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Biofiltration is the most typical type of biological air pollution control process, initially developed in the late 1970s (Leson and Winer, 1991). It is now emerging as a sustainable alternative for the treatment of air contaminated with VOCs and odorous compounds. In biofiltration, the polluted air is forced through a bed of packing media covered with a layer of aerobic microorganisms. The microorganisms are immobilized on the surface of the packing media. The primary role of the packing material in biofilter media bed is to support the microbial community through the attachment of microbial biofilm to the surface of packed media. Bohn (1992) established that an ideal biofilter bed should have a high specific surface area (>300 m^-1) for the proper development of the micro flora which can induce high gas-to-biofilm mass transfer. However, specific surface area as high as 1000 m^-1 has also been reported for polyurethane-based beds (Mudliar et al., 2010). High porosity and water retention capacity are also highly desirable to facilitate homogenous distribution of gas flow and avoid media drying, respectively. The widely used packing materials for biofiltration include soil, compost, and wood chips. These materials are adavantageous, as they satisfy the basic requirements stated above and are cost effective (Mudliar et al., 2010). In order to avoid bed crushing and compaction and to improve many other properties such as moisture hold up and microbial growth, several authors have suggested the use of advanced packing materials with complex blending such as a mixture of compost and wood chips or compost mixed with hard plastic (Taghipour et al., 2008). The pollutants are transferred from air to the water layer adhering to the bacterial growth on the media to be biologically metabolized (Upadhyay and Kumar, 2004). Biofiltration is energy efficient and cost effective while producing minimal quantities of toxic end-product. This technology has been successfully used for removing a wide range of pollutants such as VOCs, ammonia, mercaptans, and sulfurous compounds (Giri et al., 2010; Galera et al., 2008; Ho et al., 2008). The drawbacks of this technique are excessive pressure drops and gradual accumulation of acidic by-products due to dry-out, rapid degradation, and clogging; difficulty in controlling the biological operating parameters; clogging due to the accumulation of large amount of biofilm and reduced treatment efficiency at high pollutant concentrations (Mudliar et al., 2010). A typical biofilter has been observed to operate with a removal efficiency in the range of 50-95% (Park et al., 2001).

The removal efficiencies of toluene were reported to exceed 80% with an inlet concentration less than 1,274 ppm using an agro waste based biofilter (Singh et al., 2006). Moreover, Jaber et al. (2016) studied the removal of H₂S using a biofilter under extremely acidic conditions. A maximum H₂S removal capacity of 24.7 g m^-3 h^-1 was reported for a 0.07 m³ reactor at an inlet flow rate of 4 m³ h^-1 with a removal efficiency of 78% up to 360 ppm (v/v).

A bioscrubber essentially consists of a two stage unit in which absorption occurs in one stage and biodegradation by suspended microbes occurs in the other stage. Bioscrubbers are usually used for the treatment of readily soluble VOCs in the waste air stream (alcohols, ketones, etc.) having a concentration less than 5 g m^-3 (Kellener and Flauger, 1998). Reported removal efficiencies are in the range of 50-99% (Webster and Devinny, 1995). Bioscrubbers are stable enough to allow a better control of operating parameters. Also, they produce a lower pressure drop across the microbe suspension than other filter types. The major problem associated with bioscrubbers is the generation of excess sludge and liquid waste which over time reduces the efficiency of the process considerably.

Membrane bioreactors have been investigated by various researchers for VOC and odor abatement (Mudliar et al., 2010; Kumar et al., 2008a, b; Shareefdeen and Singh, 2005; Ergas and McGrath, 1997). They were designed as an alternative to conventional bioreactors for waste gas treatment. In a membrane bioreactor, the mass transfer of VOCs from the gas phase to a microbially active liquid phase occurs through the microporous hydrophobic hollow fiber membranes. The selective permeation of the pollutant across the membrane occurs due to the concentration difference between the gas phase and the biofilm phase, which provides the driving force according to Henry’s law. The advantages of membrane bioreactors are the absence of moving parts, ease of scale-up, and the ability to vary the flow of gas and liquid independently without the problems of flooding, loading, or foaming. Disadvantages associated with membrane bioreactors are the high investment cost and possible clogging of the liquid channels due to the formation of excess biomass. Removal efficiencies have been reported in the range of 50-99% (Hartmans et al., 1992).

