Developments in odour control and waste gas treatment biotechnology: a review

Biotechnology Advances 19 (2001) 35 ± 63

Research review paper

Developments in odour control and waste gas treatment biotechnology: a review Joanna E. Burgess*, Simon A. Parsons, Richard M. Stuetz School of Water Sciences, Cranfield University, Cranfield, Bedford, Bedfordshire, MK43 0AL, UK

Abstract Waste and wastewater treatment processes produce odours, which can cause a nuisance to adjacent populations and contribute significantly to atmospheric pollution. Sulphurous compounds are responsible for acid rain and mist; many organic compounds of industrial origin contribute to airborne public health concerns, as well as environmental problems. Waste gases from industry have traditionally been treated using physicochemical processes, such as scrubbing, adsorption, condensation, and oxidation, however, biological treatment of waste gases has gained support as an effective and economical option in the past few decades. One emergent technique for biological waste gas treatment is the use of existing activated sludge plants as bioscrubbers, thus treating the foul air generated by other process units of the wastewater treatment system on site, with no requirement for additional units or for interruption of wastewater treatment. Limited data are available regarding the performance of activated sludge diffusion of odorous air in spite of numerous positive reports from full-scale applications in North America. This review argues that the information available is insufficient for precise process design and optimization, and simultaneous activated sludge treatment of wastewater and airborne odours could be adopted worldwide. D 2001 Elsevier Science Inc. All rights reserved. Keywords: Biotreatment; Foul air treatment; Hydrogen sulphide; Offgas; VOCs

1. Introduction Odours in wastewater treatment arise mainly from the biodegradation of sewage, especially anaerobic degradation. Other odours associated with wastewater treatment come either * Corresponding author. Tel.: +44-1234-750111; fax: +44-1234-751671. E-mail address: [email protected] (J.E. Burgess). 0734-9750/01/$ ± see front matter D 2001 Elsevier Science Inc. All rights reserved. PII: S 0 7 3 4 - 9 7 5 0 ( 0 0 ) 0 0 0 5 8 - 6

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directly from industrial wastewater (solvents, volatile organic compounds, petroleum derivatives) or indirectly from warm, highly degradable, or sulphurous effluents (Vincent and Hobson, 1998). Odours are generated by a number of different wastewater components (Vincent and Hobson, 1998), the most significant being the sulphur compounds, hydrogen sulphide (H2S) and mercaptan. Domestic sewage contains 3±6 mg/L organic sulphur, mainly arising from proteinaceous materials, plus  4 mg/L from sulphonates contained in household detergents (Boon, 1995) and 30±60 mg/L inorganic sulphur (as sulphates) (Gostelow and Parsons, 2000). While nitrogen-based odour compounds and the organics associated with anaerobic wastewater treatment are important, hydrogen sulphide is toxic to microorganisms and corrosive to concrete and mild steel, and as the predominant odorant associated with sewers (Vincent and Hobson, 1998), it has been extensively studied (Halkjaer-Nielsen et al., 1998; Gostelow and Parsons, 2000). A range of technologies is available to treat odorous air emitted from wastewater treatment plants, sludge handling facilities, and industrial processes. There are three methods of odour treatment, biochemical (biofilters, bioscrubbers, activated sludge), chemical (chemical scrubbers, thermal oxidation, catalytic oxidation, ozonation), and physical [condensation, adsorption (activated carbon), absorption (clean water scrubbers)]. The selection of a particular technology or combination of technologies is dependent on such factors as: site characteristics including operation and maintenance capabilities, treatment objectives, foul air flowrates, contaminant loading patterns, and the characteristics and strength of odorous air. Waste gases from industry have traditionally been treated using physicochemical processes, such as scrubbing, adsorption, condensation, and oxidation (Table 1); biological treatment of waste gases has only gained support as an effective and economical option in the past few decades (Kennes and Thalasso, 1998). Biological methods of odour treatment gained much attention in Europe in the 1990s owing to their efficiency, cost-effectiveness, and environmental acceptability, and by 1994 accounted for 78% of odour treatment in Germany (Frechen, 1994). In all types of bioreactors for waste gas treatment, the pollutants diffuse into the liquid phase where microorganisms degrade them into products, such as CO2, H2O, and minerals (WEF/ASCE, 1995; Kennes and Thalasso, 1998; Vincent and Hobson, 1998). Microorganisms are known to play an important role in geochemical and biogeochemical cycles by mineralising biopolymers and xenobiotic compounds (Lie et al., 1998), although this was recognised only 30 years ago, and current understanding of the mechanisms of biodegradation of such materials is still not complete (Tan and Field, 2000). According to Brauer (1986), the transformation process can be expressed simply by: Odorous gas ‡ oxygen !via

bacteria

! more bacterial cells ‡ carbon dioxide ‡ water

When provided with an oxygen source, bacteria in wastewater consume ionic sulphide species and oxidise them to nonodorous sulphur species. This mechanism suggests an acclimated biological mixture, such as returned activated sludge, could convert dissolved sulphide into nonodorous sulphur species, if provided with an oxygen source (Joyce, 1995). It is believed that odour-producing compounds are removed by adsorption and/or dissolution in the liquid and a combination of chemical and biological oxidation (Ostojic et al., 1992). A summary of the removal methods is presented in Table 2. When Cho et al. (1992) carried out some work on H2S oxidation using a peat biofilter colonised by Xanthomonas sp. strain






Simple

Chemical counteractant

Marginal effectiveness ( < 40% odour reduction) High ongoing chemical costs

high capital and energy costs
Very Only
economical for high-strength, recalcitrant air streams or VOCs


Consistently high performance
Mechanically simple


Effective

Activated carbon

Fine mist wet scrubber

Packed tower wet scrubber

Thermal oxidiser

Disadvantages of unused chemical
Blowdown
High maintenance capital cost than a packed tower
Higher water requires softening
Scrubber Larger footprint

High O&M cost (carbon replacement/ regeneration)
Only for low contaminant loads (to ensure acceptable carbon life)

Advantages

effective and reliable
Proven
Moderate capital and O&M costs be designed for VOC removal
Can Lower chemical consumption and
hence chemical costs

Treatment process

Table 1 Summary of odour treatment chemical technologies (Bowker, 2000b)

High

Low

High

Moderate

High

Level of use

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Table 2 Mechanism responsible for the removal of odorous compounds in an activated sludge basin Removal method

Comments

Absorption

Odorous compounds are transferred from the gas phase into the bulk liquid. Water-soluble compounds, such as H2S, can be readily absorbed into the liquid. The mass of material transferred is a function of the surface area of the bubbles, the contact time, and the diffusivity coefficient. Some high molecular weight compounds with lower solubility may be physically adsorbed onto the biological floc. Relatively warm, odorous air that is transferred into a liquid of lower temperature will result in condensation of volatile organic compounds. Due to the high concentration of active, aerobic microorganisms, biological oxidation is likely responsible for a significant conversion of odorous compounds that initially are absorbed by the liquid, absorbed onto the floc, or condensed.

