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Biofiltration as an odour abatement strategy
Biochemical Engineering Journal 5 (2000) 231–242
Biofiltration as an odour abatement strategy Dennis McNevin a , John Barford b,∗ a
b
Department of Chemical Engineering, University of Sydney, Sydney, NSW 2006, Australia Department of Chemical Engineering, H.K.U.S.T., Clear Water Bay, Kowloon, Hong Kong (S.A.R.), China Received 18 December 1999; accepted 21 March 2000
Abstract The chemical, physical and biological processes occurring in biofiltration are reviewed. A survey of operating biofilter performances is also presented and includes some novel comparative methods. It is concluded that biofiltration is a simple and cost-effective technology for odour removal and that an understanding of the many interactions occuring within the biofilter is essential for the optimal performance of the biofilter. © 2000 Elsevier Science S.A. All rights reserved. Keywords: Biofiltration; Biofilter performance; Removal efficiency; Maximum load; Removal rate; Operating parameters
1. Introduction The removal of nutrients from wastewaters and odours from air is often effected through biological means in unit operations like biofilters, biotrickling filters and bioscrubbers. There are successful applications of these in the meat meal industry [1], the tobacco industry [2], hardboard production [3], wastewater treatment [4] and a number of reviews attest to their overall efficacy for odour control [5–8]. Removal of hydrogen sulfide from gas streams has been particularly successful [9,10]. A biofilter consists of a container of organic material populated with micro-organisms through which odorous air is passed, usually upwards. The influent air may be prehumidified to maintain adequate moisture in the organic bed. Alternatively, or in addition, water may be sprinkled over the surface of the bed, trickling downwards, counter-current to the odorous air. This water may contain nutrients required by the micro-organisms for growth. Liquid trickling from the bottom of the biofilter may be bled off or recycled (Fig. 1). Odorous contaminants are transferred from the air into an aqueous, bio-active layer (biofilm) surrounding organic particles in the bed. The contaminants are then aerobically degraded to various end products or incorporated into biomass. The end products will depend on the nature of the contaminants. Biofiltration enhances the natural processes of bioremediation where contaminants in the atmosphere ∗ Corresponding author. Tel.: +852-2358-7237; fax: +852-2358-0054. E-mail address: [email protected] (J. Barford)
are degraded by micro-organisms in soils after diffusion into soil pores. A well engineered biofilter provides improved contact between organic particles and air containing contaminants. Compost and soil pores are organically coated which enhances sorption of organic contaminants while the aqueous film surrounding these particles facilitates the sorption of inorganics. The distinction between the organic coating and the aqueous film is unclear and they are often collectively called the ‘biofilm’. Biofilters, biotrickling filters and bioscrubbers may be generically referred to as organic perfusion columns. They consist of three phases in intimate contact: a solid, organic phase, a liquid phase and a gas phase. All three may contain nutrients for degradation. The liquid phase (water or nutrient solution) and gas phase are passed through the solid organic medium, invoking the processes of adsorption and aerobic biological degradation of the nutrients in the liquid. These nutrients may have been transferred from the air, as is the case with odour removal. The relative extent to which adsorption or biological degradation dominates in nutrient removal may vary with the organic medium used, as well as with local conditions like pH and temperature. The interdependence of these processes is also unclear. Adsorption of solutes onto the surface of peat may benefit or hinder their subsequent degradation by micro-organisms. Some biological conversions important in biofiltration of odorous compounds, together with typical bacteria responsible for them, are shown in Table 1. Kowal et al. [11] provide evidence to suggest that removal of hydrogen sulfide in a dry activated sludge biofilter occurs by way of three distinct phases: (i)
1369-703X/00/$ – see front matter © 2000 Elsevier Science S.A. All rights reserved. PII: S 1 3 6 9 - 7 0 3 X ( 0 0 ) 0 0 0 6 4 - 4
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where Cg,in and Cg,out (g m−3 ) are the concentrations of contaminant in the inlet and outlet gas streams for the biofilter, respectively. It has been found that: 8 = 1 − exp(−µ1 τ )
Fig. 1. Generalised schematic diagram of a biofilter for odour removal.
absorption into the water present in the organic bed; (ii) adsorption onto the solid surface and biodegradation. The process denoted by (i) above is well described by standard thermodynamic vapour–liquid equilibria and is fast enough (relative to adsorption and biodegradation) to be considered instantaneous (always at equilibrium), especially when air phase residence times are typically of the order of minutes in biofilters. Either mass transfer resistance to adsorption or biodegradation will be the limiting process before equilibrium is achieved in a biofilter. The process of biodegradation is usually described by the Monod relationship, which relates the biodegradation rate to the substrate concentration Cg Rate = Maximum rate × K + Cg where Cg is the gas phase substrate concentration and K is the saturation constant. Depending on the concentration of substrate, the rate can be zero order (high concentration) or first order (low concentration).
(2)
where τ (h) is the mean residence time for the air phase in the biofilter and µ1 (h−1 ) is a rate constant, thought to be mostly a function of contaminant biodegradability. This implies that the microbial degradation rate is often first order with respect to contaminant concentration in the gas phase, as a first approximation. It also suggests that the most important design parameter for biofiltration is the biological degradation rate and biofilters have traditionally been designed thus. However, on careful consideration of the processes involved in biofiltration, mass transfer and maximum removal rate are also important. Consider a biofilter with a population of micro-organisms capable of degrading a contaminant with rate r (g m−3 h−1 ), subject to first-order kinetics according to r = µ 1 Cg
(3)
where Cg (g m−3 ) is the gas phase concentration of contaminant anywhere in the biofilter. At steady state, a mass balance on the contaminant over an incremental length, dz (m), of the biofilter yields: Fg Cg |z − Fg Cg |z+dz = µ1 Cg εA dz
(4)
where Fg (m3 h−1 ) is the gas phase flowrate, A (m2 ) the biofilter cross-sectional area and ε (m3 m−3 ) represents the voidage occupied by the gas phase in the biofilter. Collecting terms and integrating over the total length, Z (m) of the biofilter yields Cg,out εAZ = exp −µ1 (5) − exp(−µ1 τ ) Cg,in Fg where
2. Biofilter performance measures
τ=
εAZ Fg
(6)
2.1. Removal efficiency 2.2. Maximum load It is common to describe the performance of a biofilter in terms of removal efficiency (8) defined as [7] 8=1−
Cg,out Cg,in
(1)
The parameter 8 does not indicate how much contaminant is being removed in absolute terms. A 99% removal from an air stream with 10 ppm ammonia can not be distinguished from 99% removal of a 1000 ppm stream when
Table 1 Biological conversions important in biofiltration of odorous compounds Transformation
Typical bacteria
Oxygen environment
Organic carbon oxidation VOC →CO2 , H2 O Nitrification NH4 + →NO2 − , NO3 − Sulfide oxidation H2 S→S0 , SO4 2− Denitrification NO3 − →N2
Chemoheterotrophic bacteria Nitrifying bacteria Sulfur oxidising bacteria Denitrifying bacteria
Aerobic Aerobic Aerobic Anaerobic
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obviously the latter is removing a great deal more at the same flowrate. With increasing concentration in the inlet, the rate of biological degradation will eventually cease to be first order and become zeroth order (i.e. the kinetics will become ‘saturated’ or independent of concentration). Hence, a maximum load, L (g m−3 ) must be identified, above which 8 will decrease. At this maximum load, a mass balance on the contaminant over an incremental length of the biofilter yields
Table 2 Normalised variables used for simple biofilter model
Fg Cg |z − Fg Cg |z+dz = µ0 εA dz
C dC = − dz 1+C
(7)
where µ0 (g m−3 h−1 ) is a zeroth order biological degradation rate constant. Collecting terms and integrating over the length of the biofilter yields: L = Cg,in − Cg,out
εAZ = µ0 = µ0 τ Fg
(8)
The removal efficiency 8 now depends on the inlet flow concentration: 8=1−
Cg,out µ0 τ = Cg,in Cg,in
(9)
2.3. Removal rate A biofilter removal or elimination rate R (grams of contaminant removed per kg bed medium per hour), can be defined for an individual biofilter. This is the rate of removal of contaminant per unit mass of biofilter bed medium and is given by: R=
Fg µ0 τ Fg L µ0 εAZ = = m m m
(10)
where m (kg) is the mass of bed medium in the biofilter. R may also be defined on a unit volume basis (grams of contaminant removed per m3 bed medium per hour) or on a unit surface area basis (grams contaminant removed per m2 biofilter cross-sectional area per hour).
Normalised variable
Definition
Significance
C
Cg /K
z
z/z
Gas phase concentration normalised with respect to the half saturation constant Distance through the biofilter normalised with respect to the total height
where µ0 τ = K
(13)
(14)
Dimensionless groups are introduced to enable the comparison of different biofilters of different scale and operating conditions. The dimensionless parameter indicates the relative effectiveness of the biofilter. The dimensionless parameter (Cin −Cout )/8 indicates the relative removal rate. A high value of means that the biological degradation rate and/or the biofilter volume is high compared with the gas flowrate and hence removal is enhanced. Eq. (13) represents a macroscopic model of biofiltration and is the result of a steady-state mass balance on gas phase contaminants. It assumes a constant ratio of gas phase concentration to solid/liquid phase concentration. In reality, this ratio will vary with temperature and pH. The model is satisfactory as a first approximation but does not take account of mass transfer effects which are considered later. Eq. (13) was integrated numerically over a range of values for and Cin from Cin at z=0 to Cout at z=1 (Figs. 2–4). At low inlet concentrations, relative to the half saturation constant (Cin 1), a characteristic first-order concentration profile is observed through the bed (Fig. 2). At high concentrations (Cin 1), the profile is characteristically linear (Fig. 4). It is then possible to plot the dimensionless loading rate, Cin /, against the dimensionless removal rate
2.4. Effect of loading rate on removal rate The transition from first to zeroth order biological degradation kinetics due to increased load can be represented by the well known Monod [12] expression such that r = µ0
Cg K + Cg
(11)
where K (g m−3 ) is the so-called half saturation constant. When Cg
(12)
Variables can be normalised as in Table 2 which then yields
Fig. 2. Concentration profiles in a biofilter bed for a low inlet concentration relative to the half-saturation constant (Cin =0.1; predominantly first-order kinetics).
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to cope with the excess loading. However, increasing the inlet concentration will allow the biofilter to provide greater maximum elimination rates until a limit is reached here, too. This is due to the effect of high concentrations (C K or C 1) on the Monod kinetics of degradation when the kinetics become saturated. The absolute maximum elimination rate, obtained at high gas flowrates and high inlet concentrations, is given by Cg,in − Cg,out Cin − Cout = =1 µ0 τ Fig. 3. Concentration profiles in a biofilter bed for an intermediate inlet concentration relative to the half-saturation constant (Cin =1; zeroth- and first-order kinetics).
(15)
for which the conditions according to Fig. 5, assuming constant biological parameters, µ0 and K, are Cg,in Cin = >1 µ0 τ
(16)
and Cin > 100(Cg,in > 100 K)
Fig. 4. Concentration profiles in a biofilter bed for a high inlet concentration relative to the half-saturation constant (Cin =10; predominantly zeroth-order kinetics).
(Cin −Cout )/ (Fig. 5). As the loading rate increases due to increased flowrate of gas, the elimination rate increases until a maximum (saturated) elimination rate is achieved. At this inlet concentration, further increases in gas flowrate will mean larger outlet concentrations as the biofilter is unable
Fig. 5. Dimensionless removal rate (vertical axis) vs. dimensionless loading rate (horizontal axis) for a biofilter with Monod kinetics.