Capital costs for biotrickling filters vary widely according to filter size and construction materials. The required size of the biotrickling filter is a determined by such variables as air flow rate, the nature and concentration of the pollutant treated, and the required removal efficiency. The presence of corrosive gasses (e.g., H₂S) or solvent vapors is the main factor in determining construction materials (polyethylene, fiber-glass, perlite, etc.) (Popoola et al., 2013). The cost of operating a biotrickling filter is increased by, the presence of dust or fine particles, excessively high or low temperatures, highly fluctuating pollutant concentrations, etc. Hence, before commencing reactor design and construction, a detailed analysis and characterization of the reactors input air stream needs careful consideration.

A simple relationship (Equation 1) was proposed to estimate the capital cost of a biotrickling filter based on bed volume (Deshusses and Cox, 1999). The costs include basic components (e.g., pumps and level switches) for a simple biotrickling filter constructed out of inexpensive materials. The cost estimated by Equation 1 has a ±20% accuracy for reactor volumes 5 to 1000 m³. For more expensive materials such as stainless steel, the pre-exponent term (13,000) needs to be increased in equation 1.

The reactor volume can be determined from knowledge of the pollutant concentration, the intended removal efficiency, and the air flow rate. For the reactor capacities of 10, 100, and 1,000 m³, the unit cost decreases substantially with increasing reactor size, viz., 7,500, 4250, and 2400 (USD m^-3), respectively. Equation 1 is then used to estimate the capital cost (Table 1). The final installed cost is somewhat vendor dependent. Other costs such as: land, site preparation, assessment, maintenance, operating, financing, taxes, insurance, other overheads, etc. are also needed to determine the final installation and operational costs.

The operating cost estimation of a biotrickling filter should include: 1) Electricity for the blower and the recycle pump along with other electrical equipment, 2) cost of the water and nutrients, 3) maintenance, 4) costs associated with biomass growth control, 5) capital costs associated with amortization (Deshusses and Cox, 1999).

Electricity for the blower is often a major fraction of the total operating expenses. In contrast, water, nutrients, and chemicals needed for the control of moisture and pH are a relatively small fraction (10-30%) of the total operating costs. Inspection of spray nozzles for possible clogging is the most important task during maintenance, as it could lead to inadequate wetting of the bed. If the biotrickling filter is subject to clogging problems, the costs of controlling the biomass growth can be significant - up to half of the total operating costs (Deshusses and Cox, 1999). Note that there are various approaches to control biomass growth such as the use of ozone to curb biomass accumulation (Zhou et al., 2016) or by protozoan predation (Cox and Deshusses, 1999). However, those methods have not yet been reliably developed for the applications at the industrial scale. Careful evaluation of the various options is recommended. Since biotrickling filter operation is relatively inexpensive, capital cost amortization will be significant compared to other costs. Assuming the lifetime of a filter plant is 10-20 years, the amortization of capital costs represents on average 20 to 40% of the total treatment cost. This stresses the importance of careful selection of materials and proper sizing to minimize the actual capital costs.

Usual operating costs range from $0.05 to $1.5 per 1000 m³ of air treated excluding capital costs and increases to $0.1 to $3 per 1000 m³ if capital amortization is included (Deshusses and Cox, 1999). These cost estimates are inflation dependent and in the USA, there has been 45% inflation since 1999 to 2016 (http://www.bls.gov/home.htm)). The operating costs need to be carefully considered at the planning stage based on possible applications and biotrickling filter size.

Lu et al. (2001) achieved removal efficiencies as high as 95% for a mixture of acetone (20 g m^-3 h^-1) and methyl acetate (27 g m^-3 h^-1) using a bench-scale biotrickling filter. The filter comprised of an acrylic cylinder packed with phanerochaete chrysosporium immobilized on glass beads. Clogging of the medium, the complex filter structure and operation of the biotrickling filter, were the only shortcomings reported.

The effect of low dose ozonation was also investigated to prevent excess biomass accumulation and to maintain high removal efficiencies of toluene over extended BTF operation (Zhou et al., 2016). To optimize the biomass control strategy, the relative performance of five parallel BTFs was monitored at different ozone doses. The BTF was constructed from a Perspex pipe with a height of 0.95 m and an internal diameter of 9 cm. The active bed height was 48 cm with the bed volume of 3.1 L. The BTFs were packed with pelletized polyurethane foam (PUF) with a diameter of 10-15 mm. The pelletized PUF had an initial porosity of 91.0% with a specific surface area of 380 m² m^-3. The ozonefree BTF performance declined after 150 days due to excess biomass accumulation, the buildup of extracellular polymeric substances (EPS) excreta, and a decline in metabolic activity of the biofilm. An optimized dose of ozone (e.g., 5-10 mg m^-3, or 2.55-5.09 ppm) was sufficient to maintain stable operation (for 300 days) at a consistently high removal efficiency (>93%) at 400 mg m^-3 (106 ppm) toluene in the air; this prevented excess biomass accumulation. On the other hand, ozone above 20 mg m^-3 (10 ppm) inhibited excessive biomass growth to prevent poor BTF performance.