Adsorption Condensation Biological oxidation

DY44, the same levels of H2S removal were achieved with viable cells and a g-irradiated cell suspension. The importance of biological oxidation, therefore, comes into question. Contrary to Cho et al.'s findings, Kanagawa and Mikami (1989) found that in the experiments they conducted to remove H2S from contaminated air using Thiobacillus thioparus, H2S oxidation occurred in the sterilised medium, although the rate was very slow. For this reason, these workers concluded that H2S oxidation was carried out mainly by microorganisms. In a paper addressing the kinetics of chemical and biological sulphide oxidation, Buisman et al. (1994) discovered that at a sulphide concentration of around 150 mg/L the biological oxidation rate was seven times faster than the chemical oxidation rate. The biological oxidation rate below 10 mg of sulphide per liter is a factor 75 times faster than chemical oxidation. The literature documents examples of the use of phototrophic (Cork et al., 1983), heterotrophic (Cho et al., 1992), and autotrophic (Jensen and Webb, 1995; Sublette and Sylvester, 1987a,b; Kanagawa and Mikami, 1989) microorganisms for the desulphurisation of gas. Biological oxidation of sulphide to elemental sulphur has several advantages over other physicochemical processes (Buisman et al., 1991). The advantages presented include economic gain, as there is a comparably low capital outlay (Gadre, 1988; Benedek et al., 1988) and money is saved on oxidants and catalysts, resulting in a reduction in operating cost (Fox and Venkatasubbiah, 1996; Janssen et al., 1999; Li et al., 1998; Buisman et al., 1991; Comas et al., 1999; Jensen and Webb, 1995; Benedek et al., 1988; Kanagawa and Mikami, 1989). It is possible to recover sulphur from a biological process and this can be reused (Fox and Venkatasubbiah, 1996; Buisman et al., 1991). No chemical sludge is produced (Fox and Venkatasubbiah, 1996; Gadre, 1988; Buisman et al., 1991), and there is a reduction in the sulphate or thiosulphate discharge (Fox and Venkatasubbiah, 1996; Buisman et al., 1991). In general, biological sulphur removal consumes less energy than physicochemical removal methods (Fox and Venkatasubbiah, 1996; Gadre, 1988; Buisman et al., 1991; Jensen and Webb, 1995; Kasakura and Tatsukawa, 1995). According to Comas et al. (1999), chemical absorption is approximately 62% more expensive than biological processes.

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The use of biotechnologies for foul air treatment has grown dramatically because of their ability to destroy the pollutants rather than simply transfer them from the gas to the liquid phase, although chemical odour treatment remains popular. Published performance data for odour compounds treated using similar process units vary widely (Veiga et al., 1997; Kennes and Thalasso, 1998), suggesting that research into the parameters affecting the performance of biological air treatment would significantly improve our ability to use odour and air pollution control biotechnologies to their full potential. Three major types of bioreactor that currently dominate waste gas biotreatment are biofilters, trickling biofilters, and bioscrubbers (Tables 2, 4, and 5). A recent review by Kennes and Thalasso (1998) presents the state of reactor design and factors for optimization. In addition to these three most widely used technologies, other alternatives have been proposed, such as an external loop airlift bioreactor (Ritchie and Hill, 1995), a spiral bioreactor (Shim et al., 1995), membrane bioreactors (Reij et al., 1995, 1997, 1998; Parvatiyar et al., 1996a,b; Ergas and McGrath, 1997; Ergas et al., 1999), and activated sludge diffusion (Fukuyama and Honda, 1976; Fukuyama et al., 1979, 1980, 1981, 1986; Ostojic et al., 1992; Ryckman-Siegwarth and Pincince, 1992; Frechen, 1994; Stillwell et al., 1994; Bentzen et al., 1995; Johnson et al., 1995; WEF/ASCE, 1995; Bielefeldt et al., 1997; ásùy et al., 1998; Vincent and Hobson, 1998; Bowker, 1999, 2000a; Oppelt et al., 1999). 2. Bulk media biofilters The odorous compounds present are degraded as the contaminated air passes through a bed of naturally occurring microbes immobilised on a support media. The air stream and media bed are moistened to facilitate microbial activity. Odorous compounds in the air stream provide a source of carbon for the biomass, while the moisture added is used to supply other

Fig. 1. Closed aerated biofilter (WEF/ASCE, 1995).

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Fig. 2. Open soil or compost biofilter (WEF/ASCE, 1995).

nutrients (e.g. nitrogen, phosphorus, potassium) to the biofilm. Because of the low density of microbes, which are present as a mixed culture, the specific performance of a biofilter is relatively low. One type of biofilter is a closed concrete (or similar) vessel containing an aeration system (Fig. 1). Open soil and compost biofilters (Fig. 2) often consist of a trench filled with gravel to evenly distribute the foul air, containing a perforated pipe through which the gas stream is pumped at approximately 0.25±0.5 kPa (WEF/ASCE, 1995). A common construction method is to excavate a 1-m deep, appropriately sized hole, place a gravelcovered distribution network in the bottom, and fill with the excavated soil, or an alternative, such as sand or compost. Both systems are covered in wet climates to limit operational problems presented by heavy precipitation. The media employed in biofilters include peat (RothenbuÈhler et al., 1995; WEF/ASCE, 1995; Kennes and Thalasso, 1998; Wang et al., 1996; Zilli et al., 1996), heather (WEF/ASCE, 1995), soil (WEF/ASCE, 1995; Kennes and Thalasso, 1998), compost (WEF/ASCE, 1995; Kennes and Thalasso, 1998), and sand (WEF/ ASCE, 1995), with soil and compost the most common media types (WEF/ASCE, 1995). Many different materials have been used as compost, including domestic rubbish (Tang et al., 1996), digested sewage sludge (Morgenroth et al., 1996; Tang et al., 1996), forest subproducts (Morgenroth et al., 1996; Veir et al., 1996), and wastewater biosolids (Veir et al., 1996). The critical properties of biofilter media are uniform particle size and sufficient Table 3 Summary of biofiltration for odour treatment Advantages

Simple, flexible design with low capital costs
(Kolsteltz et al., 1996; Kennes and Thalasso, 1998) Good for treating high volumes of low concentration
sulphurous odorants (Smet and van Langenhove, 1998) 99% removal efficiency in streams containing
aldehydes, organic acids, sulphur dioxide, nitrous oxides, and hydrogen sulphide

90% removal of methane, propane, and isobutane
(WEF/ASCE, 1995) with air/water partition coefficients
ofCompounds up to 10 can be treated in recirculating biofilters

because the gas residence time 930 ± 60 s and specific surface area (300 ± 1000 m2/m3) are both high (Smet and van Langenhove, 1998).

Disadvantages

criteria still developing
Design land area required
Large
Media requires regular replacement (hence, high O&M costs). ppm H S can lead to rapid acidification
of >the15filter media (Vincent and Hobson, 1998). of gas into liquid is the rate-limiting
step,Dissolution so long gas residence times are required 2

(Kennes and Thalasso, 1998). Large media bed volumes are required to obtain such a long gas residence time, and operational control is limited, as there is no liquid phase involved (Smet and van Langenhove, 1998).




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porosity, particles with large surface areas, some pH buffering capability, and the ability to support microorganisms. Soil is a common choice owing to its availability, porosity, particle size distribution, and existing microbial flora. It is also able to outlast peat and compost in terms of life span before replacement is required (Kennes and Thalasso, 1998). Soil florae Ð bacteria, fungi, and actinomycetes Ð metabolise biodegradable odorants. Biofilters effectively treat air pollutants with air/water partition coefficients (Henry's Law constants) of less than 1.0 (Kennes and Thalasso, 1998), as there is no liquid phase and water solubility of the pollutants is of relatively little importance, but can only cope with relatively low loads of hydrogen sulphide (Vincent and Hobson, 1998). Advantages and disadvantages of biofiltration are shown in Table 3. Media require replacing when the pH becomes too acidic for the survival of many microorganisms, a build-up of toxins occurs, or the media becomes compacted and loses porosity. Important features in the design and operation of biofilters

Table 4 Influential factors in design and operation of media-based biofilters Factor

Target value

Influence

Reference

Oxygen concentration Moisture content

Air/odour ratio of 100:1 v/v Soil, 10 ± 25%; Compost, 20 ± 50%; Peat, 40%

a

Temperature

Optimum, 37°C

pH

Neutral

Odour residence time

Variable (minimum 30 s)

Sufficient oxygen is required for microbial oxidation of odorous compounds in the gas stream. Control of moisture content is required to counter the drying action of the gas stream, either via water sprays above the media or by humidity control in the foul air stream. Dry media channel gases do not support microorganisms. Inlet air temperature should be adjusted, especially where humidified using steam, as biodegradation generates heat and maintains high temperatures, even in biofilters in cold climates, although biofilters are reported to survive temporary conditions of < 0 ± 2°C. Microorganisms thrive at neutral or near-neutral pH, but as hydrogen sulphide is degraded, sulphuric acid is produced. Biofilters have some buffering capacity (especially soil) but long-term, high-hydrogen sulphide loading necessitates new media or lime addition for pH control. Dependent on the media used and odorants treated. The components of the foul air should be identified and quantified, so the biofilter is sized to give an appropriate residence time for the degradation rates of the components. Media porosity varies, dependent on moisture content and media settling, affecting the pressure required to pass odorous air through the filter.