Under conditions of very low Cin , and with first-order kinetics, then, (Cin −Cout )/=Cin / (1−exp(−)) can be obtained and the plot of (Cin −Cout )/−v−Cin / would be independent of Cin . Fig. 5 suggests that if first order or Monod Kinetics are in force in a biofilter, then any correlation of loading versus removal rates must supply data points that lie on or to the right of the 100% removal line (line with slope of unity passing through the origin). Obviously, the 100% removal line represents the best possible performance from a biofilter and as the biofilter becomes more overloaded, its absolute removal capacity ceases to increase. A maximum removal rate for the well studied hydrogen sulfide removing biofilters was found by Yang and Allen [9] and Brennan et al. [13]. They also reported experimental data consistent with Fig. 5. Thus, given that the kinetics of degradation are quantitatively well understood for a solid/liquid system, a biofilter can be designed to operate at a maximum removal rate and achieve a desired removal efficiency. If there is any inhibition of biological growth by the degraded contaminant then the maximum removal rate will not remain constant as loading rate increases past saturation. As the loading of contaminant increases, the removal rate will decrease from a maximum as biological growth is inhibited. This was found to be the case when Tang et al. [14] increased the loading of malodorous triethylamine past 150 g m−3 h−1 in their compost/chaff biofilter, corresponding to a maximum removal rate of 140 g m−3 h−1 . This treatment of biofilter removal capacity is based on the very limiting assumption that there is no resistance to mass transfer from the gas phase to the biofilm surrounding the solid organic particles in the bed medium. There is a risk when designing biofilters for very easily degradable contaminants that a short residence time τ will be employed which will be insufficient for mass transfer to occur. This is borne out in the results of Yang and Allen [9] where removal efficiency in their hydrogen sulfide removing biofilter
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dropped from a maximum when retention time fell below about 23 s. They were able to show that they were operating at removal rates R, well below the maximum obtainable by their biofilter. Hence, a minimum residence time for mass transfer is also required when designing gas flowrates for biofilters.
2.5. Thiele number A common method of identifying the efficiency of a biofilter is by quoting a dimensionless Thiele number φ which indicates the relative importance of biological degradation versus diffusion/mass transport with respect to the elimination of contaminants in the biofilm. For first-order degradation kinetics, the Thiele number is defined as [15–17]: s φ=
µ1 δ 2 D
(18)
where δ (m) is the biofilm thickness and D (m2 h−1 ) is the effective diffusion coefficient for the contaminant in the biofilm. Zeroth order Thiele numbers are discussed. The biofilm will become diffusion limited as φ increases. A diffusion limited biofilm requires a longer air phase residence time than suggested by consideration of biological degradation rates alone. Hence, for efficient operation of a biofilter, the air phase residence time τ should be adjusted such that it is long enough so that mass transfer of contaminants from air phase to biofilm can occur and that the kinetics of biological degradation are nearly ‘saturated’. However, it should not be much longer than this if a maximum removal rate is desired. Diluting an odorous air stream by two and doubling the flowrate of the air to a biofilter will not necessarily give the same removal rate R or the same removal efficiency 8. The maximum removal rate will only be preserved during load fluctuations if the kinetics of degradation are predominantly zeroth order or linear (Cg K). However, in this regime, the removal efficiency will suffer with increased flowrate. Conversely, at very low concentrations (Cg K), when the kinetics of degradation are first order, the removal efficiency will be independent of the inlet concentration but the removal rate will increase with concentration.
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3. Survey of biofilter performances The three parameters, 8, L and R, can be used to compare the performance of biofilters. Reported concentrations of odorous compounds are usually expressed on a volume basis, ppmv (parts per million by volume), or on a mass basis, Cg ppmm (mg m−3 ). In order to compare the two, it is necessary Cg to know the temperature of the air carrying the odorous compounds. At 25◦ C 1 M ppmm ppmv ppmv =Cg × × M × 1000 = Cg Cg 24465 24.465 where M (g mol−1 ) is the molecular mass of the compound. The removal rate R can also be expressed on a mass basis (g g−1 h−1 ), a volume basis (g m−3 h−1 ) or a biofilter cross-sectional area basis (g m−3 h−1 ). Comparison between these three bases requires knowledge of the packing density of the filter medium or the biofilter bed depth which are not always reported. Performances surveyed here have been divided into those pertaining to ammonia, hydrogen sulfide and VOC removal, according to the classification of odorous compounds elucidated earlier. These three compounds form the bulk of the work reported in the literature. In addition, performance based on olfactometry measurements has also been included. 3.1. Ammonia removal The performances of several ammonia removing biofilters found in the literature are presented in Table 3. Reported research of biofiltration of ammonia is limited, perhaps because nitrifying bacteria responsible for the oxidation of ammonia exhibit slow growth rates. Biological ammonia removal may be disguised by absorption in moisture present in the biofilter and adsorption onto the surface of the organic filter material. Any leachate removed from a biofilter exposed to ammonia in the inlet gas will contain significant ammonium [18]. 3.2. Hydrogen sulfide removal The performances of several hydrogen sulfide removing biofilters found in the literature are presented in Table 4. Removal rates vary wildly but removal efficiencies are always at least 99%. In a thorough examination of
Table 3 Performance of ammonia removing biofilters in the literature Authors
8
L (mg m−3 )
R (g m3 h−1 )
Notes
Pinnette et al. [18] Kapahi and Gross [19] Bonnin et al. [20]
0.964 0.983
3.0 580 1.43
1.0 10.6 0.2388a
Compost, bark mulch, wood chips (estimated from nitrate production) Compost, oyster shell and perlite blend at landfill Peat biofilter at wastewater treatment plant
a
Unit in g m2 h−1 .
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Table 4 Performance of hydrogen sulfide removing biofilters in the literature Authors
8
L (mg m−3 )
R (g m−3 h−1 )
Notes
Yang and Allen [9] Ergas et al. [10] Brennan et al. [13] Pinnette et al. [18] Bonnin et al. [20]
0.999 0.999 0.99
3690 28 83 0.0308 32
130 420 8.3 2.361 10
Compost, bench scale Compost, oyster shell and perlite Peat, bench scale Compost, bark mulch, wood chips Peat biofilter at wastewater treatment plant
0.99
Table 5 Performance of VOC removing biofilters in the literature Authors
8
L (g m−3 )
R (g m−3 h−1 )
Notes
Knauf and Zimmer [3] Ergas et al. [10] Tang et al. [14] Kapahi and Gross [19] Togna and Singh [21] Hodge and Devinny [22] Deshusses et al. [23] Rothenbühler et al. [24]
0.95 0.95 1.00 0.86 0.95 0.40
0.14 (organic C) 2.656 2 40 (ppmv)
0.90
6.1
20 0.130 140 0.314 175 29 121 80a
Exhaust air from a hardboard production plant Benzene from sewage treatment plant Triethylamine Non-methane hydrocarbons at landfill Ethanol, bench scale Ethanol, bench scale Methyl ethyl ketone Printing solvents; peat ball medium
a
Unit in g C
4.4
m−3 h−1 .
hydrogen sulfide biofiltration, Yang and Allen [9] determined a maximum H2 S elimination capacity for their biofilter by plotting H2 S removal rates (g(m3 of bed)−1 h−1 ) against H2 S loading rates (g(m3 of bed)−1 h−1 ). At loading rates greater than 150 g m−3 h−1 , the removal rate levelled off at approximately 130 g m−3 h−1 , implying that a maximum biological degradation rate had been achieved. Their biofiltration systems, containing various types of yard waste compost as the filter material, removed H2 S with efficiencies greater than 99.9% for H2 S inlet concentrations in the range 5–2650 ppmv. Theirs is the only biofilter for which this absolute maximum removal rate was determined.
requires a residence time greater than three minutes. The inlet gas loadings of aromatics and H2 S were subject to wide fluctuations showing that the biofilter was capable of handling load variations without a significant effect on removal efficiency. Chlorinated aliphatic compounds were removed in pilot-scale experiments although the concentrations were not reported. Williams and Miller [8] provide further reported removal capacities in their review of operational biofilters. 3.4. Odour removal as measured by olfactometry The performances of two odour removing biofilters found in the literature, as measured by dynamic olfactometry, are presented in Table 6.
3.3. VOC removal The performances of several VOC removing biofilters found in the literature are presented in Table 5. Ergas et al. [10] were able to consistently obtain greater than 90% removal of difficult-to-degrade aromatic compounds in field studies at a water treatment plant, corresponding to average mass removal rates of 8 mg benzene per minute in a 3 m2 ×1.2 m biofilter bed (130 mg m−3 h−1 ). Removal of benzene, toluene and xylene decreased when the mean residence time for air in the biofilter was decreased from 3 to 1 min, indicating that complete removal of aromatics
4. Operating parameters 4.1. Bed media Table 7 lists bed media types found in the literature for removal of odorous compounds via biofiltration. Typical materials used include soil, peat, compost, bark mulch, perlite, sewage sludge and combinations of these. Multiple
Table 6 Performance of odour removing biofilters in the literature as measured by dynamic olfactometry Authors
8
L (OU m−3 )
R (OU m−3 h−1 )
Notes
Pinnette et al. [18] Bonnin et al. [20]
0.91 0.96
309 3910
14832 651667a
Compost, bark mulch, wood chips above sludge composting facilities Peat biofilter at wastewater treatment plant
a
Unit in OU m−2 h−1 .
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Table 7 Types of bed media used for biofiltration in the literature Type
Qualities
Bark compost, woodchips Compost, sewage sludge Compost, perlite, crushed oyster shells Sod-peat nodules and extruded peat granules Sieved compost and chaff Compost, bark mulch, wood chips Compost, perlite, zeolite, oyster shells Maërl (marine mineral containing 82% CaCO3 )
Low pressure drop. Removes VOC Removes H2 S Removes aromatics, H2 S Low pressure drop. Removes H2 S, methyl mercaptan Removes triethylamine. Chaff reduces pressure drop. Removes odour from sludge composting facilities Removes NH3 , VOC Provides alkalinity for autotrophic biological degradation
layers of biofilter media have been recommended by some manufacturers. Bed depths are of the order of 1 m for many industrial applications [1,3]. In their review of operational biofilters, Williams and Miller [8] have observed that biofilter bed depths vary from 0.5 to 2.5 m. Greater bed depth requires less land area but results in higher pressure drops across the bed. Biofilter bed life varies from 3 months [19] to 4 years [13,18]. There are a number of parameters used to describe the quality of a particular biofilter bed material. Some qualities, like high surface area for adsorption, adequate moisture content and large voidage (and hence low pressure drop across the bed) are universally required for biofiltration applications while others, particularly pH, are dependent on the application. 4.2. Moisture content Williams and Miller [8] point to bed moisture content as the single most important parameter for biofilter viability. Optimal moisture contents varied from 20 to 60% in their review of operational biofilters. Biological activity ceases if the moisture content of an organic material is too low. In addition, cracks open in a dry bed and channelling occurs, further limiting performance [25]. Conversely, too much moisture leads to anaerobic zones forming in the bed where oxygen required for bio-oxidation is depleted. For this reason, the capacity of soil beds to remove odour drops markedly when they become too wet. In addition, mass transfer of odorous compounds to the organic surface through a sodden bed, as opposed to through a thin film of water surrounding organic particles, becomes rate limiting. Pinnette et al. [18] suffered a loss of biological degradation of odorous compounds above sludge composting facilities when moisture content dropped below approximately 40%. The biofilter recovered its previous performance within 3 months after watering of the filter bed. Yang and Allen [9] found that above a water content of 30 wt.%, there was not much effect of water content on H2 S removal capacity in their bench scale biofilter. Below 30 wt.% water, however, the removal efficiency decreased proportionately. Composts were able to recover from being dry to regain their former
pH 3.2 6.5–7.0 7
Authors Knauf and Zimmer [3] Yang and Allen [9] Ergas et al. [10] Brennan et al. [13] Tang et al. [14] Pinnette et al. [18] Kapahi & Gross [19] Bonnin et al. [20]
H2 S eliminating capacity. The time required for recovery was from 1 to 3 days. Ottengraf and Van den Oever [25] kept moisture content in their VOC-eliminating biofilter at 50–70%. At low water levels, the organic packing lost its microbial activity while high water content promoted the development of anaerobic zones in the bed. Removal of aromatic VOC and hydrogen sulfide in the odour-removing biofilter of Ergas et al. [10] increased dramatically and almost immediately when moisture content was increased from below 50 to 55%. Peat and compost have good water holding capacities. Microbial activity in peat falls off if water content drops below 70% and rises above 85%. Below about 30%, biological activity usually ceases [26]. Soil is much less permeable than compost due to smaller pore sizes. It is, however, hydrophilic where as dry compost is hydrophobic and is difficult to rewet after drying out due to its highly porous structure. Compost also cakes when lime is added [7]. Heat generated by biological activity in a biofilter may increase the temperature of the bed medium above that of the inlet gas phase. Even if this gas enters the biofilter saturated with water, it will become unsaturated as its temperature rises after contact with the bed medium. Drying of the bed medium will inevitably occur. This will be offset to some degree by biological degradations that produce water (e.g. biological oxidation of VOC) [27]. Hence it is important to supply 100% humid air to a biofilter and/or water the bed to replace moisture lost to the gas phase if the organic bed is to remain viable. 4.3. Porosity Porosity, voidage and bulk density essentially relate to the same property of organic bed media and are important primarily for the effect they have on the gas phase pressure drop across the bed. Porosities for organic media range from 40 to 50% for soils and 50–80% for compost [7]. Peat has typically 90% porosity with one-fifth of the pores being less than 30 m in diameter [26]. A mixture of peat fibres and twigs having bulk density of 0.4 t m−3 is used in the meat-meal industry [1]. Peat (in particular, fibrous sphagnum peat) has an open structure providing minimum resistance to air flow. It must
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be supported on a ‘stiffening agent’ like fibrous wood chips to stop the bed clogging [18,26]. Shareefdeen et al. [28] used peat mixed with polyurethane foam, vermiculite and perlite for biofiltration of methanol vapour because of the large surface area provided for microbial adhesion and minimal pressure drop. In a review of waste treatment operations utilising peat, Martin [29] identifies high surface area (>200 m2 g−1 ), porosity, adsorptive capacity and ability to support biofilm formation as among the attributes of peat. It also has high potential nutrient value, comprising 90% carbohydrates and 5% minerals. Course materials such as bark chips, inert packings and perlite can be used to increase the porosity of an organic medium. Ergas et al. [10] used a filter media consisting of air-dried compost, perlite and crushed oyster shell. The compost consisted of 50% digested sewage sludge and 50% forest products. Perlite increased the porosity of the bed and the oyster shell provided calcium carbonate as a pH buffer. Bark compost and wood chips were found to be suitable for cleaning the exhaust air above 40◦ C emitted from a hardboard production plant down to 10 mg organic carbon per m3 . The compost could tolerate 40% higher throughputs of gas than wood chips but suffered a greater pressure loss [3]. 4.4. Pressure drop Gas phase pressure drop through a biofilter bed increases with gas flowrate and diminishing particle size. Yang and Allen [9] report increase in pressure drop of (0–35 kPa m−1 ) which are reasonably linear with increasing gas velocity (0–0.3 m s−1 ) over a range of bed medium particle sizes (1–12 mm). However, pressure drop appears to increase exponentially with decreasing particle size, especially with particles less than 1 mm. Brennan et al. [13] found pressure drop to increase significantly across a fibrous mixture compared with peat nodules and granules. Compaction of the filter bed over extended periods of usage and due to over watering will also give rise to prohibitive pressure drops [18]. Pressure drop increased from less than 500 to greater than 2500 Pa after 3 months of continuous operation of a biofilter treating 40◦ C exhaust from a hardboard plant. After a 25 day pause in operation, the pressure drop was reduced to less than 1000 Pa [3]. Ergas et al. [10] report pressure drops of 100–600 Pa m−1 at corresponding superficial air velocities through the biofilter of 0.3–1.8 m3 m−2 min−1 . These pressure drops were achieved by addition of 50% by volume of perlite to the filter media. In their review of operational biofilters, Williams and Miller [8] note the unpredictability of pressure drop across differing bed media and recommend pilot testing of individual media. Monitoring of the pressure drop across the bed was considered important for detection of cracks in the media and resultant short-circuiting of the bed by the air stream. Any increase in pressure drop adds to the operating cost of the biofilter as odorous air must be supplied at a greater pressure to achieve the same flowrate.
