until now were considered to be beyond the reach of aerobic biological waste gas treatment. 7. Comparison of conventional bioreactors and membrane bioreactor for VOC and odour control The application areas and comparative performance evaluation of bioreactors (biofilter, biotrickling filter, bioscrubber, membrane bioreactor) widely reported for VOC and odour control is presented in Tables 2a and 2b, respectively. The target pollutant concentration for bioscrubber is relatively higher than biofilter and biotrickling filter, while in case of membrane bioreactor the membrane flux limit should allow for higher VOC and odour concentration. Biofilters usually have a media layer of 1–2 m to prevent excessive air velocities through the media, which easily results in high-pressure drop or airflow preferences. Biotrickling filters and bioscrubbers do not require an upfront humidifier, like biofilter do, to increase the inlet air humidity up to a preferable 100%. The footprint of biotrickling and bioscrubber reactors is normally much smaller, since they usually contain a more open packing that can be more than 2 m high. The media in biofilters needs to be replaced frequently, due to the deterioration of the (usually organic) media, or the increasingly worse process conditions like pH decrease, nutrient depletion, or extensive biomass accumulation, salt content or pressure drop. In biotrickling and bioscrubber reactors, no medium change-out is required, since inert packing material is used and process conditions can be better controlled. Operational costs can be saved with biotrickling or bioscrubber reactors, since up to 40% of the operational cost of a biofilter is typically related to the medium change-out. Further, the biofilter reactor do not have a continuous and distinct liquid phase as found in the case of biotrickling and bioscrubber. Therefore, important process conditions like pH, salt content, nutrients, toxic intermediates or end products of microbial degradation are much easier to analyze and to control in biotrickling and bioscrubber. Also, the liquid phase itself (the water content) can be better controlled to obtain the optimal water content through the bioreactor system, and to minimize drying out

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Table 2a Comparative performance evaluation of bioreactors for VOC and odour control. Bioreactor Type

Target VOCs and Odours conc. g/m3

Treatment efficiency for Low conc. of VOCs /Odours

High conc. of VOCs /Odours

High water soluble VOCs

Low water insoluble VOCs

Fluctuating feed conditions

Biofilter Biotrickling filter Bioscrubber Membrane reactor

<1 <0.5

High High

Low Low

High High

Low Low

<5 High

High High

High High

High High

Low High

a

Pressure drop

Capital cost

Operational cost

Bioprocess controla

Low Low

Low Low

Low Low

Low Low

Low Low

High Need longterm evaluation

Very low Need longterm evaluation

Medium High

Medium High

High Need longterm evaluation

Comparative evaluation on bioprocess control parameter is given in Table 2b.

Table 2b Critical bioprocess control parameters for different bioreactor configurations. Type

Moisture

Nutrient/pH

Clogging

Transient response

Airflow channeling

Startup

Biofilter Biotrickling filter Rotating contactors Bioscrubber Suspended growth Membrane reactor

Highly sensitive Not sensitive Not sensitive Not sensitive Not sensitive Not sensitive

Highly sensitive Not sensitive Not sensitive Not sensitive Not sensitive Not sensitive

Sensitive Highly sensitive Not sensitive Not sensitive Sensitive Highly sensitive

Sensitive Highly sensitive Highly sensitive Highly sensitive Sensitive Sensitive

Highly sensitive Sensitive Not sensitive Not sensitive Not sensitive Not sensitive

Sensitive Highly sensitive Highly sensitive Sensitive Sensitive Highly sensitive

of the packing material or the biofilm on the packing material. Heslinga (1994) mentioned that probably 50–75% of the problems with conventional biofilters are related to a poor control of the water content in the biofilter media. An important disadvantage of biotrickling and bioscrubber reactors is the higher complexity to construct and to operate. The startup of biotrickling and bioscrubber reactors is also more complicated, since the inert medium does not contain microorganisms at the start. As soon as the microorganisms are present, they can be washed out by the required drainage of the process water; a problem that is encountered with full-scale operations of bioscrubbers. Pressure drop control is more complex especially with biotrickling reactors. Compared to a biofilter, higher air velocities through the media results more easily in a higher pressure drop. When biotrickling or bioscrubber reactors are applied for the treatment of pollutants like hydrogen sulfide, ammonia or chlorinated compounds, the degradation produces acid end products in the drain water, which needs further processing. Biotrickling and bioscrubber reactors make a better process control possible, which require, on the other hand, measurement or control instrumentation. The location of the water film with respect to the biomass differs in biotrickling filter, bioscrubber and membrane bioreactor. In the trickling bed reactor and in the bioscrubber pollutants have to diffuse through the water phase, before they can be consumed by the microorganisms. For pollutants with a poor water-solubility, such a layer of water causes a substantial additional resistance for mass transfer. In the membrane bioreactor, on the contrary, the liquid phase is situated at he opposite side of the biofilm and hardly forms a barrier for mass transfer of the poorly water soluble pollutants. Therefore, the removal of poorly water soluble pollutants from air can be considered as the most promising application for membrane bioreactors. The mass transfer resistance of membranes for this poorly water soluble pollutants is negligible. Moreover, the large gas–liquid interface of membrane modules enables efficient removal of these pollutants, which in general are difficult to remove from air. Other niches for the application of membrane bioreactors are indoor application and the removal of pollutants that require a specific microbial population, like TCE and nitrogen monoxide. Very recently,