Pressure drop

a, b

c

a, d, e, f, g, h, i

a, j

a

a, WEF/ASCE, 1995; b, Wang et al., 1996; c, Holubar and Braun, 1995; d, Smet et al., 1996; e, Devinny and Hodge, 1995; f, Morgenroth et al., 1996; g, Veir et al., 1996; h, Ergas et al., 1995; i, AcunÄa et al., 1996; j, Kennes and Thalasso, 1998.

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include the availability of oxygen and moisture and the physical conditions that prevail (Table 4). 3. Trickling biofilters The odour-laden air stream is passed over a microbial consortium immobilised on support media with a high surface area. Recirculating water maintains humidity in the media bed and allows nutrient supply. Odorants dissolve into the aqueous phase and are degraded by the biofilm present. Foul air can pass through a trickling biofilter either co- or counter-currently to the liquid that provides the biofilm with nutrients. Trickling filter media can be ceramic or plastic structures, activated carbon, celite, or mixtures of materials (Pedersen and Arvin, 1995; Shinabe et al., 1995; Weber and Hartmans, 1996; Kennes and Thalasso, 1998). Filters in which the foul air is recycled (to allow maximum dissolution of the odorants into the liquid and biofilm), moderate dissolved oxygen (DO) is maintained throughout, and the wastewater does not short-circuit through the media are the most effective for odour treatment. One example of a trickling filter successfully treating odorous air employs a pumping station to transfer the foul air from the covered headworks and primary clarifiers to the bottom of a bed of plastic media (Fig. 3). As the air passes through the media, it is collected from inside a dome at the top of the filter, the majority for recycling through the filter media and a small portion for final treatment in a mist scrubber. Trickling biofilters (Table 5) represent a method in which reaction products are washed out of the media and acidification can be avoided. The major drawback with this system is the problem of transferring the odorous pollutants from the foul air to the liquid phase, but trickling biofilters can still be effective in treatment of gaseous compounds with an air/water partition coefficient of less than 0.1 (Kennes and Thalasso, 1998). The wet area (the active area) of a filter is usually less than 50% of the total specific area (Green and Maloney, 1997), a figure that can be improved by increasing the

Fig. 3. Trickling biofilter for odour control, Mesa, AZ (WEF/ASCE, 1995).

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Table 5 Summary of biotrickling filters for odour treatment Advantages

flexible design
Simple, Low capital and O&M costs, especially
where existing trickling filters can be used Can treat up to 500 ppm H S (Vincent
and Hobson, 1998) Acidification slower in nitrifying filters
(WEF/ASCE, 1995) and with calcareous 2

filter media (such as mussel shells Ð good buffering capacity to counteract acidification) (Vincent and Hobson, 1998).

Disadvantages

criteria still developing
Design Dissolution of gas into liquid is the
rate-limiting step, so long gas residence

times are required (Kennes and Thalasso, 1998), necessitating recirculation of foul air. Media require regular replacement (hence, high O&M costs). The best filter media require replacement frequently (Vincent and Hobson, 1998). Only approximately 60% hydrogen sulphide removal efficiency (WEF/ASCE, 1995) Increased structure maintenance (corrosion of concrete units) Accumulation of excess biomass in the media bed reduces the specific surface area and bed volume and causes pressure drop, resulting in performance fall-off (Kennes and Thalasso, 1998) or requiring control techniques which compromise long-term performance (Morgenroth et al., 1996; Thalasso et al., 1996).








flow rate of the liquid, although this action increases operation costs and in the case of a filter also being used to treat wastewater, the level of wastewater treatment will be compromised. However, it has also been shown that decreasing the liquid flow rate to the minimum required for microbial growth can result in more efficient gas treatment (Thalasso et al., 1996), and also avoid one operational problem specific to trickling biofilters Ð the accumulation of excess biomass in the media bed (Kennes and Thalasso, 1998). Excess biofilm can completely clog a filter bed, although this does not always occur and can also be prevented by liquid nutrient minimization and backwashing (Smith et al., 1996; Thalasso et al., 1996). However, long-term minimization of liquid supply leads to reduced microbial activity and gas treatment (Morgenroth et al., 1996; Thalasso et al., 1996), so backwashing is a better option for maintenance of filter media (Smith et al., 1996). Carbon, hydrogen, and oxygen are not usually the limiting nutrients present in wastewater, but supplies of nitrogen, phosphorus, or micronutrients (vitamins and trace elements) may require control in order to balance biomass growth and removal efficiency against media clogging. 4. Bioscrubbers Bioscrubbing has several advantages over media-based filtration (Table 6). The process is more easily controlled because the pH, temperature, nutrient balance, and removal of metabolic products can be altered in the water of the reactor (Smet and van Langenhove, 1998). Bioscrubbing is reliant on good gas dissolution, as it employs the absorption of

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Table 6 Summary of bioscrubbers for odour treatment biotechnologies Advantages

Bioscrubbing is more easily controlled because
the pH, temperature, nutrient balance, and






removal of metabolic products can be altered in the water of the reactor (Smet and van Langenhove, 1998). Removal of the products of pollutant degradation by washout (thus, avoiding inhibition of the biomass) (Kennes and Thalasso, 1998) Easy control of the liquid medium composition (hence, a very controllable unit process) (Kennes and Thalasso, 1998) Acclimation capacity of the biomass provides efficient degradation of the pollutants (Kennes and Thalasso, 1998).

Disadvantages




Reliant on good gas dissolution, thus, it removes only highly soluble contaminants efficiently (gaseous pollutants with an air/water partition coefficient of less than 0.01) (Kennes and Thalasso, 1998). Biomass growth has to be controlled to reduce solid waste output and to increase gas treatment efficiency (WuÈbker and Friedrich, 1996). Controlled inputs of phosphate and potassium in the liquid media required for efficient pollutant degradation, but this is not suitable for low concentration, wastewater treatment-generated odorants (WuÈbker and Friedrich, 1996).





pollutants into the aqueous phase in a gas/liquid exchange column, followed by degradation in a liquid phase bioreactor (Fig. 4). The liquid phase bioreactor effluent is recirculated into the absorption column, providing excellent gas cleaning of highly soluble pollutants. The disadvantages of this system are based on the requirement to dissolve the pollutants during the short residence time in the absorption column, a major consideration as many air pollutants and odorants are volatile and poorly water-soluble. In addition, bioscrubber biomass growth has to be controlled to reduce solid waste output and to increase gas treatment efficiency. Reduction of biomass can be done in two ways: (1) increase the requirement for maintenance energy by increasing the mean cell residence time and (2) decrease efficiency of energy generation for biomass growth by limiting nutrient supply. Sewage could be used as the liquid media, although this removes the operators' control over the nutrition available in the liquid phase (WuÈbker and Friedrich, 1996). Bioscrubbers are less popular than biofilters, probably due to this feature, although some examples of successful applications are reported in the literature (Kennes and Thalasso, 1998). Some newer applications of bioscrubbers have developed as it is possible to biologically desulphurise large volumes of gas (up to 2  106 m3/h) (Buisman et al., 1994), and bioscrubbers are one of the very few means of anaerobic gas treatment. 5. Membrane bioreactors In membrane bioreactors, the gaseous pollutants are transferred from the gas to the liquid phase (where they are degraded) via a membrane. Two membrane materials are available for treating odours: dense (e.g. silicone rubber) and hydrophobic microporous (e.g. polysulphone); dense materials are more selective and microporous materials more permeable (Reij et al., 1998). There are also two types of biomass available: fixed film cultures (biofilms) and suspended growth cultures. The membrane forms the gas±liquid interface and, therefore, its size can be closely controlled (Fig. 5). Passage of the contaminated air across a membrane

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Fig. 4. Principal features of bioscrubber design (Kennes and Thalasso, 1998).

surface allows passive diffusion of contaminants through the membrane into the liquid phase on the other side, driven by the concentration gradient (Reij et al., 1998). The mass transfer coefficient inside a dense membrane depends on the solubility and diffusivity of the compound in the material matrix. Solubility and diffusivity vary between each contaminant, and mass transfer resistances of different membrane materials for the same compound vary, so waste gas compounds can be selectively extracted from or retained within the gas phase by careful choice of membrane material (Reij et al., 1995, 1998), presenting an advantage over thin film and bubble diffusion in which it is not possible to select removal of certain components of the foul air. The presence of the membrane prevents microorganisms from contaminating the gas phase, thus, such systems have been tested for treatment of indoor air (Freitas dos Santos et al., 1995) or air in a manned space cabin (Binot et al., 1994). The gas can be used as either the carbon or nitrogen source for the microorganisms in the liquid phase. While membrane bioreactors for odour treatment have yet to be tested at full scale, a number of features have emerged from the lab- and pilot-scale work carried out to date (Table 7). 6. Activated sludge There are fewer examples of liquid-based odour control systems than media-based systems (WEF/ASCE, 1995). The comparative merits of liquid- and media-based systems

Fig. 5. Principle of membrane bioreactor treatment of contaminated air using either a fixed film or suspended growth reactor.