4.5. Surface area Specific surface areas for soils and compost range from 1–100 m2 g−1 [7]. Tang et al. [14] estimated the surface area of their sieved compost and chaff medium to be 180 m2 m−3 . Peat has even greater surface areas which result in high adsorptive capacity [29,30]. A high surface area provides access for biological growth and adsorption which are the two mechanisms by which depletion of odour from the gas phase may occur. Adsorptive capacity in the organic medium allows the biofilter to withstand fluctuations in loading rates without compromising removal rate. Higher than average loads of odorous compounds are accommodated on the solid surface. Conversely, micro-organisms can gain their carbon and energy requirements during lean loading periods from compounds that are then desorbed from the surface. Many organic materials utilised for biofiltration have a large capacity for adsorption of odorous compounds. Yang and Allen [27] consistently removed greater than 99.5% of the hydrogen sulfide supplied to their compost biofilter when loading rates were varied from 2 to 101 g S m−3 h−1 . 4.6. pH Biological metabolism is strongly dependent on pH. Many micro-organisms will only grow within a particular pH range. As a rough rule of thumb, most biological growth occurs near a neutral pH and wide deviations from this will impair the efficiency of the biofilter. A notable exception is the sulfur oxidising bacteria which thrive at low pHs. Biological oxidation more often than not results in a net increase in acidity as protons are exchanged for electrons. This includes nitric and nitrous acids as products of ammonia degradation, sulfuric acid as a product of sulfide degradation and carbon dioxide as a product of oxidation of organic compounds. Input parameters for the biofilter model of Hodge and Devinny [22] for removal of ethanol vapour in air include adsorptive capacity, porosity and buffer capacity of the filter media which enabled comparison of the effectiveness of various media. Insufficient buffering capacity as well as carbon dioxide evolution yielded a pH gradient in one of their biofilters. Yang and Allen [9] found that H2 S removal efficiency in their biofilter decreased markedly at pH below 3.2 but was almost independent of pH at higher values. They suggest that the dominant active species present were acidophiles which prefer an optimum pH near 3. The pH at the inlet to their biofilter decreased from 8.0 to 2.5 after 32 days operation. The pH was higher at the end of the biofilter as biological activity decreases through the biofilter with concentration of hydrogen sulfide [27]. Brennan et al. [13] also recorded decreases in pH from 6.5–7.0 to 3.6–4.8 after 3 weeks in a bank of biofilters exposed to hydrogen sulfide and methyl mercaptan. After 6 months, the pH in untreated biofilters fell below 2. Similarly, Kowal et al. [11] observed a decrease
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in pH from 6.5 to 3.5 after 60 days of operation of their activated sludge biofilter. Crushed oyster shells are used extensively as a source of alkalinity for buffering and as a carbon source for the autotrophic bacteria responsible for nitrification and sulfide oxidation [10,19]. Maërl, a mineral support of marine origin containing 82% calcium carbonate, was found by Bonnin et al. [20] to increase hydrogen sulfide removal in biofilters at sewage treatment plants from 10 g m−3 h−1 (peat without Maërl) to 50 g m−3 h−1 . 4.7. Conditioning Peat suffers from its low natural pH which makes establishment of many biological populations difficult without some form of pretreatment. Peat, compost and soils are all characterised by the presence of humic acids which contribute to ion exchange capacity on the solid surface. Pretreatment or conditioning of biofilter media may enhance operation. This may involve adjusting the pH of an acidic peat to encourage the growth of nitrifying bacteria, adding alkalinity such as carbonate as a carbon source for autotrophic bacteria or stabilising compost so that it is at equilibrium before adding it to a biofilter. Yang and Allen [9] noticed no significant differences in operating characteristics as a result of varying compost storage times before using it as a bed material in a biofilter column.
5. Air phase 5.1. Flowrate Gas phase residence times in biofilters can range from minutes for alcohols to 3 h for 90% removal of trichloroethylene [7]. Examples include 15 s for removal of meat-meal industry odours [1] and 23 s for 99% removal of hydrogen sulfide [9]. Maximum removal in a biofilter is obtained when the gas flowrate is as high as possible such that the maximum removal capacity R of the biofilter is not exceeded. A further constraint is imposed if mass transfer of odorous compounds from the gas phase to the biofilm is limiting. Yang and Allen [9] found that for gas phase residence times less than 23 s, their biofilter removal rates suffered from resistance to transfer of hydrogen sulfide from the gas phase to the biofilm. Superficial gas phase velocities are typically 30–200 m h−1 (50–300 cm min−1 ) in odour removing biofilters. Ottengraf and Van den Oever [25] supplied toluene from 0.3 to 5.6 g m−3 to their 3 m peat moss biofilter column. Their inlet superficial gas velocities varied correspondingly from 1 to 15 cm s−1 (30–530 m h−1 ). A superficial gas velocity of 75 m h−1 was suitable for degradation of 40◦ C hard board plant exhaust [3]. Ergas et al. [10] performed a series of pilot scale experiments removing chlorinated aliphatic compounds from air and then tested their biofilter on site
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at a water treatment plant, monitoring odorous compound removal, including aromatic VOC. Pilot scale studies involved air flowrates ranging from 0.5 to 2 m3 m−2 min−1 (30 to 120 m h−1 ) while the water treatment plant air was fed at 0.3 m3 m−2 min−1 (18 m h−1 ) to one bed and 1 m3 m−2 min−1 (60 m h−1 ) to the other bed. 5.2. Concentration Concentration loads of 0–500 ppm are common in biofilter applications [7]. When first-order kinetics are in force in a biofilter, the removal rate 8 is independent of concentration of odorous compounds in the gas inlet. This will be the case as long as the inlet concentration is low enough that the kinetics of biological degradation do not become saturated. Yang and Allen [9] were able to consistently remove greater than 99% of hydrogen sulfide from the gas phase inlet to their biofilter in the range 5–2650 ppm while operating below the maximum removal capacity. In their review of operational biofilters, Williams and Miller [8] recommend that at least 100 parts of oxygen should be provided in the air stream for every part of oxidisable odorous gas. Because odours are typically dilute, this condition is usually met. 5.3. Humidity If the biofilter bed is not watered, odorous air should be prehumidified before entering to prevent drying of the organic bed. Koch et al. [1] describe how off gases to biofilters in the meat-meal industry must be scrubbed with water to remove entrained fat and dust that will otherwise clog the biofilter bed. The scrubber also humidifies the air to greater than 95% so that additional sprinkling of the bed to keep it moist is not necessary [26]. Ergas et al. [10] performed a series of pilot scale experiments removing chlorinated aliphatic compounds from air and then tested their biofilter on site at a water treatment plant, monitoring odorous compound removal, including aromatic VOC. Humidification of inlet air was performed for the pilot-scale studies but was unnecessary at the water treatment plant as the air was already saturated. Moisture content in the filter beds was supplemented by soaker hoses at the top of the beds.
6. Liquid phase A liquid phase flowing through the biofilter bed may serve a number of purposes. It may maintain an adequate moisture content, dilute toxic metabolic products, provide additional nutrients for biological growth and provide buffering for pH stability. Dilution has been found to be particularly important for hydrogen sulfide biofiltration. Brennan et al. [13] stress the importance of a watering system for dilution and maintaining pH levels within an optimal range for biofilters
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removing reduced sulfur compounds. While sulfide removing bacteria may thrive at lower pH, other bacteria responsible for degrading different odorous compounds generally do not. Hence, a high sulfide removal rate in a biofilter does not necessarily imply a high overall odour removal. Yang and Allen [27] washed the compost in their hydrogen sulfide removing biofilter periodically with distilled water in order to dilute H2 SO4 which accumulated in the bed. This was found to potentially extend the life of the biofilter. This was also the experience of Bonnin et al. [20] who supplied water at a flowrate sufficient to dilute sulfate in their hydrogen sulfide removing biofilter such that calcium sulfate did not precipitate (solubility 3 g l−1 ). Required superficial velocities for a liquid phase are generally very low compared with those for the gas phase due to the slow rates of bed drying and biological growth.
7. Temperature Low operating temperatures will enhance sorption of odorous compounds into the biofilm but will slow down the microbial growth. Higher temperatures will have the reverse effect. High performance for most odour removing applications occurs within a temperature range of 25–40◦ C with an optimum occurring around 37◦ C [7]. Bed temperature is primarily dictated by input gas temperature as heat liberated by biological oxidation is negligible. Knauf and Zimmer [3] found that removal efficiency for organics in hot exhaust from a hardboard plant decreased steadily as temperature of the exhaust increased. For bark compost, removal efficiency dropped from 95 to 85% as temperature rose from 40 to 55◦ C. For wood chips, efficiency dropped from 80 to 70% as temperature rose from 35 to 50◦ C. Hydrogen sulfide oxidising bacteria present in the biofilter of Yang and Allen [9] were most active in the temperature range 25–50◦ C. Activity, measured as hydrogen sulfide removal efficiency at constant loading, dropped markedly on either side of this temperature range. Brennan et al. [13] found that removal rates in their hydrogen sulfide and methyl mercaptan removing biofilter decreased by over 50% when ambient temperature decreased from 20–22◦ C to 9–12◦ C. Biofiltration can be effective at low temperatures. Pinnette et al. [18] found that, once a biological population was established in their biofilter, odour removal above sludge composting facilities was not compromised at temperatures below 10◦ C.