a membrane bioreactor with both an anaerobic and an aerobic zone was proposed. Such a bioreactor might enable the biodegradation of pollutants, such as highly chlorinated hydrocarbons, that until now are considered to be beyond the reach of (aerobic) biological waste gas treatment. Disadvantages of membrane bioreactors are the high investment cost, particularly compared to other bioreactor, and possible clogging of the liquid channels due the formation of excess biomass. Compared to other types of bioreactors, the membrane may form an additional barrier for mass transfer. So far, membranebased biological waste gas treatment has only been tested on laboratory scale. If the long-term stability of these reactors can be demonstrated, membrane bioreactor technology can be useful in the treatment of gas streams containing poorly water soluble pollutants and highly chlorinated hydrocarbons, which are difficult to treat with conventional methods for biological waste gas treatment. 8. Other bioreactor configurations Apart from bioreactors described in the preceding sections for VOC and odour treatment, several other bioreactors have also been reported. An innovative design was reported by Yang et al. (2002), which consists of a rotating drum BF. Open pore reticulated polyurethane foam was used as the BF packing medium and this new design resulted in better distribution of VOCs, oxygen, nutrients, and biomass, over conventional BFs. Two types of rotating drum BFs were investigated to study the effect of medium configuration on BF performance for VOC treatment. One was a single-layer BF that consisted of a thick layer of open pore reticulated polyurethane foam media. The other was a multi-layer BF that used a set of four concentric thinner layers of the media. The effect of the two different media configurations was examined using diethyl ether at various organic loading rates. The results showed that the multilayer BF could maintain more stable and higher ether removal efficiencies at gas empty bed contact time (EBCT) of 30 s, when compared to the single-layer BF at gas EBCT of 90 s, and at organic loading rates ranging from 32.1 to 128.4 g (ether) m3 l1. The multilayer BF also exhibited a more even biomass distribution on the concentric surface at medium depth than the single-layer BF, which

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suggested a reduced possibility of short-circuiting of gas streams and, consequently, better performance (Yang et al., 2002). A novel rotating rope bioreactor was described by Mudliar et al. (2008a,b), especially for the treatment of VOCs characterized by high volatility along with high water-solubility. The bioreactor could be used for treatment of vapor phase VOCs by suitably scrubbing the compound in water and subjecting it to the new bioreactor referred to as the rotating rope bioreactor (RRB) for treatment. The novel bioreactor provided higher interfacial area (per unit reactor liquid volume) along with high oxygen mass transfer rate, greater microbial culture stability and consequently higher substrate loadings and removal rates in comparison to other conventional reactors (e.g. BFs) widely used for the treatment of VOCs. Pyridine was used as a model compound to demonstrate the enhanced performance of RRB. The experimental results showed that the novel RRB system was able to degrade synthetic wastewater containing pyridine with removal efficiency of more than 85% up to a loading of 66.86 g m3 h1. Further, the authors also described a single stage reactor called the rotating rope biofilter for direct treatment of VOCs instead of a two-stage process described above. This reactor is a modified closed RRB where the waste air containing the VOC is directly sparged through the water hold-up of the reactor and the water soluble VOCs are absorbed in the aqueous phase. It is then degraded by the microbial consortium immobilized on the RRB rope media. Kan and Deshusses (2003) reported a new type of bioreactor for air pollution control referred to as a foamed emulsion bioreactor (FEBR). The new reactor was based on an organic-phase emulsion and actively growing pollutant-degrading microorganisms, made into foam with the air being treated. As there is no packing in the reactor, the FEBR is not subjected to clogging. Mathematical modeling of the process and proof of concept using a laboratory prototype revealed that the foamed emulsion bioreactor greatly surpasses the performance of existing gas phase bioreactors. Experimental results showed a toluene elimination capacity as high as 285 g (toluene) m3 (reactor) h1 with removal efficiency of 95% at a gas residence time of 15 s and a toluene inlet concentration of 1–1.3 g m3. Study of toluene degradation in a flat plate vapor phase bioreactor by a Pseudomonas putida 54 G biofilm using oxygen micro sensors was reported by Villaverde et al. (1997). Oxygen microelectrodes were used to measure oxygen concentration profiles through the gas, liquid, and biofilm phases. The linear shape of the dissolved oxygen concentration profile in the outer 87% of the biofilm thickness suggested an absence of reaction in this layer. Oxygen consumption in the remaining 13% biofilm layer (0.3 mm) followed zero order kinetics with a rate constant of 102.2 g m3 h1, for toluene gas concentration of 1.5 g m3. The increase in respiratory activity near the substratum was confirmed by microscopic study of cryogenic biofilm sections, and the lack of activity in the surface film was interpreted as a consequence of injury exerted by the toxic substrate. The accumulation of dead cells on the top of the biofilm contributed resistance to the transport of substrates to deeper layers of the biofilm suggesting a protective role of the outer layer against the harmful effect of the toxic substrate. These results highlight a new conceptual biofilm model in which both microbial growth and inactivation are controlled by substrate transport, leading to a structure that by itself controls substrate availability. 9. Conclusions This review provides an overview of the various bioreactors used for waste gas treatment, limitations of existing bioreactors, and new avenues required in bioreactor development and design. Clearly, many of the bioreactor designs discussed herein still