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Table 7 Summary of membrane bioreactors for odour treatment biotechnologies Advantages

Disadvantages

Water phase allows optimal humidification of to be proven on any system other than lab-scale
the
Yet biomass and removal of the products of Rate of transmembrane transport depends strongly on
metabolism, avoiding acidification/inactivation the partition coefficient of the contaminant and its rate of the biomass.

of degradation in the liquid phase, therefore, use may be limited to highly soluble pollutants (Reij et al., 1995). Low removal efficiencies with poorly soluble pollutants (58% removal of propene). Propene is poorly soluble and (Ergas et al., 1999). most resistance to mass transfer was in the liquid phase The gas flow rate and liquid substrate flow (Reij et al, 1997). A total of 35% trichloroethylene removal rate can be controlled independently. most resistance to mass transfer was in the liquid phase Waste gas compounds can be selectively with a polysulphone hollow fibre membrane bioreactor extracted from or retained in the gas phase with a polysulphone hollow fibre membrane bioreactor by careful choice of membrane material (Parvatiyar et al., 1996b). (Reij et al., 1995). >95% removal of toluene and dichloromethane Very high capital cost (Ergas et al., 1999) No routine removal of biomass, so there is a possibility of excess solids retention within the reactor. Potential for 97% toluene removal with a polyethylene hollow fibre membrane bioreactor, but removal declines over time, as the biofilm increases in thickness (Ergas and McGrath, 1997).

Nutrients, substances for cometabolism, and
buffers can be easily applied to the biomass














differ, and so their suitability to the conditions in different wastewater treatment plants also differs (Table 8). Activated sludge diffusion is used as an alternative to more established bioreactors for waste gas treatment, such as biofilters, bioscrubbers, and biotrickling filters. Contaminant removal mechanisms in activated sludge diffusion of waste gas include absorption (the solution of gases into the mixed liquor; limited by bubble size and gas residence time), adsorption (high RMM compounds with low solubility adsorb onto flocs), or condensation (volatile organic compounds in warm air condense on contact with the cooler mixed liquor), followed by biodegradation (Bowker, 2000a). Foul air is collected from its source and transferred via blowers through a delivery pipework system to submerged nozzles in the activated sludge aeration tank (Fig. 6). The odorous air bubbles diffuse into the mixed liquor, where the contaminants dissolve and are subsequently adsorbed onto the floc or absorbed into bacterial cells and biodegraded. A number of workers have reported results for removal of odorous compounds by the sparging of gas through a suspended microorganism culture. Table 9 shows results obtained for removal of H2S and other odorous compounds from gas streams by suspended culture microbial processes. 6.1. Operational considerations The problems reported in using activated-sludge treatment of waste gases are varied. Primary issues with treating odorous air using such systems are corrosion of pipework and air blowers by the moist, acidic air, the transfer of the foul air from the gas to the liquid phase,

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Table 8 Advantages and disadvantages of liquid- and media-based odour control biotechnologies Liquid-based systems Advantages

Disadvantages

of existing facilities (no footprint)
Use treatment of large volumes
Economical pressure requirements in trickling filters
Low operation
Simple No chemical
Gas dissolutionrequirements is the rate-limiting step.
Ability to treat odorants other than
H S limited can be difficult to control.
Process
Questionable consistency of performance 2

Media-based systems

to treat a wide range of odorants
Ability capital and operating costs
Low operation
Simple chemical requirements
Minimal No sidestreams
Large footprint requiring treatment
High energy demand to force air
through media can be difficult to control.
Process in foul air plug media
Particles and pH control more difficult
Moisture Downtime required for media replacement


and the potential toxic effects of the odorants once in solution (Ryckman-Siegwarth and Pincince, 1992; Bowker, 1999). 6.1.1. Corrosion Concrete and carbon±steel items suffer corrosion from hydrogen sulphide and sulphuric acid (Ryckman-Siegwarth and Pincince, 1992; WEF/ASCE, 1995; Bowker, 2000a), however, PVC, fibreglass, and stainless steel are all suitable for the foul air delivery system (RyckmanSiegwarth and Pincince, 1992; Bowker, 2000a; Oppelt et al., 1999). Some sites reported corrosion of the concrete aeration tank, ameliorated by the provision of a protective coating at the waterline (Ryckman-Siegwarth and Pincince, 1992). Stainless steel and PVC have been shown not to be affected by  100 ppm hydrogen sulphide over a 2-year period of operation, and the provision of effective moisture traps prevents condensation of sulphuric acid in blowers and pipework, avoiding pitting of metal parts (Bowker, 2000a). Corrosion of diffusers (both coarse and fine bubble) was not found to be a significant problem in a survey of

Fig. 6. Activated sludge diffusion of contaminated air.

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Table 9 Performance of suspended growth odour control systems Substance

Reactor type

Level of removal (%)

Loading rate (g/kg MLSS/day)

Microorganism

Reference

H2S

Activated sludge

99

< 15

Mixed culture

(CH3)2S

Activated sludge

99

<9

Mixed culture

Trimethylamine

Activated sludge

99

168

Mixed culture

Monoethylamine

Activated sludge

99

192

Mixed culture

N-butanol (n-BA)

Activated sludge

±

130

Mixed culture

Isopropanol (I-PA) Activated sludge

99.5

94

Mixed culture

BTEXa

Activated sludge

>99

Mixed culture

H2S

Activated sludge

83

15 ± 17 BTEX/L day

H2S

Activated sludge

96

7

Mixed culture

H2S

Activated sludge

95

15

Mixed culture

Dimethyl sulphide Activated sludge

35

Carbon disulphide Activated sludge

33.9

Ammonia

Activated sludge

96

0.20 mL/L, when Mixed culture AIb = 30 m3/m3/h 0.059 mL/L, when Mixed culture AI = 30 m3/m3/h Mixed culture

H2S

Activated sludge

99

Mixed culture

Methylmercaptan

Activated sludge

99

Mixed culture

Methyl disulphide Activated sludge

98

Mixed culture

Fukuyama et al., 1979 Fukuyama et al., 1979 Fukuyama et al., 1980 Fukuyama et al., 1980 Fukuyama et al., 1981 Fukuyama et al., 1981 Bielefeldt et al., 1997 Stillwell et al., 1994 Fukuyama et al., 1986 Fukuyama et al., 1986 Fukuyama et al., 1986 Fukuyama et al., 1986 Kasakura and Tatsukawa, 1995 Kasakura and Tatsukawa, 1995 Kasakura and Tatsukawa, 1995 Kasakura and Tatsukawa, 1995 Oppelt et al., 1999

Mixed gasc

H2S H2S (anoxic) H2S (aerobic) H2S H2S

Activated sludge 92.6 ± 93.7 basin Ð chemical manufacturing plant Bioscrubber 99

Mixed culture

Mixed culture

2000 ppm

Suspended culture Breakthrough 5.4 ± 7.6 mmol H2S/h g biomass Suspended culture Breakthrough 15.1 ± 20.9 mmol H2S/h g biomass Suspended culture 100 2.9-g dry cell weight/day Suspended culture 99.9