8. Microbial population Bohn [7] estimates that microbiological populations in biofilters are of the order of one billion micro-organisms per gram of organic material. Quite often, an appropriate biological population for degradation of a particular odour
can be found near the source of the odour. For example, micro-organisms capable of degrading VOC from oil refining operations have been found in soil adjacent to tanks where overflow and spillage were common. Because of their relatively fast population doubling times, bacteria may evolve within a few years to grow more efficiently on a predominant nutrient. 8.1. Acclimatisation Acclimatisation periods are often required for microorganisms to adapt to new conditions imposed by a new biofilter bed medium. These will include biological nutrients, pH, temperature and solid surface characteristics. Because of their ability to evolve, micro-organisms often improve their odour removing ability through the life of the biofilter. In their review of operational biofilters, Williams and Miller [8] reported acclimatisation periods of greater than 10 days. Even longer times were recorded by Wang et al. [4] in their investigation of biological water filters. They found that biomass accumulation was a slow process, taking up to 5 months for biomass to reach an equilibrium concentration. Biomass was measured via a phospholipid extraction technique that showed a decreasing biomass concentration with depth in the bed. This indicates that most biological degradation takes place at the front of the bed where initial contact with the waste fluid takes place. Further, it was found that the capacity of the bed to remove organic material did not always increase with biomass concentration. It was suggested that above a critical biomass level, the removal capacity of the filter was somewhat independent of biomass concentration. Shareefdeen et al. [28] conditioned their methanolremoving biofilter with cultures of bacteria that had been exposed to methanol vapour. Agricultural soil was placed in an enclosure with the vapour and, after a week, soil samples were placed in a mineral medium with 1% methanol and sufficient airspace to prevent oxygen limitation. It was found that a consortium of eight methanol-utilising bacterial strains was found to grow consistently at higher rates than any of its individual members. Ergas et al. [10] inoculated their odour removing biofilter with a compost containing 50% digested sewage sludge and 50% forest products. Mixed liquor suspended solids from a waste water treatment plant were also added at 7 l per m3 of filter material, followed by a Pseudomonas putida strain at 106 cells per gram of filter medium. Brennan et al. [13] inoculated hydrogen sulfide and methyl mercaptan removing biofilter with cultures of Thiosphaera pantotropha and Thiobacillus neapolitanus. They found that these inoculated biofilters provided significant improvement in removal over uninoculated biofilters. Yang and Allen [9] found that an acclimatisation period was required to reach maximum removal efficiency for H2 S. This was presumably required for the microbial population
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to build up to its optimum size after exposure to H2 S. The time required was up to 2 weeks. They did not inoculate with sulfide oxidising bacteria or previously expose the medium to hydrogen sulfide. Contrary to Brennan et al. [13], they concluded that the bacteria are ubiquitous in soils and compost. They found no advantage from adding sewage sludge to their media. 8.2. Inhibition It is common for nutrients required for biological growth and biological end products to inhibit growth in significant concentrations. As such, biofiltration is well suited to odour removal as concentrations of odorous compounds are typically dilute. Pinnette et al. [18] concluded that high levels of ammonium in their biofilter bed contributed to poor odour removing performance during start up due to inhibition of biological populations by ammonium. Yang and Allen [9] found that sulfate concentrations of more than 25 mg S (g dry compost) inhibited the performance of their hydrogen sulfide removing biofilter. This was significant because sulfate is a product of the biological oxidation of sulfide. Elemental sulfur also accumulated at the entrance to the biofilter at high loading rates [27]. 8.3. Kinetics It is difficult to extract consistent quantitative biological rate constants for biofiltration of odours due to the widely varying experimental data reported. Biological rates are very dependent on biofilter bed medium, pH, alkalinity and operating temperature, none of which are likely to be identical in any pair of experiments. Most of the available data pertains to biofiltration of hydrogen sulfide and a very thorough investigation of this by Yang and Allen [27] is analysed here. They determined values for µ0 (26.1 ppmv s−1 =131 g m−3 h−1 ) and µ1 (0.54 s−1 =1900 h−1 ) in their hydrogen sulfide removing biofilter. Thus, at high inlet gas phase concentrations of hydrogen sulfide dCg µ0 εA Cg µ0 1 =− ≈ = −131 dz Fg K + Cg ug ug where ug (m h−1 ) is the gas phase interstitial velocity given by ug =
Fg εA
(19)
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Hence the kinetics for hydrogen sulfide degradation are given by r = 131 ×
Cg 67 + Cg
where the units of Cg are mg m−3 . These results should be treated with some caution as µ1 was determined for Cg,in <280 mg m−3 . According to their results, first-order kinetics would only be expected for Cg K=67 mg m−3 . However, their values for µ0 (131 g m−3 h−1 ) and, by implication, K (67 mg m−3 ) are indicative and provide insight into when first- and zeroth-order kinetics might be expected in a hydrogen sulfide removing biofilter. It is difficult to determine biological rate constants for biofiltration of ammonia due to the limited success reported in the literature. In biological nutrient removal from wastewater, the International Association on Water Quality recommends K=1.3 g m−3 (1300 mg m−3 ) for degradation in the liquid phase. Given that, at 25◦ C, the Henry’s law coefficient for ammonia in water is approximately 0.0037 [31], then this liquid phase concentration corresponds to K=4.8 mg m−3 in air. It is even less meaningful to quantify rate constants for VOC and ODU biofiltration due to the vast numbers of individual odorous compounds involved.
9. Conclusions It is difficult to compare the performances of biofilters in the literature when it is not clear whether they are operating subject to zeroth- or first-order kinetics of contaminant degradation. An absolute maximum elimination rate can only be determined by measuring the removal rate of contaminants as first flowrate and then concentration are increased. A qualitative kinetic analysis of biological degradation in biofilters suggests that the mean gas phase residence time in a biofilter should be large enough for mass transfer of contaminants from gas to biofilm to occur and for the biological kinetics to remain unsaturated. If the residence time is too large, however, then the maximum elimination capacity of the biofilter will not be attained. Besides kinetics, attention must be paid to biofilter bed media to optimise performance. Parameters such as moisture content, porosity, surface area and pH will affect adsorption and biological degradation of contaminants.
At low inlet gas phase concentrations of hydrogen sulfide Cg dCg µ0 εA Cg µ0 =− Cg = 1900 ≈ dz Fg K + Cg ug K ug It is then possible to calculate a value for K according to 131 µ0 = = 1900 h−1 K K
and
K = 67 mg m−3
References [1] W. Koch, H.-G. Liebe, B. Striefler, Operating experience with biofilters to reduce odorous emissions, Staub-Reinhaltung Der Luft 42 (12) (1982) 488–493. [2] M. Walters, Love thy neighbour, PACE: Process Chem. Eng. 48 (1) (1995) 18.
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[3] S. Knauf, H. Zimmer, Biofiltration at a temperature above 40◦ C: comparison of the biofilter materials bark compost and woodchips, Staub-Reinhaltung der Luft 54 (2) (1994) 41–44. [4] J.Z. Wang, R.S. Summers, R.J. Miltner, Biofiltration performance. Part 1. Relationship to biomass, J. Am. Water Works Assoc. 87 (12) (1995) 64–70. [5] S.P.P. Ottengraf, Biological systems for waste gas elimination, Trends Biotechnol. 5 (5) (1987) 132–136. [6] G. Leson, A.M. Winer, Biofiltration: an innovative air pollution control technology for VOC emissions, J. Air Waste Manage. Assoc. 41 (8) (1991) 1045–1054. [7] H. Bohn, Consider biofiltration for decontaminating gases, Chem. Eng. Progress 88 (4) (1992) 34–40. [8] T.O. Williams, F.C. Miller, Biofilters and facility operations, BioCycle 33 (11) (1992) 75–79. [9] Y. Yang, E.R. Allen, Biofiltration control of hydrogen sulfide: 1. Design and operational parameters, J. Air Waste Manage. 44 (1994) 863–868. [10] S.J. Ergas, E.D. Schroeder, D.P.Y. Chang, R.L. Morton, Control of volatile organic compound emissions using a compost biofilter, Water Environ. Res. 67 (5) (1995) 816–821. [11] S. Kowal, J.L. Fanlo, J.R. Degorce-Dumas, P. Le Cloirec, Removal of H2 S by a biofilter with dry activated sludge: an approach of mechanisms, in: Proceedings of the International Symposium on Environmental Biotechnology Oostende, Royal Flemish Society of Engineers, 1991, pp. 135–138. [12] J. Monod, The growth of bacterial cultures, Annu. Rev. Microbiol. 3 (1949) 371–394. [13] B.M. Brennan, M. Donlon, E. Bolton, Peat biofiltration as an odour control technology for sulphur-based odours, J. Chartered Institution Water Environmen. Manage. 10 (3) (1996) 190–198. [14] H-M. Tang, S-J. Hwang, S-C. Hwang, Waste gas treatment in biofilters, J. Air Waste Manage. Assoc. 46 (4) (1996) 349– 354. [15] J.C. Van Den Heuvel, Mass transfer in and around biofilms, in: L.F. Melo et al. (Eds.), Biofilms: Science and Technology, Kluwer Academic Publishers, Dordrecht, 1992, pp. 239–250. [16] S.J. Ergas, K. Kinney, M.E. Fuller, K.M. Scow, Characterization of a compost biofiltration system degrading dichloromethane, Biotechnol. Bioeng. 44 (1994) 1048–1054. [17] S.M. Zarook, A.A. Shaikh, Analysis and comparison of biofilter models, Chem. Eng. J. 65 (1997) 55–61.
[18] J.R. Pinnette, M.D. Giggey, G.J. Marcy, M.A. O’Brien, Performance of biofilters at two agitated bin composting facilities, in: Proceedings of the 87th annual meeting of the Air and Waste Management Association, Ohio, Air and Waste Management Association, 1994. [19] R. Kapahi, M. Gross, Biofiltration for VOC and ammonia emissions control, BioCycle 36 (2) (1995) 87–90. [20] C. Bonnin, G. Martin, G. Gragnic, Biopurification of odorous gases in urban wastewater treatment plants, in: S. Vigneron, J. Hermia, J. Chaouki (Eds.), Characterization and Control of Odours and VOC in the Process Industries, Proceedings of the Second International Symposium on Characterization and Control of Odours and VOC in the Process Industries, Louvain-la-Neuve, Belgium, 3–5 November 1993, Elsevier, Amsterdam, 1994, pp. 431–444. [21] A.P. Togna, M. Singh, Biological vapor-phase treatment using biofilter and biotrickling filter reactors: practical operating regimes, Environ. Progress 13 (2) (1994) 94–97. [22] D.S. Hodge, J.S. Devinny, Modelling removal of air contaminants by biofiltration, J. Environ. Eng. 121 (1) (1995) 21–32. [23] M.A. Deshusses, G. Hamer, I.J. Dunn, Behaviour of biofilters for waste air biotreatment. 2. Experimental evaluation of a dynamic model, Environ. Sci. Technol. 29 (4) (1995) 1059–1068. [24] M. Rothenbühler, M. Heitz, M. Beerli, B. Marcos, Biofiltration of volatile organic emissions in reference to flexographic printing processes, Water Air Soil Pollut. 83 (1995) 37–50. [25] S.P.P. Ottengraf, A.H.C. Van Den Oever, Kinetics of organic compound removal from waste gases with a biological filter, Biotechnol. Bioeng. 25 (1983) 3089–3102. [26] F.H.H. Valentin, Peat beds for odour control: recent developments and practical details, Filtration Sep. 23 (4) (1986) 224–227. [27] Y. Yang, E.R. Allen, Biofiltration control of hydrogen sulfide. 2. Kinetics biofilter performance and maintenance, J. Air Waste Manage. Assoc. 44 (11) (1994) 1315–1321. [28] Z. Shareefdeen, B.C. Baltzis, Y.S. Oh, R. Bartha, Biofiltration of methanol vapor, Biotechnol. Bioeng. 41 (1993) 512–524. [29] A.M. Martin, Biofilms in peat biofilters, in: L.F. Melo et al. (Eds.), Biofilms: Science and Technology, Kluwer Academic Publishers, Dordrecht, 1992, pp. 461–466. [30] D. Mcnevin, J. Barford, J. Hage, Adsorption and biological degradation of ammonium and sulfide on peat, Water Res. 33 (6) (1999) 1449–1459. [31] E. Wilhelm, R. Battino, R.J. Wilcock, Low-pressure solubility of gases in liquid water, Chem. Rev. 77 (2) (1977) 219–230.