The biodegradation of toluene vapor was investigated by a lab-scale biofilter impregnated with pseudomonas putida DK-1 (Park et al., 2001). Removal efficiencies in the range of 75-90% were observed for an inlet loading ranging from 250-350 g h^-1 m^-3. They used a fiber-glass column with an inner diameter of 50 mm and height of 200 mm. The pressure drop in the bed was 20-100 (Pa m^-1 of packing) and had limited impact on the process efficiency. At the bottom of the reactor, a perforated sieve plate was to support the medium. The medium (wood chips) was impregnated with pseudomonas putida DK-1 and positioned over samples of contaminated soil. The problems associated with the filter were the difficulty in controlling the moisture and pH in order to maintain the optimum environment for the growth of degrading microorganisms and also the relatively high rates of clogging and deterioration of the medium. From these two studies. Biofilters have more drawbacks compared to biotrickling filters. Moreover, biotrickling filters have been shown to reach higher removal efficiencies for relatively low concentrations of VOCs and odorous compounds.

First ever reported laboratory scale BTF for the elimination of nitrobenzene vapors was reported by Oh and Bartha (1997). They used a stable microbial consortium enriched by sewage sludge and immobilized on dry perlite (the reactor occupied 59% (0.4 L) of the total column volume (1.5 L)). During the startup period of four weeks, the inlet nitrobenzene concentration was kept relatively low (<16 ppm) 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 m^-3 and an empty bed gas contact time of 21 seconds. The resultant elimination capacity was of 50 g m^-3 h^-1 at a stream flow rate of 200 m³ h^-1. This is a significant removal rate that could lead to an economically viable process. A nitrogen balance showed that 98% of the removed nitrobenzene was converted into ammonia while a small amount of nitrite was also produced.

Also, noteworthy is the study by (Sun and Wood, 1997). They immobilized a pure culture of Burkholderia Cepacia PR1₂₃, a Tn5transposon mutant of B. cepacia G4 that constitutively expresses the trichloroethylene (TCE) degrading enzyme and toluene orthomonooxygenase (TOM). Aerobic biodegradation of TCE only occurs through co-metabolism with the addition of a growth substrate (usually toluene, methane, propane, phenol or ammonia). This is required to induce the expression of the appropriate TCE-degrading enzyme. However, the bacterium strain Burkholderia Cepacia PR1₂₃ expresses toluene ortho-monooxygenase at a constant rate, regardless of physiological demand. This circumvents the problem of competitive inhibition of TCE oxidation by the usual inducers during the growth phase. They used glucose as a carbon and energy source and observed TCE eliminations up to 200 times higher than previously reported 90% TCE removal at an inlet loading of 2.4-100 mg TCE L^-1 together with 95 mg toluene L^-1 (Guo et al., 2001). As observed previously in other bioreactors for TCE aerobic cometabolism, rapid inactivation of the TCE-degrading enzyme by TCE breakdown products (e.g. TCE-epoxide) still remains to be resolved (Guo et al., 2001; McFarland et al., 1992).

In the past few years, much research has been done to improve the BTF technology (Valero et al., 2017; Zhou et al., 2016; Tsang et al., 2015). Several successful conversions of full-scale chemical scrubbers to biotrickling filters have been demonstrated (Gabriel and Deshusses, 2003; Kraakman, 2003, 2001).

There are over 16,000 publicly owned treatment works (POTW) in the United States serving 75 percent of the total population (U.S. DHS, 2016). The POTWs treat 32 billion US gallons (120 gigalitres) of wastewater every day (EPA, 2014). Emission of objectionable odors from these facilities is a major problem. POTW off-gases contain a wide range of odorous compounds, air toxics and volatile organic compounds (VOCs). These include volatile sulfur compounds, ammonia, benzene, toluene, chloroform, dichloromethane, and trichloroethylene (Lewkowska et al., 2016; Zhou et al., 2016). H₂S is the principal odorous component that causes nuisances even at volume fractions as low as 8 ppb (Smet et al., 1998). In addition to its unpleasant odor, H₂S gas is highly toxic (Roth, 1993). Continuous exposure to low (15-50 ppm) concentrations will generally cause irritation to mucous membranes, and may also cause headaches, dizziness, and nausea. Higher concentrations (200-300 ppm) may result in respiratory arrest leading to coma and unconsciousness. Exposure for more than 30 minutes at concentrations greater than 700 ppm have been fatal (MSDS, 1996). Concerns about the odor nuisance to the surrounding communities as well as the implementation of more stringent regulations are forcing POTWs to treat their off-gases.