Mixed culture T. denitrificans T. denitrificans Thiobacillus thioparus Chlorobium limicola forma thiosulfatophilum

Nishimura and Yoda, 1997 Sublette et al., 1998 Sublette et al., 1998 Kanagawa and Mikami, 1989 Cork et al., 1983

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wastewater treatment plants employing activated sludge odour diffusion (Ryckman-Siegwarth and Pincince, 1992). Diffusion minimises the amount of equipment involved in introducing the air into the sludge, but could increase the amount of system maintenance required. 6.1.2. Deposition Accumulation of a `tar-like substance' on diffuser nozzles and internal components of air blowers has been reported (Ryckman-Siegwarth and Pincince, 1992), but it is also reported that such deposits are easily removed by household oven cleaner, or steam cleaning. Positioning of the foul air intake higher than the splash zone above the surface of the liquid/sludge generating the odour minimises this problem (Bowker, 1999). Blowers may require periodic cleaning (every 2±3 years) to remove grease and grit. A particulate filter and a moisture trap at the foul air collection point prevent most of the corrosion/deposition problems in the air collection and delivery system (Ryckman-Siegwarth and Pincince, 1992; Bowker, 2000a,b). 6.1.3. Increased odour emission Activated sludge diffusion of odorous air reduces the presence of liquid phase odorants via biological oxidation, but can produce odours via gas stripping, especially where systems become overloaded (Ryckman-Siegwarth and Pincince, 1992; Frechen, 1994; Vincent and Hobson, 1998). This is not a significant operating problem, however, as in most cases there is no detectable difference between the odour off an activated sludge plant treating odorous offgas and the odour from an activated sludge plant operating `normally' (Bowker, 2000a). In general, odours are significantly reduced at all wastewater treatment sites, provided sufficient DO is maintained in the mixed liquor (Ryckman-Siegwarth and Pincince, 1992), as the odour monitored at site boundaries is the product of the entire site as opposed to the activated sludge tanks alone. Case studies of full-scale sites using activated sludge diffusion (Ryckman-Siegwarth and Pincince, 1992) found that aerating with offgas from grit chambers and primary clarifiers and fine bubble nozzles affected the tank air emissions, effluent concentrations, and the quantity of volatile organic compounds biodegraded. The odour emission from the aeration tank increased, but the emissions from the site as a whole decreased, owing to the odours from the grit chambers and primary clarifiers being eliminated. The concentrations of volatile organic compounds emitted to the environment via the reactor effluent increased, but the total emissions from the site decreased as a substantially higher proportion of the volatile organic compounds received by the site were biodegraded. The use of foul air for aeration carries many advantages for sites at which all emissions to the atmosphere must be treated before discharge. Notes to Table 9 a Gas containing benzene, toluene, ethylbenzene, and/or xylenes. b AI: Aeration intensity, calculated according to Fukuyama et al. (1986) as cubic meter of gas entering tank per cubic meter of deodorising tank volume per hour. c The mixed gas contained benzene, chloroethane, chloroform, ethyl benzene, hexane, toluene, 1,2,4trimethylbenzene, vinyl acetate, m-xylene, o-xylene, and p-xylene.

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6.1.4. Treatment capacity Activated sludge diffusion of odorous air works well in situations where the activated sludge plant is not heavily loaded and DO levels are maintained, and has been in use by the Los Angeles County Sanitation Districts, California, USA for many years (WEF/ASCE, 1995). The degradability of odorants other than hydrogen sulphide is in doubt (Bielefeldt et al., 1997), although the performance data presented later in this review contest this statement. If it were to be the case, activated sludge tanks could still be employed as primary odour treatment units, designed for maximum hydrogen sulphide removal. However, one stated problem is that aeration tanks cannot always accept the total volume of foul air generated at a wastewater treatment works (WEF/ASCE, 1995). A survey of several North American full-scale treatment plants using activated sludge foul air diffusion assessed the extent of the operational problems experienced with this form of odour treatment (Ryckman-Siegwarth and Pincince, 1992). Foul air was found to account for 20±100% of aeration air supply (Ryckman-Siegwarth and Pincince, 1992), but in no case was an excess of foul air reported. 6.1.5. Effects on wastewater treatment Introduction of odorous air into heavily loaded activated sludge plants can cause loss of process performance, although foul air drawn from waste treatment processes such as activated sludge compost systems can be high in oxygen, thus providing an advantage (WEF/ASCE, 1995). Activated sludge plant operation is affected by the amount of sulphide entering the reactor. All sulphurous compounds, anilines, and phenols are inhibitory to nitrification (Henze et al., 1995); if the prevailing pH drops below 7, then nitrification declines (ásùy et al., 1998). It is known that the H2S removal was significantly influenced by the pH in the range of 3.0±5.5, and less affected by the pH in the range of 5.5±7.5. The maximum sulphide removal was found to occur at the pH level of around 4 (Li et al., 1998). Yang and Allen (1994) also carried out work on the use of biofilters for H2S removal and noted that removal efficiency was highly dependent on pH below 3.2 but was almost independent of pH at higher values. Sublette et al. (1998) reported the fact that as the pH shifts to a higher value, there is a propensity for dissolved H2S to remain in solution as the disulphide ion. Hydrogen sulphide input into activated sludge either via air or wastewater has been seen to result in nitrification inhibition and sludge bulking (Bentzen et al., 1995, ásùy et al., 1997). Activated sludge nitrification decreased by 27%, 67%, and 76% when sulphide concentrations are found to be present in wastewater at 1, 5, and 10 mg/L, respectively (ásùy et al., 1998). Sulphide inhibition depends on the composition and acclimation of the biomass, the concentration of hydrogen sulphide and other components in the wastewater, and temperature (as it affects solubility and bacterial growth rates) (ásùy et al., 1998). The efficiency of sulphide oxidation by T. denitrificans was noted by Sublette et al. (1998) to drop significantly at low temperatures and became inhibited below 15.6°C. The effect of temperature on sulphide removal capacity was also studied by Buisman et al. (1990). Experiments conducted on biomass acclimatised to a temperature of 20°C showed that the optimal temperature range was located between 25°C and 35°C, but at 5°C, up to 30% of the maximum sulphide oxidation capacity was reached.

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Laboratory scale activated sludge reactors showed higher concentrations of filamentous bacteria (responsible for sludge handling problems) when high loadings of hydrogen sulphide were applied (Johnson et al., 1995). Increases in the levels of filamentous bacteria present in the mixed liquor of full scale activated sludge plants accepting foul air have been reported, but no cause and effect has been identified by the wastewater treatment plants allegedly experiencing this problem (Bowker, 2000a). This may be due to the fact that sulphide concentrations in wastewater are usually higher than hydrogen sulphide concentrations in foul air, and the impact of the gaseous sulphide load is therefore less than that of the aqueous load. A number of researchers have made reference to sulphide as an inhibitor of nitrification (ásùy et al., 1998; Nielsen and Keiding, 1998; Beccari et al., 1980), however, there are few documented reports of these effects. ásùy et al. (1998) found that a sulphide concentration of 0.5 mg S/L in the influent wastewater caused a considerable reduction in the nitrification rate. The extent of sulphide inhibition is supposed to be dependent on the composition of the biomass, degree of acclimatisation, the concentration of sulphide, and the content of other inorganic and organic compounds in the wastewater. The toxic effect may be particularly harmful at low temperature because of the low growth rate of nitrifying bacteria (ásùy et al., 1998). ásùy et al. (1998) do however suggest that since both the presence of biodegradable organic matter and the sulphide influenced the nitrification rate negatively, it is difficult to single out the effect of the sulphide alone. In experimental work on a biofilter carried out by Pomeroy (1982), it was noted that while nitrification was occurring, a better removal of H2S was accomplished. This correlation may be the result of nitrification and H2S degradation both indicating optimum operating conditions rather than a cause and effect relationship. ásùy et al. (1998) reported that settling efficiency of suspended solids was reduced when high sulphide loads were present. This was attributed to either microbial activity (hydrolysis) or the types of organism present, for example filamentous or nonfloc forming. Nielsen and Keiding (1998), WileÂn et al. (2000), and dos Santos Afonso and Stumm (1992) found that the addition of aqueous sulphide to activated sludge led to a change in the floc structure due to a specific reduction of Fe3 + to FeS. The change was observed as a weakening of the floc strength leading to an increased shear sensitivity. The degree of disintegration was thus dependent on the shear stress within a particular system. dos Santos Afonso and Stumm (1992) claim that the rate of dissolution is dependent on the concentration of sulphide, the dissolution promoting species, on the surface of the floc. The observation of these changes has important implications on the settleability of the sludge and, thus, the efficiency of the process. The breakup of the floc occurs because Fe3 + has better flocculating properties than Fe2 + (in the form of FeS) due to its higher valence. Nielsen and Keiding (1998) suggest that the reduction of Fe3 + may take place via microbial reduction or by chemical reduction with sulphide. In this paper, it was shown that floc strength decreased with increasing sulphide levels and that the decrease was proportional to the amount of Fe3 + being reduced. When 60% of all sulphide reducible Fe3 + was reduced, 5±10% of the floc material was found in the bulk water, resulting in a highly turbid final effluent (Nielsen and Keiding, 1998). Experimental work using a bubble column (Lee and Sublette, 1993) showed that when the sulphide concentration in the feed water was about 19.6 mM, the floc began to change gradually from floc-type to granular-type. The size of the granular particles was then seen