Biofiltration as an odour abatement strategy Dennis McNevin a , John Barford b,∗ a
b
Department of Chemical Engineering, University of Sydney, Sydney, NSW 2006, Australia Department of Chemical Engineering, H.K.U.S.T., Clear Water Bay, Kowloon, Hong Kong (S.A.R.), China Received 18 December 1999; accepted 21 March 2000
Abstract The chemical, physical and biological processes occurring in biofiltration are reviewed. A survey of operating biofilter performances is also presented and includes some novel comparative methods. It is concluded that biofiltration is a simple and cost-effective technology for odour removal and that an understanding of the many interactions occuring within the biofilter is essential for the optimal performance of the biofilter. © 2000 Elsevier Science S.A. All rights reserved. Keywords: Biofiltration; Biofilter performance; Removal efficiency; Maximum load; Removal rate; Operating parameters
1. Introduction The removal of nutrients from wastewaters and odours from air is often effected through biological means in unit operations like biofilters, biotrickling filters and bioscrubbers. There are successful applications of these in the meat meal industry [1], the tobacco industry [2], hardboard production [3], wastewater treatment [4] and a number of reviews attest to their overall efficacy for odour control [5–8]. Removal of hydrogen sulfide from gas streams has been particularly successful [9,10]. A biofilter consists of a container of organic material populated with micro-organisms through which odorous air is passed, usually upwards. The influent air may be prehumidified to maintain adequate moisture in the organic bed. Alternatively, or in addition, water may be sprinkled over the surface of the bed, trickling downwards, counter-current to the odorous air. This water may contain nutrients required by the micro-organisms for growth. Liquid trickling from the bottom of the biofilter may be bled off or recycled (Fig. 1). Odorous contaminants are transferred from the air into an aqueous, bio-active layer (biofilm) surrounding organic particles in the bed. The contaminants are then aerobically degraded to various end products or incorporated into biomass. The end products will depend on the nature of the contaminants. Biofiltration enhances the natural processes of bioremediation where contaminants in the atmosphere ∗ Corresponding author. Tel.: +852-2358-7237; fax: +852-2358-0054. E-mail address: [email protected] (J. Barford)
are degraded by micro-organisms in soils after diffusion into soil pores. A well engineered biofilter provides improved contact between organic particles and air containing contaminants. Compost and soil pores are organically coated which enhances sorption of organic contaminants while the aqueous film surrounding these particles facilitates the sorption of inorganics. The distinction between the organic coating and the aqueous film is unclear and they are often collectively called the ‘biofilm’. Biofilters, biotrickling filters and bioscrubbers may be generically referred to as organic perfusion columns. They consist of three phases in intimate contact: a solid, organic phase, a liquid phase and a gas phase. All three may contain nutrients for degradation. The liquid phase (water or nutrient solution) and gas phase are passed through the solid organic medium, invoking the processes of adsorption and aerobic biological degradation of the nutrients in the liquid. These nutrients may have been transferred from the air, as is the case with odour removal. The relative extent to which adsorption or biological degradation dominates in nutrient removal may vary with the organic medium used, as well as with local conditions like pH and temperature. The interdependence of these processes is also unclear. Adsorption of solutes onto the surface of peat may benefit or hinder their subsequent degradation by micro-organisms. Some biological conversions important in biofiltration of odorous compounds, together with typical bacteria responsible for them, are shown in Table 1. Kowal et al. [11] provide evidence to suggest that removal of hydrogen sulfide in a dry activated sludge biofilter occurs by way of three distinct phases: (i)
1369-703X/00/$ – see front matter © 2000 Elsevier Science S.A. All rights reserved. PII: S 1 3 6 9 - 7 0 3 X ( 0 0 ) 0 0 0 6 4 - 4
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where Cg,in and Cg,out (g m−3 ) are the concentrations of contaminant in the inlet and outlet gas streams for the biofilter, respectively. It has been found that: 8 = 1 − exp(−µ1 τ )
Fig. 1. Generalised schematic diagram of a biofilter for odour removal.
absorption into the water present in the organic bed; (ii) adsorption onto the solid surface and biodegradation. The process denoted by (i) above is well described by standard thermodynamic vapour–liquid equilibria and is fast enough (relative to adsorption and biodegradation) to be considered instantaneous (always at equilibrium), especially when air phase residence times are typically of the order of minutes in biofilters. Either mass transfer resistance to adsorption or biodegradation will be the limiting process before equilibrium is achieved in a biofilter. The process of biodegradation is usually described by the Monod relationship, which relates the biodegradation rate to the substrate concentration Cg Rate = Maximum rate × K + Cg where Cg is the gas phase substrate concentration and K is the saturation constant. Depending on the concentration of substrate, the rate can be zero order (high concentration) or first order (low concentration).
(2)
where τ (h) is the mean residence time for the air phase in the biofilter and µ1 (h−1 ) is a rate constant, thought to be mostly a function of contaminant biodegradability. This implies that the microbial degradation rate is often first order with respect to contaminant concentration in the gas phase, as a first approximation. It also suggests that the most important design parameter for biofiltration is the biological degradation rate and biofilters have traditionally been designed thus. However, on careful consideration of the processes involved in biofiltration, mass transfer and maximum removal rate are also important. Consider a biofilter with a population of micro-organisms capable of degrading a contaminant with rate r (g m−3 h−1 ), subject to first-order kinetics according to r = µ 1 Cg
(3)
where Cg (g m−3 ) is the gas phase concentration of contaminant anywhere in the biofilter. At steady state, a mass balance on the contaminant over an incremental length, dz (m), of the biofilter yields: Fg Cg |z − Fg Cg |z+dz = µ1 Cg εA dz
(4)
where Fg (m3 h−1 ) is the gas phase flowrate, A (m2 ) the biofilter cross-sectional area and ε (m3 m−3 ) represents the voidage occupied by the gas phase in the biofilter. Collecting terms and integrating over the total length, Z (m) of the biofilter yields Cg,out εAZ = exp −µ1 (5) − exp(−µ1 τ ) Cg,in Fg where
2. Biofilter performance measures
τ=
εAZ Fg
(6)
2.1. Removal efficiency 2.2. Maximum load It is common to describe the performance of a biofilter in terms of removal efficiency (8) defined as [7] 8=1−
Cg,out Cg,in
(1)
The parameter 8 does not indicate how much contaminant is being removed in absolute terms. A 99% removal from an air stream with 10 ppm ammonia can not be distinguished from 99% removal of a 1000 ppm stream when
Table 1 Biological conversions important in biofiltration of odorous compounds Transformation
Typical bacteria
Oxygen environment
Organic carbon oxidation VOC →CO2 , H2 O Nitrification NH4 + →NO2 − , NO3 − Sulfide oxidation H2 S→S0 , SO4 2− Denitrification NO3 − →N2
Chemoheterotrophic bacteria Nitrifying bacteria Sulfur oxidising bacteria Denitrifying bacteria
Aerobic Aerobic Aerobic Anaerobic
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obviously the latter is removing a great deal more at the same flowrate. With increasing concentration in the inlet, the rate of biological degradation will eventually cease to be first order and become zeroth order (i.e. the kinetics will become ‘saturated’ or independent of concentration). Hence, a maximum load, L (g m−3 ) must be identified, above which 8 will decrease. At this maximum load, a mass balance on the contaminant over an incremental length of the biofilter yields
Table 2 Normalised variables used for simple biofilter model
Fg Cg |z − Fg Cg |z+dz = µ0 εA dz
C dC = − dz 1+C
(7)
where µ0 (g m−3 h−1 ) is a zeroth order biological degradation rate constant. Collecting terms and integrating over the length of the biofilter yields: L = Cg,in − Cg,out
εAZ = µ0 = µ0 τ Fg
(8)
The removal efficiency 8 now depends on the inlet flow concentration: 8=1−
Cg,out µ0 τ = Cg,in Cg,in
(9)
2.3. Removal rate A biofilter removal or elimination rate R (grams of contaminant removed per kg bed medium per hour), can be defined for an individual biofilter. This is the rate of removal of contaminant per unit mass of biofilter bed medium and is given by: R=
Fg µ0 τ Fg L µ0 εAZ = = m m m
(10)
where m (kg) is the mass of bed medium in the biofilter. R may also be defined on a unit volume basis (grams of contaminant removed per m3 bed medium per hour) or on a unit surface area basis (grams contaminant removed per m2 biofilter cross-sectional area per hour).
Normalised variable
Definition
Significance
C
Cg /K
z
z/z
Gas phase concentration normalised with respect to the half saturation constant Distance through the biofilter normalised with respect to the total height
where µ0 τ = K
(13)
(14)
Dimensionless groups are introduced to enable the comparison of different biofilters of different scale and operating conditions. The dimensionless parameter indicates the relative effectiveness of the biofilter. The dimensionless parameter (Cin −Cout )/8 indicates the relative removal rate. A high value of means that the biological degradation rate and/or the biofilter volume is high compared with the gas flowrate and hence removal is enhanced. Eq. (13) represents a macroscopic model of biofiltration and is the result of a steady-state mass balance on gas phase contaminants. It assumes a constant ratio of gas phase concentration to solid/liquid phase concentration. In reality, this ratio will vary with temperature and pH. The model is satisfactory as a first approximation but does not take account of mass transfer effects which are considered later. Eq. (13) was integrated numerically over a range of values for and Cin from Cin at z=0 to Cout at z=1 (Figs. 2–4). At low inlet concentrations, relative to the half saturation constant (Cin 1), a characteristic first-order concentration profile is observed through the bed (Fig. 2). At high concentrations (Cin 1), the profile is characteristically linear (Fig. 4). It is then possible to plot the dimensionless loading rate, Cin /, against the dimensionless removal rate
2.4. Effect of loading rate on removal rate The transition from first to zeroth order biological degradation kinetics due to increased load can be represented by the well known Monod [12] expression such that r = µ0
Cg K + Cg
(11)
where K (g m−3 ) is the so-called half saturation constant. When Cg
(12)
Variables can be normalised as in Table 2 which then yields
Fig. 2. Concentration profiles in a biofilter bed for a low inlet concentration relative to the half-saturation constant (Cin =0.1; predominantly first-order kinetics).
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to cope with the excess loading. However, increasing the inlet concentration will allow the biofilter to provide greater maximum elimination rates until a limit is reached here, too. This is due to the effect of high concentrations (C K or C 1) on the Monod kinetics of degradation when the kinetics become saturated. The absolute maximum elimination rate, obtained at high gas flowrates and high inlet concentrations, is given by Cg,in − Cg,out Cin − Cout = =1 µ0 τ Fig. 3. Concentration profiles in a biofilter bed for an intermediate inlet concentration relative to the half-saturation constant (Cin =1; zeroth- and first-order kinetics).
(15)
for which the conditions according to Fig. 5, assuming constant biological parameters, µ0 and K, are Cg,in Cin = >1 µ0 τ
(16)
and Cin > 100(Cg,in > 100 K)
Fig. 4. Concentration profiles in a biofilter bed for a high inlet concentration relative to the half-saturation constant (Cin =10; predominantly zeroth-order kinetics).
(Cin −Cout )/ (Fig. 5). As the loading rate increases due to increased flowrate of gas, the elimination rate increases until a maximum (saturated) elimination rate is achieved. At this inlet concentration, further increases in gas flowrate will mean larger outlet concentrations as the biofilter is unable
Fig. 5. Dimensionless removal rate (vertical axis) vs. dimensionless loading rate (horizontal axis) for a biofilter with Monod kinetics.
Under conditions of very low Cin , and with first-order kinetics, then, (Cin −Cout )/=Cin / (1−exp(−)) can be obtained and the plot of (Cin −Cout )/−v−Cin / would be independent of Cin . Fig. 5 suggests that if first order or Monod Kinetics are in force in a biofilter, then any correlation of loading versus removal rates must supply data points that lie on or to the right of the 100% removal line (line with slope of unity passing through the origin). Obviously, the 100% removal line represents the best possible performance from a biofilter and as the biofilter becomes more overloaded, its absolute removal capacity ceases to increase. A maximum removal rate for the well studied hydrogen sulfide removing biofilters was found by Yang and Allen [9] and Brennan et al. [13]. They also reported experimental data consistent with Fig. 5. Thus, given that the kinetics of degradation are quantitatively well understood for a solid/liquid system, a biofilter can be designed to operate at a maximum removal rate and achieve a desired removal efficiency. If there is any inhibition of biological growth by the degraded contaminant then the maximum removal rate will not remain constant as loading rate increases past saturation. As the loading of contaminant increases, the removal rate will decrease from a maximum as biological growth is inhibited. This was found to be the case when Tang et al. [14] increased the loading of malodorous triethylamine past 150 g m−3 h−1 in their compost/chaff biofilter, corresponding to a maximum removal rate of 140 g m−3 h−1 . This treatment of biofilter removal capacity is based on the very limiting assumption that there is no resistance to mass transfer from the gas phase to the biofilm surrounding the solid organic particles in the bed medium. There is a risk when designing biofilters for very easily degradable contaminants that a short residence time τ will be employed which will be insufficient for mass transfer to occur. This is borne out in the results of Yang and Allen [9] where removal efficiency in their hydrogen sulfide removing biofilter
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dropped from a maximum when retention time fell below about 23 s. They were able to show that they were operating at removal rates R, well below the maximum obtainable by their biofilter. Hence, a minimum residence time for mass transfer is also required when designing gas flowrates for biofilters.