Controlling H₂S is usually achieved by wet or chemical scrubbers. Chemical scrubbing in a packed tower is an established technique and is effective at gas contact times as short as 1.3-2 s (Gabriel and Deshusses, 2003). However, chemical scrubbing suffers from important drawbacks such as high operating costs, generation of halo-methanes that are known air toxics and can contribute to global warming (Wenhai et al., 2016), and the requirement for hazardous chemicals that pose serious health and safety concerns. Hence, more research is being done to convert chemical scrubbers into biotrickling units.

Gabriel and Deshusses (2003) demonstrated the conversion of a chemical scrubber located at the Orange County sewerage treatment facility in California. The converted bioscrubber was 9.75 m high, 1.82 m in i.d., and made of fiber-glass reinforced plastic with a nominal packing bed height of 2.8 m and bed volume of 7.3 m³. The water trickling flow rate was 4.5 m³ h^-1 and the foul air was fed to the reactor at atmospheric pressure with an average flow rate of 16,300 m³ h^-1. The researchers operated a laboratory pilot unit under conditions similar to those expected at a publicly owned treatment work and tested selected packing materials for their sustainability for biotrickling filtration. The conversion consisted of (i) replacing the existing packing, which had a low interfacial area and was not suitable for microorganism attachment; (ii) replacing the liquid recycle pump with a smaller one; (iii) disconnecting the chemical feeds; and (iv) modifying controls of the reactor. Before startup, the reactor was impregnated with 0.8 m³ of activated sludge from wastewater-treatment plant. The pH started to decline after three days of operation to reach pH 2, in seven days after startup. The decline in pH (from the production of H⁺ and sulfate from the oxidation of H₂S) was correlated with the increase in H₂S removal efficiency. Acclimation lasted for about ten days, after which H₂S removal was more than 99% for H₂S inlet volume fractions ranging from 5 to 25 ppm and remained high for the rest of the operation.

Vikromvarasiri and Pisutpaisal (2016), studied the removal of H₂S in a biotrickling system using a new bacterial strain of obligately chemolithoautotrophic, Halothiobacillus neapolitanus NTV01 (HTN). The biotrickling filter column was made from glass (0.475 m inner diameter and 0.72 m height). The column was packed with randomly structured packing media (GEA2H Water Technologies GmbH) to its working height of 0.282 m. The packing media had a random structure and made from high-density polyethylene (HDPE) 12 mm beads, surface 859 m² m^-3, and density 150 kg m^-3. A mixed gas stream containing CH₄, CO₂, and H₂S in the air was supplied to the inlet of the biotrickling filter; and [H₂S] varied between 0-255 ppm (v/v). The system could completely remove 45 ppm H₂S within 70 min of operation. The removal efficiency was in the range of 95-98%, for 225 ppm (v/v) H₂S.

The removal of H₂S in high-performance biotrickling filters was investigated by Kim and Deshusses (2005) using a differential biotrickling filter. The filter was designed to reach high gas velocities through a miniature packed bed. A small differential filter was used in this study. It was filled with a single cube (64 mL) of open-pore polyurethane foam packing. The biotrickling filter was operated in a counter-current mode. The air flow was circulated in a closed loop from an 85 L Tedlar bag to the differential biotrickling filter by a 0.2 horsepower blower up to a maximum linear velocity of 3 m s^-1. Pure H₂S was injected into the differential biotrickling filter system using a 20 mL syringe. Within a short time, gas velocity was varied to determine its effect on H₂S elimination capacity. The H₂S elimination capacity (35 to 125 g m^-3 h^-1) was achieved at a liquid trickling velocity of 1.5 m h^-1 and inlet [H₂S] was between 50 and 65 ppm. Interestingly, the addition of dissolved sodium sulfide (1-3.5 mg L^-1) resulted in reduced H₂S gas degradation at pH 1.9-2.1. H₂S is utilized by the microbial population as an S source. Hence, the addition of sodium sulfide competes with H₂S gas for biodegradation as the only ionic species (from dissolved H₂S and Na₂S), i.e. SH-, are metabolized by the bacteria.