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J.E. Burgess et al. / Biotechnology Advances 19 (2001) 35±63

to increase as the sulphide concentration in the water increased. A maximum diameter of 5 mm was obtained for the granular-type particles. The solid content of the waste sludge is determined by the ability of the sludge to yield water within a limited time scale by mechanical means (Mikkelsen et al., 1996). Disintegration of the activated sludge floc will affect the dewaterability of the excess sludge and, therefore, may have economic implications. General effects of aeration with foul air from wastewater treatment on the activated sludge were noted throughout a pilot-scale trial of odour diffusion (Fukuyama et al., 1986). The mixed liquor pH remained constant (around neutral), the mixed liquor suspended solids (MLSS) decreased, effluent suspended solids increased, and changes were seen in the community structure of the biomass (decreased numbers of Protozoa and increased Euglypha). The authors also observed the effects of treating night-soil odorous air on activated sludge. They found that pH was reduced from 7.6 to 3.15 over one 33-day experiment, and from 6.3 to 5.5 over 22 days, with consequential effects on nitrification. Autolysis of the biomass at pH 3.15 led to ammonia in the reactor effluent in excess of the influent ammonia concentrations. Mixed liquor volatile suspended solids (MLVSS) and MLSS decreased by 130±160 mg/L, and the sulphur and nitrogen content of the biomass increased during the experiments. A total of 22±39% of the sulphur present was metabolised to SO42 ÿ , and no residual sulphides were measured in the wastewater. An economic analysis of one installation of a dedicated air diffusion system for odorous gas treatment in an activated sludge tank (previously aerated by mechanical surface aerators, at Valley Forge WWTP, Phoenixville, PA; Bowker, 2000a) showed that the capital expenditure required was significantly less (36%) than for a wet scrubber at the site. Annual power consumption costs of the wet scrubbing system examined were 17% less than those of diffusing waste gas, although this cost can be offset by the reduction in power required for the mechanical surface aeration system, a saving that could not be calculated prior to the installation and, hence, was not included in the economic analysis. Wastewater treatment plants, which already have a diffuser system for aeration, need only make minimal capital expenditure (adding pipework and moisture traps, replacing mild steel with stainless steel, or PVC) to begin activated sludge diffusion of waste gas, making it more economically viable (Bowker, 2000a). In treating foul air from waste treatment plants, some control of pH and MLSS would be required to maintain ongoing performance (Fukuyama et al., 1986). The constant input of new wastewater and routine sludge surplussing and recycling will provide this control and avoid the accumulation of toxic metabolites. It has been found that sulphurous compounds are converted to sulphate and partly taken up by the sludge; nitrogenous compounds are converted to nitrate and nitrite. High loading rates, plus a pH of >5.0, facilitate nitrification of ammoniacal-N, and as no nitrate or nitrite was found in the effluent or MLSS supernatant, denitrification was also occurring. Fukuyama et al. (1986) also concluded that activated sludge treatment is a feasible option for treatment of odours from waste treatment sources, especially where an activated sludge plant already exists. It has been stated that activated sludge diffusion of odorous air is useful only where the sludge is fully aerobic (preferably nitrifying) and the concentration of hydrogen sulphide to

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53

be treated is relatively low (Vincent and Hobson, 1998), as the concentration of oxygen required to oxidise H2S to H2SO4 is high. However, wastewater treatment plants have been known to successfully operate diffusion of waste gases at  100 ppm for long periods of time (Bowker, 2000a), so secondary treatment systems, such as activated sludge offer an economically viable solution to the problem of odour generation from other parts of the wastewater treatment plant, however, there is a paucity of published information regarding this application. However, the operational difficulties outlined can be and have been overcome at several sites, examples of which are summarised in Table 10. Activated sludge bioscrubbing has been successful at two Los Angeles, CA wastewater treatment plants, the much-cited Hyperion Treatment Plant and the Donald C Tillman Water Reclamation Plant (Bielefeldt et al., 1997; Bowker, 1999, 2000a; Ostojic et al., 1992; Stillwell et al., 1994). Vented air from odorous sources can be routed through the aeration tanks of an activated sludge plant in addition to fresh air, or instead of fresh air in cases where sufficient foul air is generated to provide aeration (Ryckman-Siegwarth and Pincince, 1992). Activated sludge diffusion effectively removes odour and degrades volatile organic compounds. The sites summarised (Table 10) successfully employ activated sludge diffusion odour treatment at full scale, but there are limited performance data available reporting the effects of various factors on process effectiveness. 6.2. Performance data One potential disadvantage of activated sludge diffusion is the need for a deep aeration tank to provide a long gas residence time, when a shallow reactor represents a reduction in energy requirements. However, shallow activated sludge basins have been shown to effectively degrade a mixture of benzene, toluene, ethylbenzene, and xylene (BTEX) (Bielefeldt et al., 1997). A bench-scale activated sludge reactor with a working volume of 2 L and liquid depth of 40 cm was run with sludge ages of 1.7, 2.7, and 9.2 days [hydraulic retention time was equal to sludge retention time (SRT)] and 15±17 mg/L BTEX in the air entering the reactor. The BTEX in the offgas was below the limit of detection (0.01 mg/L), indicating > 99% removal in all cases and showing that shallow activated sludge tanks are able to biodegrade BTEX in contaminated air (Bielefeldt et al., 1997). In further studies on the effects of mixed liquor depth on odour treatment, a pilot activated sludge plant was run to treat foul air from the headspace of a dissolved air flotation sludge thickener (Bielefeldt et al., 1997). The 35-L working volume reactor held 127-cm depth of activated sludge, with 250 mg/L MLVSS. The contaminated air contained low levels of hydrogen sulphide, amines, ammonia, and mercaptans, all of which were removed to < 0.1 ppm in the tank offgas. Reducing the height of the liquid to 60 cm had no effect on the levels of contaminants present in the effluent gas (Bielefeldt et al., 1997). As in wastewater treatment, acclimation of the activated sludge was crucial for effective gas treatment, as unacclimated activated sludge gas treatment was biodegradation limited, removing only  45% of some contaminants. Longer SRTs allowed degradation of indole and skatole. Bubble size is also an important factor in gas treatment efficiency. Coarse bubble diffusers employed at Concord, NH wastewater treatment plant were positioned in the aeration basins

± ± ± ± ± ± ± ± 0.05 0.1

0.12 ± ± ± ± ± ± ± 0.1 0.1

5.6 ± ± ± ± ± ± ± 77

± 18 8 30 83 180 240 130 5 30

16 74 9 27 49 130 230 110 5 7

In-vessel composting reactor

Influent, clarifiers, effluent launders

Springfield Regional, Springfield, MA

Valley Forge Regional, Phoenixville, PA

230 430 3500 19 000

3900

660 2600

0.2 0.2 ±

Background

1° clarifiers In-vessel composting reactor

7.8 0.3 <1

Out

Conserv II, Orlando, FL Reedy Creek, Orlando, FL

 100  100 6

In

10 14 ±

Background

In

Out

H2S (ppm)

Odour detectability (OU) 850 18 ±

Odour source

Sludge holding tank 24 150 39 000 ± Hyperion, Los Angeles, CA Headworks, sludge handling, effluent pumping

Concord, Concord, NH

Location

Table 10 Performance of activated sludge diffusion for odour control (Ryckman-Siegwarth and Pincince, 1992; Bowker, 2000a)

Started 1988. No corrosion reported. Coarse bubble diffusers. Odour control collection and delivery system separate from fresh air system. No corrosion or deposition problems. Synthetic resin coatings on impellers. Coarse bubble diffusers.