2.5. Thiele number A common method of identifying the efficiency of a biofilter is by quoting a dimensionless Thiele number φ which indicates the relative importance of biological degradation versus diffusion/mass transport with respect to the elimination of contaminants in the biofilm. For first-order degradation kinetics, the Thiele number is defined as [15–17]: s φ=
µ1 δ 2 D
(18)
where δ (m) is the biofilm thickness and D (m2 h−1 ) is the effective diffusion coefficient for the contaminant in the biofilm. Zeroth order Thiele numbers are discussed. The biofilm will become diffusion limited as φ increases. A diffusion limited biofilm requires a longer air phase residence time than suggested by consideration of biological degradation rates alone. Hence, for efficient operation of a biofilter, the air phase residence time τ should be adjusted such that it is long enough so that mass transfer of contaminants from air phase to biofilm can occur and that the kinetics of biological degradation are nearly ‘saturated’. However, it should not be much longer than this if a maximum removal rate is desired. Diluting an odorous air stream by two and doubling the flowrate of the air to a biofilter will not necessarily give the same removal rate R or the same removal efficiency 8. The maximum removal rate will only be preserved during load fluctuations if the kinetics of degradation are predominantly zeroth order or linear (Cg K). However, in this regime, the removal efficiency will suffer with increased flowrate. Conversely, at very low concentrations (Cg K), when the kinetics of degradation are first order, the removal efficiency will be independent of the inlet concentration but the removal rate will increase with concentration.
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3. Survey of biofilter performances The three parameters, 8, L and R, can be used to compare the performance of biofilters. Reported concentrations of odorous compounds are usually expressed on a volume basis, ppmv (parts per million by volume), or on a mass basis, Cg ppmm (mg m−3 ). In order to compare the two, it is necessary Cg to know the temperature of the air carrying the odorous compounds. At 25◦ C 1 M ppmm ppmv ppmv =Cg × × M × 1000 = Cg Cg 24465 24.465 where M (g mol−1 ) is the molecular mass of the compound. The removal rate R can also be expressed on a mass basis (g g−1 h−1 ), a volume basis (g m−3 h−1 ) or a biofilter cross-sectional area basis (g m−3 h−1 ). Comparison between these three bases requires knowledge of the packing density of the filter medium or the biofilter bed depth which are not always reported. Performances surveyed here have been divided into those pertaining to ammonia, hydrogen sulfide and VOC removal, according to the classification of odorous compounds elucidated earlier. These three compounds form the bulk of the work reported in the literature. In addition, performance based on olfactometry measurements has also been included. 3.1. Ammonia removal The performances of several ammonia removing biofilters found in the literature are presented in Table 3. Reported research of biofiltration of ammonia is limited, perhaps because nitrifying bacteria responsible for the oxidation of ammonia exhibit slow growth rates. Biological ammonia removal may be disguised by absorption in moisture present in the biofilter and adsorption onto the surface of the organic filter material. Any leachate removed from a biofilter exposed to ammonia in the inlet gas will contain significant ammonium [18]. 3.2. Hydrogen sulfide removal The performances of several hydrogen sulfide removing biofilters found in the literature are presented in Table 4. Removal rates vary wildly but removal efficiencies are always at least 99%. In a thorough examination of
Table 3 Performance of ammonia removing biofilters in the literature Authors
8
L (mg m−3 )
R (g m3 h−1 )
Notes
Pinnette et al. [18] Kapahi and Gross [19] Bonnin et al. [20]
0.964 0.983
3.0 580 1.43
1.0 10.6 0.2388a
Compost, bark mulch, wood chips (estimated from nitrate production) Compost, oyster shell and perlite blend at landfill Peat biofilter at wastewater treatment plant
a
Unit in g m2 h−1 .
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Table 4 Performance of hydrogen sulfide removing biofilters in the literature Authors
8
L (mg m−3 )
R (g m−3 h−1 )
Notes
Yang and Allen [9] Ergas et al. [10] Brennan et al. [13] Pinnette et al. [18] Bonnin et al. [20]
0.999 0.999 0.99
3690 28 83 0.0308 32
130 420 8.3 2.361 10
Compost, bench scale Compost, oyster shell and perlite Peat, bench scale Compost, bark mulch, wood chips Peat biofilter at wastewater treatment plant
0.99
Table 5 Performance of VOC removing biofilters in the literature Authors
8
L (g m−3 )
R (g m−3 h−1 )
Notes
Knauf and Zimmer [3] Ergas et al. [10] Tang et al. [14] Kapahi and Gross [19] Togna and Singh [21] Hodge and Devinny [22] Deshusses et al. [23] Rothenbühler et al. [24]
0.95 0.95 1.00 0.86 0.95 0.40
0.14 (organic C) 2.656 2 40 (ppmv)
0.90
6.1
20 0.130 140 0.314 175 29 121 80a
Exhaust air from a hardboard production plant Benzene from sewage treatment plant Triethylamine Non-methane hydrocarbons at landfill Ethanol, bench scale Ethanol, bench scale Methyl ethyl ketone Printing solvents; peat ball medium
a
Unit in g C
4.4
m−3 h−1 .
hydrogen sulfide biofiltration, Yang and Allen [9] determined a maximum H2 S elimination capacity for their biofilter by plotting H2 S removal rates (g(m3 of bed)−1 h−1 ) against H2 S loading rates (g(m3 of bed)−1 h−1 ). At loading rates greater than 150 g m−3 h−1 , the removal rate levelled off at approximately 130 g m−3 h−1 , implying that a maximum biological degradation rate had been achieved. Their biofiltration systems, containing various types of yard waste compost as the filter material, removed H2 S with efficiencies greater than 99.9% for H2 S inlet concentrations in the range 5–2650 ppmv. Theirs is the only biofilter for which this absolute maximum removal rate was determined.
requires a residence time greater than three minutes. The inlet gas loadings of aromatics and H2 S were subject to wide fluctuations showing that the biofilter was capable of handling load variations without a significant effect on removal efficiency. Chlorinated aliphatic compounds were removed in pilot-scale experiments although the concentrations were not reported. Williams and Miller [8] provide further reported removal capacities in their review of operational biofilters. 3.4. Odour removal as measured by olfactometry The performances of two odour removing biofilters found in the literature, as measured by dynamic olfactometry, are presented in Table 6.
3.3. VOC removal The performances of several VOC removing biofilters found in the literature are presented in Table 5. Ergas et al. [10] were able to consistently obtain greater than 90% removal of difficult-to-degrade aromatic compounds in field studies at a water treatment plant, corresponding to average mass removal rates of 8 mg benzene per minute in a 3 m2 ×1.2 m biofilter bed (130 mg m−3 h−1 ). Removal of benzene, toluene and xylene decreased when the mean residence time for air in the biofilter was decreased from 3 to 1 min, indicating that complete removal of aromatics
4. Operating parameters 4.1. Bed media Table 7 lists bed media types found in the literature for removal of odorous compounds via biofiltration. Typical materials used include soil, peat, compost, bark mulch, perlite, sewage sludge and combinations of these. Multiple
Table 6 Performance of odour removing biofilters in the literature as measured by dynamic olfactometry Authors
8
L (OU m−3 )
R (OU m−3 h−1 )
Notes
Pinnette et al. [18] Bonnin et al. [20]
0.91 0.96
309 3910
14832 651667a
Compost, bark mulch, wood chips above sludge composting facilities Peat biofilter at wastewater treatment plant
a
Unit in OU m−2 h−1 .
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Table 7 Types of bed media used for biofiltration in the literature Type
Qualities
Bark compost, woodchips Compost, sewage sludge Compost, perlite, crushed oyster shells Sod-peat nodules and extruded peat granules Sieved compost and chaff Compost, bark mulch, wood chips Compost, perlite, zeolite, oyster shells Maërl (marine mineral containing 82% CaCO3 )
Low pressure drop. Removes VOC Removes H2 S Removes aromatics, H2 S Low pressure drop. Removes H2 S, methyl mercaptan Removes triethylamine. Chaff reduces pressure drop. Removes odour from sludge composting facilities Removes NH3 , VOC Provides alkalinity for autotrophic biological degradation
layers of biofilter media have been recommended by some manufacturers. Bed depths are of the order of 1 m for many industrial applications [1,3]. In their review of operational biofilters, Williams and Miller [8] have observed that biofilter bed depths vary from 0.5 to 2.5 m. Greater bed depth requires less land area but results in higher pressure drops across the bed. Biofilter bed life varies from 3 months [19] to 4 years [13,18]. There are a number of parameters used to describe the quality of a particular biofilter bed material. Some qualities, like high surface area for adsorption, adequate moisture content and large voidage (and hence low pressure drop across the bed) are universally required for biofiltration applications while others, particularly pH, are dependent on the application. 4.2. Moisture content Williams and Miller [8] point to bed moisture content as the single most important parameter for biofilter viability. Optimal moisture contents varied from 20 to 60% in their review of operational biofilters. Biological activity ceases if the moisture content of an organic material is too low. In addition, cracks open in a dry bed and channelling occurs, further limiting performance [25]. Conversely, too much moisture leads to anaerobic zones forming in the bed where oxygen required for bio-oxidation is depleted. For this reason, the capacity of soil beds to remove odour drops markedly when they become too wet. In addition, mass transfer of odorous compounds to the organic surface through a sodden bed, as opposed to through a thin film of water surrounding organic particles, becomes rate limiting. Pinnette et al. [18] suffered a loss of biological degradation of odorous compounds above sludge composting facilities when moisture content dropped below approximately 40%. The biofilter recovered its previous performance within 3 months after watering of the filter bed. Yang and Allen [9] found that above a water content of 30 wt.%, there was not much effect of water content on H2 S removal capacity in their bench scale biofilter. Below 30 wt.% water, however, the removal efficiency decreased proportionately. Composts were able to recover from being dry to regain their former
pH 3.2 6.5–7.0 7
Authors Knauf and Zimmer [3] Yang and Allen [9] Ergas et al. [10] Brennan et al. [13] Tang et al. [14] Pinnette et al. [18] Kapahi & Gross [19] Bonnin et al. [20]
H2 S eliminating capacity. The time required for recovery was from 1 to 3 days. Ottengraf and Van den Oever [25] kept moisture content in their VOC-eliminating biofilter at 50–70%. At low water levels, the organic packing lost its microbial activity while high water content promoted the development of anaerobic zones in the bed. Removal of aromatic VOC and hydrogen sulfide in the odour-removing biofilter of Ergas et al. [10] increased dramatically and almost immediately when moisture content was increased from below 50 to 55%. Peat and compost have good water holding capacities. Microbial activity in peat falls off if water content drops below 70% and rises above 85%. Below about 30%, biological activity usually ceases [26]. Soil is much less permeable than compost due to smaller pore sizes. It is, however, hydrophilic where as dry compost is hydrophobic and is difficult to rewet after drying out due to its highly porous structure. Compost also cakes when lime is added [7]. Heat generated by biological activity in a biofilter may increase the temperature of the bed medium above that of the inlet gas phase. Even if this gas enters the biofilter saturated with water, it will become unsaturated as its temperature rises after contact with the bed medium. Drying of the bed medium will inevitably occur. This will be offset to some degree by biological degradations that produce water (e.g. biological oxidation of VOC) [27]. Hence it is important to supply 100% humid air to a biofilter and/or water the bed to replace moisture lost to the gas phase if the organic bed is to remain viable. 4.3. Porosity Porosity, voidage and bulk density essentially relate to the same property of organic bed media and are important primarily for the effect they have on the gas phase pressure drop across the bed. Porosities for organic media range from 40 to 50% for soils and 50–80% for compost [7]. Peat has typically 90% porosity with one-fifth of the pores being less than 30 m in diameter [26]. A mixture of peat fibres and twigs having bulk density of 0.4 t m−3 is used in the meat-meal industry [1]. Peat (in particular, fibrous sphagnum peat) has an open structure providing minimum resistance to air flow. It must
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be supported on a ‘stiffening agent’ like fibrous wood chips to stop the bed clogging [18,26]. Shareefdeen et al. [28] used peat mixed with polyurethane foam, vermiculite and perlite for biofiltration of methanol vapour because of the large surface area provided for microbial adhesion and minimal pressure drop. In a review of waste treatment operations utilising peat, Martin [29] identifies high surface area (>200 m2 g−1 ), porosity, adsorptive capacity and ability to support biofilm formation as among the attributes of peat. It also has high potential nutrient value, comprising 90% carbohydrates and 5% minerals. Course materials such as bark chips, inert packings and perlite can be used to increase the porosity of an organic medium. Ergas et al. [10] used a filter media consisting of air-dried compost, perlite and crushed oyster shell. The compost consisted of 50% digested sewage sludge and 50% forest products. Perlite increased the porosity of the bed and the oyster shell provided calcium carbonate as a pH buffer. Bark compost and wood chips were found to be suitable for cleaning the exhaust air above 40◦ C emitted from a hardboard production plant down to 10 mg organic carbon per m3 . The compost could tolerate 40% higher throughputs of gas than wood chips but suffered a greater pressure loss [3]. 4.4. Pressure drop Gas phase pressure drop through a biofilter bed increases with gas flowrate and diminishing particle size. Yang and Allen [9] report increase in pressure drop of (0–35 kPa m−1 ) which are reasonably linear with increasing gas velocity (0–0.3 m s−1 ) over a range of bed medium particle sizes (1–12 mm). However, pressure drop appears to increase exponentially with decreasing particle size, especially with particles less than 1 mm. Brennan et al. [13] found pressure drop to increase significantly across a fibrous mixture compared with peat nodules and granules. Compaction of the filter bed over extended periods of usage and due to over watering will also give rise to prohibitive pressure drops [18]. Pressure drop increased from less than 500 to greater than 2500 Pa after 3 months of continuous operation of a biofilter treating 40◦ C exhaust from a hardboard plant. After a 25 day pause in operation, the pressure drop was reduced to less than 1000 Pa [3]. Ergas et al. [10] report pressure drops of 100–600 Pa m−1 at corresponding superficial air velocities through the biofilter of 0.3–1.8 m3 m−2 min−1 . These pressure drops were achieved by addition of 50% by volume of perlite to the filter media. In their review of operational biofilters, Williams and Miller [8] note the unpredictability of pressure drop across differing bed media and recommend pilot testing of individual media. Monitoring of the pressure drop across the bed was considered important for detection of cracks in the media and resultant short-circuiting of the bed by the air stream. Any increase in pressure drop adds to the operating cost of the biofilter as odorous air must be supplied at a greater pressure to achieve the same flowrate.