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require improvement, and confirmation of significantly better performance compared to existing designs. For example, considering the volatile nature of VOCs and odourous compounds, suspended growth reactor systems/bioscrubbers may find limited application in the field at higher substrate concentration and loadings, and oxygen transfer can become a limitation, since air purging cannot be used in such systems. In future, improved ‘‘rational bioreactor design’’ techniques need to be gradually developed. The knowledge needed to do this can be developed by further research on topics such as the mechanisms of clogging, the kinetics of biofilm growth, and fundamental microbial ecology. Flow characterization through the bioreactor is important, since gas flow, liquid flow and gas velocity have an important impact on process parameters like mean gas resident time, gas dispersion in the reactor, and pressure drop over the system. These parameters are important to scale up and to operate a bioreactor at optimum conditions. The bioreactor development should be also focused on issues like robustness (flexible to process fluctuations/failures), large pollutant loadings, high temperatures, halogenated compounds and poorly water soluble compounds. It is a big and challenging task to design a bioreactor from fundamental theory, but definitely the understanding of biological treatment is growing with time. Further, developments of innovative combined bioreactor designs remain a high priority, since a single bioreactor configuration will never provide a universal solution to existing VOC and odour problems. In many instances, progresses in reactor design and development will require similar advances in understanding the fundamentals of the bioprocess, so that a more logical, creative and focused approach in bioreactor design can be implemented. Hence, to improve the performance of the biological air treatment system for VOCs and odours, there is a need for continuous innovation in bioreactor configurations. References Abumaizar, R.J., Kocher, W., Smith, E.H., 1998. Biofiltration of BTEX contaminated streams using compost-activated carbon filter media. Journal of Hazardous Material 60, 111–126. Acuna, M.E., Perez, F., Auria, R., Revah, S., 1999. Microbiological and kinetic aspects of a biofilter for the removal of toluene from waste gases. Biotechnology and Bioengineering 63, 175–184. Alexander, R., 1999. Compost markets grow with environmental applications. Biocycle 40, 43–48. Alonso, C., Zhu, X., Suidan, M.T., Kim, B.R., Kim, B.J., 2000. Parameter estimation in biofilter systems. Environmental Science and Technology 34, 2318–2323. Alonso, C., Zhu, X., Suidan, M.T., Kim, B.R., Kim, B.J., 2001. Mathematical model of biofiltration of VOCs: effect of nitrate concentration and backwashing. Journal of Environmental Engineering 127, 655–664. Armon, R., Laot, N., Lev, O., Shuval, H., Fattal, B., 2000. Controlling biofilm formation by hydrogen peroxide and silver combined disinfectant. Water Science and Technology 42, 187–193. Arnold, M., Reittu, A., Von, W.A., Martikainen, P.J., Suikho, M.L., 1997. Bacterial degradation of styrene in waste gases using a peat filter. Applied Microbiology and Biotechnology 48, 738–744. Arulneyam, D., Swaminathan, T., 2000. Biodegradation of ethanol vapour in a biofilter. Bioprocess Engineering 22, 63–67. Attaway, H., Gooding, C.H., Schmidt, M.G., 2001. Biodegradation of BTEX vapors in a silicone membrane bioreactor system. Journal of Industrial Microbiology and Biotechnology 26, 316–325. Attaway, H., Gooding, C.H., Schmidt, M.G., 2002. Comparison of microporous and nonporous membrane bioreactor systems for the treatment of BTEX in vapor streams. Journal of Industrial Microbiology and Biotechnology 28, 245–251. Aziz, C.E., Fitch, M.W., Linquist, L.K., Pressman, J.G., Georgiou, G., Speitel, G.E., 1995. Methanotrophic biodegradation of trichloroethylene in a hollow fiber membrane bioreactor. Environmental Science and Technology 29, 2574–2583. Bauerle, U., Fischer, K., 1987. Verfahren und Vorrichtung zur Eliminierung schwer wasserloslicher und leicht fluchtiger Verunreinigungen aus einem Abluft-bzw. Abgasstrom durch biologische Oxidation. Patent DE 3542599. Bauerle, U., Fischer, K., Bardtke, D., 1986. Biologische Abluftreinigung mit Hilfe eines neuartigen Permeationsreaktors. Staub Reinhalt Luft 46, 233–235. Binot, R.A., Paul, P., Keuning, S., Hartmans, S., de Hoop, D., 1994. Biological air filters. Part 1 – conception and design. Preparing for the future. ESA Technology Progress Quarterly 4, 14–15.

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