Coarse bubble Fine bubble Blowers cleaned twice (1959 ± 1992). No corrosion reported. Fine bubble diffusers. 96 ± 99% odour removal, 80% VOC removal.

Comments

54 J.E. Burgess et al. / Biotechnology Advances 19 (2001) 35±63

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at 3-m depth (Bowker, 2000a). These diffusers provided 95% odour reduction (measured by olfactometry) and 92% hydrogen sulphide reduction (  100 ppm inlet concentration), but the odour from the activated sludge plant was detectably higher than before activated sludge diffusion of foul air began. Upon changing the system to fine bubble diffusers, a > 99.5% reduction in both odour and hydrogen sulphide was achieved, the activated sludge plant odour being indistinguishable from its odour prior to waste gas treatment (Bowker, 2000a). Much of the information available on activated sludge diffusion is in the form of qualitative, almost anecdotal reports by site operators, with little information available regarding the activated sludge plant types and operating conditions leading to good odour treatment. Experiments at lab-scale have shown that activated sludge tanks can effectively degrade sulphurous compounds (Fukuyama et al., 1979), aliphatic amines (Fukuyama et al., 1980), toluene (Fukuyama and Honda, 1976), and low relative molecular mass compounds (Fukuyama et al., 1981). Lab-scale experiments using odorous air rather than sample odour gases achieved  99% removal efficiencies of these compounds; the pilot plant approximately 90%. Later work at pilot-scale (Fukuyama et al., 1986) reported that the pilot plant coped well with variable loads, which had not been an issue in the lab experiments. The wastewater treatment plant in the study received > 25% of its load from industrial sources, so has a high proportion of odorous components in the liquid phase. During one study, continuous activated sludge tank deodorization of exhaust gas from wastewater treatment and night-soil treatment plants was carried out for several months (Fukuyama et al., 1986). Efficiency was measured in terms of the concentrations of the main odorants prior to and after treatment; influent concentrations varied greatly, but outlet concentrations were more consistent. Mean removal efficiencies were 90% for aromatic hydrocarbons and dimethyl sulphide, 96% for hydrogen sulphide (mean influent concentration of 7 mg/g MLSS/day), and 100% for ammonia (Fukuyama et al., 1986). Investigation into the effect of aeration intensity employed an aeration tank (depth 1.0 m, working volume 150 L, MLSS 11.20 g/L, SRT of 1) received two levels of aeration intensity: 6 and 12 m3 air/m3 tank volume/h. Aeration intensity was found to affect the degree of removal of volatiles by gas stripping. Total aromatic hydrocarbons, dimethyl sulphide, and carbon disulphide in the activated sludge plant offgas decreased when the aeration rate of the plant was increased from 6 to 12 m3 air/m3 tank volume/h and Fukuyama et al. (1986) concluded that the decrease in these measured components at the higher aeration intensity (except carbon disulphide, which did not contribute to odour, as the outlet concentration was below the limit of human detection) and simultaneous increase in odour units leads to the conclusion that increased aeration causes other, unmeasured odorants present in the wastewater to be stripped from the liquid phase. Night-soil treatment plant foul air was also treated using an activated sludge aeration tank and odorant removal was compared to a control tank filled with clean water (Fukuyama et al., 1986). The first run was carried out using an aeration intensity of 4.7 m3 air/m3 tank volume/ day, MLSS of 16.28 g/L, and SRT of 1. The second run was carried out using a lower loading rate, finer bubbles, and an aeration intensity of 2.0 m3 air/m3 tank volume/day, MLSS of 15.55 g/L, SRT of 1, and resulted in greater contaminant removal efficiency than the first run. Increases in removal efficiency ranging from 0.80%, for ammonia, to 44.62%, for dimethyl sulphide, were recorded. The control tank attained comparable removal efficiencies

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when loading rates were consistent and normal, but did not remove peak loads, which were removed by the activated sludge tank (up to 0.58 ml H2S/ml air), resulting in a consistent outlet hydrogen sulphide concentration from the test aeration tank (Fukuyama et al., 1986). This indicates that a proportion of the odorants are dissolved in the liquid and go no further, but that much of the odour removal reported is dependent on biodegradation to avoid saturation of the liquid present with the odour compounds. Fukuyama et al. (1986) established the relationship between hydrogen sulphide loading and removal rates as: y ˆ ÿ0:981x ‡ 99:26

…1†

where y = removal efficiency (%); x = hydrogen sulphide load (mg/g MLSS/day). There must be a threshold top loading rate at which Eq. (1) ceases to be true, as hydrogen sulphide exerts a toxic effect on biomass when present in excess. This threshold value clearly exceeds 7 mg H2S/g MLSS/day, the mean loading rate applied by Fukuyama et al. (1986), but has yet to be established. Fukuyama et al. (1986) studied the effectiveness of a two-stage diffusion process. Two identical aeration tanks (depth 1.0 m, working volume 150 L, MLSS 8.82 g/L, SRT of 1, aeration intensity of 30 m3 air/m3 tank volume/day) were used in series. The results showed large standard deviations in the extra removal obtained in the second stage (e.g. extra removal figures between 0% and 40% for carbon disulphide removal) and low values for extra contaminant removal (up to 94.38% dimethyl sulphide removal in the first stage compared with 4.89% removal in the second stage), which mean that the cost of duplication in adding a second aeration tank is not justified in most cases. The use of two-stage treatment is most useful for very variable loads of airborne contaminants, instead of recycling the outlet air (Fukuyama et al., 1986). Investigation into the effects of sludge reaeration was carried out, using one aeration tank (depth 1.0 m, working volume 150 L, MLSS 4.65 g/L, two different SRTs, 1 and 4 h) and an aeration intensity of 30 m3 air/m3 tank volume/day (Fukuyama et al., 1986). The removal efficiencies for the test compounds (carbon disulphide, dimethyl disulphide, and total aromatic hydrocarbons) obtained varied from 9.26% to 35.00%, with an SRT of 1 h, and from 0% to 21.05%, when SRT = 4 h, indicating that increasing SRT in this experiment led to decreasing removal efficiencies. However, the removal efficiencies at both SRTs were very low in comparison to the authors' other experiments, in which SRT = 1. This leads to the conclusion that the SRTs were both too low to compare to the typical SRTs used in wastewater treatment (6±10 days for domestic wastewater, longer for industrial effluent; Eckenfelder and Grau, 1992) and the data are not representative of the effects of SRT on odour treatment operated in a `real' system. Loading rates of 15 mg H2S/g MLSS/day were degraded very well (  95% removal efficiency) in the laboratory. The pilot plant was subject to variable loading rates and the effects of other odorants not present in the laboratory experiments, but still achieved  90% removal efficiency up to 7 mg H2S/g MLSS/day ( < 90% over mg H2S/g MLSS/day). The order of removal by the sludge was: hydrogen sulphide > methyl mercaptan > dimethyl sulphide > dimethyl disulphide. Removal efficiencies may suffer when peak concentrations occur, as foul air compounds have to be acclimated to just as wastewater components do, but this rarely affects aeration tank outlet concentrations of odorant (Fukuyama et al., 1986), and