4.5. Surface area Specific surface areas for soils and compost range from 1–100 m2 g−1 [7]. Tang et al. [14] estimated the surface area of their sieved compost and chaff medium to be 180 m2 m−3 . Peat has even greater surface areas which result in high adsorptive capacity [29,30]. A high surface area provides access for biological growth and adsorption which are the two mechanisms by which depletion of odour from the gas phase may occur. Adsorptive capacity in the organic medium allows the biofilter to withstand fluctuations in loading rates without compromising removal rate. Higher than average loads of odorous compounds are accommodated on the solid surface. Conversely, micro-organisms can gain their carbon and energy requirements during lean loading periods from compounds that are then desorbed from the surface. Many organic materials utilised for biofiltration have a large capacity for adsorption of odorous compounds. Yang and Allen [27] consistently removed greater than 99.5% of the hydrogen sulfide supplied to their compost biofilter when loading rates were varied from 2 to 101 g S m−3 h−1 . 4.6. pH Biological metabolism is strongly dependent on pH. Many micro-organisms will only grow within a particular pH range. As a rough rule of thumb, most biological growth occurs near a neutral pH and wide deviations from this will impair the efficiency of the biofilter. A notable exception is the sulfur oxidising bacteria which thrive at low pHs. Biological oxidation more often than not results in a net increase in acidity as protons are exchanged for electrons. This includes nitric and nitrous acids as products of ammonia degradation, sulfuric acid as a product of sulfide degradation and carbon dioxide as a product of oxidation of organic compounds. Input parameters for the biofilter model of Hodge and Devinny [22] for removal of ethanol vapour in air include adsorptive capacity, porosity and buffer capacity of the filter media which enabled comparison of the effectiveness of various media. Insufficient buffering capacity as well as carbon dioxide evolution yielded a pH gradient in one of their biofilters. Yang and Allen [9] found that H2 S removal efficiency in their biofilter decreased markedly at pH below 3.2 but was almost independent of pH at higher values. They suggest that the dominant active species present were acidophiles which prefer an optimum pH near 3. The pH at the inlet to their biofilter decreased from 8.0 to 2.5 after 32 days operation. The pH was higher at the end of the biofilter as biological activity decreases through the biofilter with concentration of hydrogen sulfide [27]. Brennan et al. [13] also recorded decreases in pH from 6.5–7.0 to 3.6–4.8 after 3 weeks in a bank of biofilters exposed to hydrogen sulfide and methyl mercaptan. After 6 months, the pH in untreated biofilters fell below 2. Similarly, Kowal et al. [11] observed a decrease
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in pH from 6.5 to 3.5 after 60 days of operation of their activated sludge biofilter. Crushed oyster shells are used extensively as a source of alkalinity for buffering and as a carbon source for the autotrophic bacteria responsible for nitrification and sulfide oxidation [10,19]. Maërl, a mineral support of marine origin containing 82% calcium carbonate, was found by Bonnin et al. [20] to increase hydrogen sulfide removal in biofilters at sewage treatment plants from 10 g m−3 h−1 (peat without Maërl) to 50 g m−3 h−1 . 4.7. Conditioning Peat suffers from its low natural pH which makes establishment of many biological populations difficult without some form of pretreatment. Peat, compost and soils are all characterised by the presence of humic acids which contribute to ion exchange capacity on the solid surface. Pretreatment or conditioning of biofilter media may enhance operation. This may involve adjusting the pH of an acidic peat to encourage the growth of nitrifying bacteria, adding alkalinity such as carbonate as a carbon source for autotrophic bacteria or stabilising compost so that it is at equilibrium before adding it to a biofilter. Yang and Allen [9] noticed no significant differences in operating characteristics as a result of varying compost storage times before using it as a bed material in a biofilter column.
5. Air phase 5.1. Flowrate Gas phase residence times in biofilters can range from minutes for alcohols to 3 h for 90% removal of trichloroethylene [7]. Examples include 15 s for removal of meat-meal industry odours [1] and 23 s for 99% removal of hydrogen sulfide [9]. Maximum removal in a biofilter is obtained when the gas flowrate is as high as possible such that the maximum removal capacity R of the biofilter is not exceeded. A further constraint is imposed if mass transfer of odorous compounds from the gas phase to the biofilm is limiting. Yang and Allen [9] found that for gas phase residence times less than 23 s, their biofilter removal rates suffered from resistance to transfer of hydrogen sulfide from the gas phase to the biofilm. Superficial gas phase velocities are typically 30–200 m h−1 (50–300 cm min−1 ) in odour removing biofilters. Ottengraf and Van den Oever [25] supplied toluene from 0.3 to 5.6 g m−3 to their 3 m peat moss biofilter column. Their inlet superficial gas velocities varied correspondingly from 1 to 15 cm s−1 (30–530 m h−1 ). A superficial gas velocity of 75 m h−1 was suitable for degradation of 40◦ C hard board plant exhaust [3]. Ergas et al. [10] performed a series of pilot scale experiments removing chlorinated aliphatic compounds from air and then tested their biofilter on site
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at a water treatment plant, monitoring odorous compound removal, including aromatic VOC. Pilot scale studies involved air flowrates ranging from 0.5 to 2 m3 m−2 min−1 (30 to 120 m h−1 ) while the water treatment plant air was fed at 0.3 m3 m−2 min−1 (18 m h−1 ) to one bed and 1 m3 m−2 min−1 (60 m h−1 ) to the other bed. 5.2. Concentration Concentration loads of 0–500 ppm are common in biofilter applications [7]. When first-order kinetics are in force in a biofilter, the removal rate 8 is independent of concentration of odorous compounds in the gas inlet. This will be the case as long as the inlet concentration is low enough that the kinetics of biological degradation do not become saturated. Yang and Allen [9] were able to consistently remove greater than 99% of hydrogen sulfide from the gas phase inlet to their biofilter in the range 5–2650 ppm while operating below the maximum removal capacity. In their review of operational biofilters, Williams and Miller [8] recommend that at least 100 parts of oxygen should be provided in the air stream for every part of oxidisable odorous gas. Because odours are typically dilute, this condition is usually met. 5.3. Humidity If the biofilter bed is not watered, odorous air should be prehumidified before entering to prevent drying of the organic bed. Koch et al. [1] describe how off gases to biofilters in the meat-meal industry must be scrubbed with water to remove entrained fat and dust that will otherwise clog the biofilter bed. The scrubber also humidifies the air to greater than 95% so that additional sprinkling of the bed to keep it moist is not necessary [26]. Ergas et al. [10] performed a series of pilot scale experiments removing chlorinated aliphatic compounds from air and then tested their biofilter on site at a water treatment plant, monitoring odorous compound removal, including aromatic VOC. Humidification of inlet air was performed for the pilot-scale studies but was unnecessary at the water treatment plant as the air was already saturated. Moisture content in the filter beds was supplemented by soaker hoses at the top of the beds.
6. Liquid phase A liquid phase flowing through the biofilter bed may serve a number of purposes. It may maintain an adequate moisture content, dilute toxic metabolic products, provide additional nutrients for biological growth and provide buffering for pH stability. Dilution has been found to be particularly important for hydrogen sulfide biofiltration. Brennan et al. [13] stress the importance of a watering system for dilution and maintaining pH levels within an optimal range for biofilters
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removing reduced sulfur compounds. While sulfide removing bacteria may thrive at lower pH, other bacteria responsible for degrading different odorous compounds generally do not. Hence, a high sulfide removal rate in a biofilter does not necessarily imply a high overall odour removal. Yang and Allen [27] washed the compost in their hydrogen sulfide removing biofilter periodically with distilled water in order to dilute H2 SO4 which accumulated in the bed. This was found to potentially extend the life of the biofilter. This was also the experience of Bonnin et al. [20] who supplied water at a flowrate sufficient to dilute sulfate in their hydrogen sulfide removing biofilter such that calcium sulfate did not precipitate (solubility 3 g l−1 ). Required superficial velocities for a liquid phase are generally very low compared with those for the gas phase due to the slow rates of bed drying and biological growth.
7. Temperature Low operating temperatures will enhance sorption of odorous compounds into the biofilm but will slow down the microbial growth. Higher temperatures will have the reverse effect. High performance for most odour removing applications occurs within a temperature range of 25–40◦ C with an optimum occurring around 37◦ C [7]. Bed temperature is primarily dictated by input gas temperature as heat liberated by biological oxidation is negligible. Knauf and Zimmer [3] found that removal efficiency for organics in hot exhaust from a hardboard plant decreased steadily as temperature of the exhaust increased. For bark compost, removal efficiency dropped from 95 to 85% as temperature rose from 40 to 55◦ C. For wood chips, efficiency dropped from 80 to 70% as temperature rose from 35 to 50◦ C. Hydrogen sulfide oxidising bacteria present in the biofilter of Yang and Allen [9] were most active in the temperature range 25–50◦ C. Activity, measured as hydrogen sulfide removal efficiency at constant loading, dropped markedly on either side of this temperature range. Brennan et al. [13] found that removal rates in their hydrogen sulfide and methyl mercaptan removing biofilter decreased by over 50% when ambient temperature decreased from 20–22◦ C to 9–12◦ C. Biofiltration can be effective at low temperatures. Pinnette et al. [18] found that, once a biological population was established in their biofilter, odour removal above sludge composting facilities was not compromised at temperatures below 10◦ C.
8. Microbial population Bohn [7] estimates that microbiological populations in biofilters are of the order of one billion micro-organisms per gram of organic material. Quite often, an appropriate biological population for degradation of a particular odour
can be found near the source of the odour. For example, micro-organisms capable of degrading VOC from oil refining operations have been found in soil adjacent to tanks where overflow and spillage were common. Because of their relatively fast population doubling times, bacteria may evolve within a few years to grow more efficiently on a predominant nutrient. 8.1. Acclimatisation Acclimatisation periods are often required for microorganisms to adapt to new conditions imposed by a new biofilter bed medium. These will include biological nutrients, pH, temperature and solid surface characteristics. Because of their ability to evolve, micro-organisms often improve their odour removing ability through the life of the biofilter. In their review of operational biofilters, Williams and Miller [8] reported acclimatisation periods of greater than 10 days. Even longer times were recorded by Wang et al. [4] in their investigation of biological water filters. They found that biomass accumulation was a slow process, taking up to 5 months for biomass to reach an equilibrium concentration. Biomass was measured via a phospholipid extraction technique that showed a decreasing biomass concentration with depth in the bed. This indicates that most biological degradation takes place at the front of the bed where initial contact with the waste fluid takes place. Further, it was found that the capacity of the bed to remove organic material did not always increase with biomass concentration. It was suggested that above a critical biomass level, the removal capacity of the filter was somewhat independent of biomass concentration. Shareefdeen et al. [28] conditioned their methanolremoving biofilter with cultures of bacteria that had been exposed to methanol vapour. Agricultural soil was placed in an enclosure with the vapour and, after a week, soil samples were placed in a mineral medium with 1% methanol and sufficient airspace to prevent oxygen limitation. It was found that a consortium of eight methanol-utilising bacterial strains was found to grow consistently at higher rates than any of its individual members. Ergas et al. [10] inoculated their odour removing biofilter with a compost containing 50% digested sewage sludge and 50% forest products. Mixed liquor suspended solids from a waste water treatment plant were also added at 7 l per m3 of filter material, followed by a Pseudomonas putida strain at 106 cells per gram of filter medium. Brennan et al. [13] inoculated hydrogen sulfide and methyl mercaptan removing biofilter with cultures of Thiosphaera pantotropha and Thiobacillus neapolitanus. They found that these inoculated biofilters provided significant improvement in removal over uninoculated biofilters. Yang and Allen [9] found that an acclimatisation period was required to reach maximum removal efficiency for H2 S. This was presumably required for the microbial population
D. McNevin, J. Barford / Biochemical Engineering Journal 5 (2000) 231–242
to build up to its optimum size after exposure to H2 S. The time required was up to 2 weeks. They did not inoculate with sulfide oxidising bacteria or previously expose the medium to hydrogen sulfide. Contrary to Brennan et al. [13], they concluded that the bacteria are ubiquitous in soils and compost. They found no advantage from adding sewage sludge to their media. 8.2. Inhibition It is common for nutrients required for biological growth and biological end products to inhibit growth in significant concentrations. As such, biofiltration is well suited to odour removal as concentrations of odorous compounds are typically dilute. Pinnette et al. [18] concluded that high levels of ammonium in their biofilter bed contributed to poor odour removing performance during start up due to inhibition of biological populations by ammonium. Yang and Allen [9] found that sulfate concentrations of more than 25 mg S (g dry compost) inhibited the performance of their hydrogen sulfide removing biofilter. This was significant because sulfate is a product of the biological oxidation of sulfide. Elemental sulfur also accumulated at the entrance to the biofilter at high loading rates [27]. 8.3. Kinetics It is difficult to extract consistent quantitative biological rate constants for biofiltration of odours due to the widely varying experimental data reported. Biological rates are very dependent on biofilter bed medium, pH, alkalinity and operating temperature, none of which are likely to be identical in any pair of experiments. Most of the available data pertains to biofiltration of hydrogen sulfide and a very thorough investigation of this by Yang and Allen [27] is analysed here. They determined values for µ0 (26.1 ppmv s−1 =131 g m−3 h−1 ) and µ1 (0.54 s−1 =1900 h−1 ) in their hydrogen sulfide removing biofilter. Thus, at high inlet gas phase concentrations of hydrogen sulfide dCg µ0 εA Cg µ0 1 =− ≈ = −131 dz Fg K + Cg ug ug where ug (m h−1 ) is the gas phase interstitial velocity given by ug =
Fg εA
(19)
241
Hence the kinetics for hydrogen sulfide degradation are given by r = 131 ×
Cg 67 + Cg
where the units of Cg are mg m−3 . These results should be treated with some caution as µ1 was determined for Cg,in <280 mg m−3 . According to their results, first-order kinetics would only be expected for Cg K=67 mg m−3 . However, their values for µ0 (131 g m−3 h−1 ) and, by implication, K (67 mg m−3 ) are indicative and provide insight into when first- and zeroth-order kinetics might be expected in a hydrogen sulfide removing biofilter. It is difficult to determine biological rate constants for biofiltration of ammonia due to the limited success reported in the literature. In biological nutrient removal from wastewater, the International Association on Water Quality recommends K=1.3 g m−3 (1300 mg m−3 ) for degradation in the liquid phase. Given that, at 25◦ C, the Henry’s law coefficient for ammonia in water is approximately 0.0037 [31], then this liquid phase concentration corresponds to K=4.8 mg m−3 in air. It is even less meaningful to quantify rate constants for VOC and ODU biofiltration due to the vast numbers of individual odorous compounds involved.
9. Conclusions It is difficult to compare the performances of biofilters in the literature when it is not clear whether they are operating subject to zeroth- or first-order kinetics of contaminant degradation. An absolute maximum elimination rate can only be determined by measuring the removal rate of contaminants as first flowrate and then concentration are increased. A qualitative kinetic analysis of biological degradation in biofilters suggests that the mean gas phase residence time in a biofilter should be large enough for mass transfer of contaminants from gas to biofilm to occur and for the biological kinetics to remain unsaturated. If the residence time is too large, however, then the maximum elimination capacity of the biofilter will not be attained. Besides kinetics, attention must be paid to biofilter bed media to optimise performance. Parameters such as moisture content, porosity, surface area and pH will affect adsorption and biological degradation of contaminants.
At low inlet gas phase concentrations of hydrogen sulfide Cg dCg µ0 εA Cg µ0 =− Cg = 1900 ≈ dz Fg K + Cg ug K ug It is then possible to calculate a value for K according to 131 µ0 = = 1900 h−1 K K
and
K = 67 mg m−3
References [1] W. Koch, H.-G. Liebe, B. Striefler, Operating experience with biofilters to reduce odorous emissions, Staub-Reinhaltung Der Luft 42 (12) (1982) 488–493. [2] M. Walters, Love thy neighbour, PACE: Process Chem. Eng. 48 (1) (1995) 18.
242
D. McNevin, J. Barford / Biochemical Engineering Journal 5 (2000) 231–242
[3] S. Knauf, H. Zimmer, Biofiltration at a temperature above 40◦ C: comparison of the biofilter materials bark compost and woodchips, Staub-Reinhaltung der Luft 54 (2) (1994) 41–44. [4] J.Z. Wang, R.S. Summers, R.J. Miltner, Biofiltration performance. Part 1. Relationship to biomass, J. Am. Water Works Assoc. 87 (12) (1995) 64–70. [5] S.P.P. Ottengraf, Biological systems for waste gas elimination, Trends Biotechnol. 5 (5) (1987) 132–136. [6] G. Leson, A.M. Winer, Biofiltration: an innovative air pollution control technology for VOC emissions, J. Air Waste Manage. Assoc. 41 (8) (1991) 1045–1054. [7] H. Bohn, Consider biofiltration for decontaminating gases, Chem. Eng. Progress 88 (4) (1992) 34–40. [8] T.O. Williams, F.C. Miller, Biofilters and facility operations, BioCycle 33 (11) (1992) 75–79. [9] Y. Yang, E.R. Allen, Biofiltration control of hydrogen sulfide: 1. Design and operational parameters, J. Air Waste Manage. 44 (1994) 863–868. [10] S.J. Ergas, E.D. Schroeder, D.P.Y. Chang, R.L. Morton, Control of volatile organic compound emissions using a compost biofilter, Water Environ. Res. 67 (5) (1995) 816–821. [11] S. Kowal, J.L. Fanlo, J.R. Degorce-Dumas, P. Le Cloirec, Removal of H2 S by a biofilter with dry activated sludge: an approach of mechanisms, in: Proceedings of the International Symposium on Environmental Biotechnology Oostende, Royal Flemish Society of Engineers, 1991, pp. 135–138. [12] J. Monod, The growth of bacterial cultures, Annu. Rev. Microbiol. 3 (1949) 371–394. [13] B.M. Brennan, M. Donlon, E. Bolton, Peat biofiltration as an odour control technology for sulphur-based odours, J. Chartered Institution Water Environmen. Manage. 10 (3) (1996) 190–198. [14] H-M. Tang, S-J. Hwang, S-C. Hwang, Waste gas treatment in biofilters, J. Air Waste Manage. Assoc. 46 (4) (1996) 349– 354. [15] J.C. Van Den Heuvel, Mass transfer in and around biofilms, in: L.F. Melo et al. (Eds.), Biofilms: Science and Technology, Kluwer Academic Publishers, Dordrecht, 1992, pp. 239–250. [16] S.J. Ergas, K. Kinney, M.E. Fuller, K.M. Scow, Characterization of a compost biofiltration system degrading dichloromethane, Biotechnol. Bioeng. 44 (1994) 1048–1054. [17] S.M. Zarook, A.A. Shaikh, Analysis and comparison of biofilter models, Chem. Eng. J. 65 (1997) 55–61.
[18] J.R. Pinnette, M.D. Giggey, G.J. Marcy, M.A. O’Brien, Performance of biofilters at two agitated bin composting facilities, in: Proceedings of the 87th annual meeting of the Air and Waste Management Association, Ohio, Air and Waste Management Association, 1994. [19] R. Kapahi, M. Gross, Biofiltration for VOC and ammonia emissions control, BioCycle 36 (2) (1995) 87–90. [20] C. Bonnin, G. Martin, G. Gragnic, Biopurification of odorous gases in urban wastewater treatment plants, in: S. Vigneron, J. Hermia, J. Chaouki (Eds.), Characterization and Control of Odours and VOC in the Process Industries, Proceedings of the Second International Symposium on Characterization and Control of Odours and VOC in the Process Industries, Louvain-la-Neuve, Belgium, 3–5 November 1993, Elsevier, Amsterdam, 1994, pp. 431–444. [21] A.P. Togna, M. Singh, Biological vapor-phase treatment using biofilter and biotrickling filter reactors: practical operating regimes, Environ. Progress 13 (2) (1994) 94–97. [22] D.S. Hodge, J.S. Devinny, Modelling removal of air contaminants by biofiltration, J. Environ. Eng. 121 (1) (1995) 21–32. [23] M.A. Deshusses, G. Hamer, I.J. Dunn, Behaviour of biofilters for waste air biotreatment. 2. Experimental evaluation of a dynamic model, Environ. Sci. Technol. 29 (4) (1995) 1059–1068. [24] M. Rothenbühler, M. Heitz, M. Beerli, B. Marcos, Biofiltration of volatile organic emissions in reference to flexographic printing processes, Water Air Soil Pollut. 83 (1995) 37–50. [25] S.P.P. Ottengraf, A.H.C. Van Den Oever, Kinetics of organic compound removal from waste gases with a biological filter, Biotechnol. Bioeng. 25 (1983) 3089–3102. [26] F.H.H. Valentin, Peat beds for odour control: recent developments and practical details, Filtration Sep. 23 (4) (1986) 224–227. [27] Y. Yang, E.R. Allen, Biofiltration control of hydrogen sulfide. 2. Kinetics biofilter performance and maintenance, J. Air Waste Manage. Assoc. 44 (11) (1994) 1315–1321. [28] Z. Shareefdeen, B.C. Baltzis, Y.S. Oh, R. Bartha, Biofiltration of methanol vapor, Biotechnol. Bioeng. 41 (1993) 512–524. [29] A.M. Martin, Biofilms in peat biofilters, in: L.F. Melo et al. (Eds.), Biofilms: Science and Technology, Kluwer Academic Publishers, Dordrecht, 1992, pp. 461–466. [30] D. Mcnevin, J. Barford, J. Hage, Adsorption and biological degradation of ammonium and sulfide on peat, Water Res. 33 (6) (1999) 1449–1459. [31] E. Wilhelm, R. Battino, R.J. Wilcock, Low-pressure solubility of gases in liquid water, Chem. Rev. 77 (2) (1977) 219–230.