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particularly high concentrations can be degraded by recycling the air during peak loads (Frechen, 1994). Performance data relating to volatile organic compound removal are reported by Oppelt et al. (1999) operating an activated sludge plant (liquid depth 6.6 m, MLVSS 2242 mg/L, DO 2.0 mg/L) through which headspace air from a lift station was diffused, mixed with fresh air. The wastewater passing through the aeration tank contained concentrations of the volatile organic compounds present at the wastewater treatment site far in excess of the concentrations measured in the headspace air, so it was not possible to calculate volatile organic compound removal efficiencies during normal operation of the plant. Instead, the system's effectiveness was measured during the construction phase, with no wastewater flowing through the activated sludge plant. The aeration tanks were filled with mixed liquor from another activated sludge tank and allowed a 2-week period for acclimation to the foul air contaminants, after which the aeration tank headspace was sampled. Eleven volatile organic compounds entered the aeration tank (Table 11) and were biodegraded, but the variation in the aeration tank emission data is so great that longer-term results, generated with a working system are still required to build on these very promising data. A modelling software package (Toxchem+) was applied to the design of the process and was shown to consistently overestimate the degradation rates obtainable by this aeration tank (Oppelt et al., 1999), indicating that currently available modelling packages are not accurate when applied to activated sludge odour treatment and that more developmental work is required for a full understanding of the reactor and microbial kinetics involved in activated sludge treatment of odorous gases. 6.3. Advantages over media-based systems Odour control for offgas from sludge composting was studied and the methods of wet scrubbing, biofiltration, and activated sludge diffusion were compared (Ostojic et al., 1992). Wet scrubbing is one of the most popular methods of odour control in the USA, but Table 11 Lift station and aeration tank volatile organic compound emission data (Oppelt et al., 1999) Compound

Benzene Chloroethane Chloroform Ethyl benzene Hexane Toluene 1,2,4-Trimethylbenzene Vinyl acetate Xylene o-Xylene

Lift station headspace air concentration (mg/m3)

Aeration tank headspace air Mean concentration (mg/m3)

Standard deviation

Relative standard deviation (%)

315 912 596 1589 20 059 11106 4097 82 308 4325 9708

20 81 84 14 3815 121 30 45 17 21

7.9 40.4 19.4 12.5 2917.9 167.5 32.5 40.7 12.7 20.0

39.5 49.9 23.0 87.0 76.5 138.3 110.7 89.4 74.7 95.5

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suffers from recurring problems and averages 70±75% removal efficiencies. Biofiltration using compost or wood chips averages around 90±95% removal efficiency, and has replaced wet scrubbing at some sites, but suffers badly when media humidification fails (45% removal efficiency). Activated sludge averages  100% removal efficiency at full scale, where 2.0±2.5-m depth of MLSS is maintained (Springfield, MA and Orlando, FL), and even improve on the original level of background odour in cases where surface mechanical aerators are replaced by submerged nozzle when the foul air diffusion system is fitted. Activated sludge outperforms wet scrubbing for treatment of air from sludge composting, as the air contains a number of odorants in addition to sulphurous compounds (alcohols, ketones, aldehydes, acids), which are biodegradable but which persist after wet scrubber treatment. Activated sludge diffusion avoids many common problems with biofilters, biotrickling filters, and membrane bioreactors, i.e. media plugging, excess biomass accumulation, gas short-circuiting, moisture control, and maintaining the correct biofilm thickness (Bielefeldt et al., 1997). The advantages and disadvantages of activated sludge diffusion are summarised in Table 12. Any filter consisting of a bed of media has to be supplied with prehumidified waste gas to prevent dehydration of the filter microorganisms. The biofilter and trickling biofilter both consist of a packed-bed of media onto which water is sprayed and the odorants then diffuse into the thin water layer within the filter from which they are taken up by microorganisms. Pollutants with low water solubility may not diffuse into the thin layer, as the water surface area is small by comparison to the area available in activated sludge diffusion using small bubbles (Stillwell et al., 1994). Filter biotreatment of gases containing chlorinated pollutants, sulphur compounds, or ammonia results in accumulation of chloride, sulphate, or nitrate ions and subsequent acidification of the biofilter; acidification can be buffered by chemical additions, such as lime, but the mineral end products can neither be neutralised in, nor removed from the filter. The use of the microorganisms in suspension in the liquid means that toxic end products are removed from the liquid phase as components of the reactor effluent or as solids incorporated into the biomass removed for disposal. Humidity does not require control, the volume of mixed liquor stabilises the reactor temperature, and nutrients are supplied in the wastewater. Activated sludge systems represent low-cost solutions for similar performance as membrane bioreactors, which require the installation of a unit process, thus carrying considerable capital costs, and, which have yet to be proven for odour treatment at pilot- or full-scale. 7. Conclusions and research needs Activated sludge treatment of odorous air has been increased over the last 25 years. Ryckman-Siegwarth and Pincince (1992) and Bowker (1999, 2000a) have shown that aeration tanks can and have been successfully employed at wastewater treatment sites covering much of North America. The restrictions of corrosion of structures and deposition problems have been overcome. However, the longest trial reported in the literature ran for only 33 days (Fukuyama et al., 1986), and the longer term, full-scale results are mainly

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Table 12 Summary of activated sludge diffusion Advantages

and effective. No chemical requirements.
Simple Easily controlled (pH, temperature, nutrient balance)
via the water of a liquid reactor (Smet and van










Langenhove, 1998). Removal of excess biomass and the products of pollutant degradation by washout (avoids biomass inhibition). Biomass acclimation capacity provides efficient pollutant degradation (Kennes and Thalasso, 1998). Low O&M, low capital cost. Very low capital costs if existing blowers/diffusers are used. Use of existing facilities (no footprint) and equipment (operator familiarity with unit, minimal capital cost). Economical treatment of large volumes. Can treat up to  100 ppm H2S long term (Bowker, 1999). Can treat up to  100 ppm H2S long term (Bowker, 1999). Avoids the problems with biofilms, i.e. media plugging, excess biomass accumulation, gas short-circuiting, moisture control, and maintaining correct biofilm thickness (Bielefeldt et al., 1997). Very high removal efficiencies ( > 99.5%) obtainable (Bowker, 2000a,b). No detectable difference between the odour off an activated sludge plant treating H2S and the odour from a `normal' activated sludge plant (Bowker, 1999).

Disadvantages

Increased blower maintenance, and related to
deposition in the blower, increased power costs

associated with compression of gases required to force them through clogged blowers and diffusers. Media-based systems do not generally require the same level of gas compression. Gas dissolution is the rate-limiting step. Ability to treat odorants other than H2S limited. Process can be difficult to control, as composition of wastewater is not controlled. Some authors question consistency of performance (WEF/ASCE, 1995). Useful only where the sludge is aerobic, nitrifying, and the concentration of H2S is low (Vincent and Hobson, 1998). Overloaded systems can produce odours via gas stripping (Vincent and Hobson, 1998). Odorants inhibit nitrification (ásùy et al., 1998; Henze et al., 1995) H2S input may result in bulking sludge (Bentzen et al., 1995; Johnson et al., 1995; ásùy et al., 1997).











quantitative and concerned with overcoming operational restraints. Existing modelling packages were shown to consistently overestimate degradation rates (Oppelt et al., 1999), and while a model has been developed specifically for BTEX (Bielefeldt et al., 1997), the principles have yet to be applied to other odorous compounds. The threshold-loading rate at which Fukuyama et al.'s (1986) equation describing the relationship between hydrogen sulphide loading rate and removal efficiency no longer applies has not been established, and the masses of odorous compounds, a set volume of activated sludge can remove, are not known. The proportions of odour removal achieved by physicochemical and biological mechanisms are not yet established, nor are the effects of foul air diffusion on activated sludge chemistry and properties. Long-term influent and effluent quality data are required in order to study the performance of the simultaneous odour and wastewater treatment processes occurring in the activated sludge aeration tank. Many questions remain to be answered before models can be developed to take the variation in odour compounds and plant operating regimes into account, and diffusion treatment can be successfully designed, installed, and optimised at any type of activated sludge plant. More information is required regarding the

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fates of common odorous pollutants during bubble dissolution, the metabolic pathways involved in the degradation of such compounds, and the long-term effects on concurrent wastewater treatment by the undefined impacts of odour pollutants on mixed liquor chemistry and microbiology.

Acknowledgments The authors would like to thank Anglian Water Services for financial support of this work.

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