Volatile Organic Compounds Removal Methods: A Review

American Journal of Biochemistry and Biotechnology, 2012, 8 (4), 220-229
ISSN: 1553-3468
© 2012 A. Berenjian et al., This open access article is distributed under a Creative Commons Attribution
(CC-BY) 3.0 license
doi:10.3844/ajbbsp.2012.220.229 Published Online 8 (4) 2012 (http://www.thescipub.com/ajbb.toc)
Volatile Organic Compounds Removal Methods: A Review
1
Aydin Berenjian, 1Natalie Chan and 2Hoda Jafarizadeh Malmiri
1
School of Chemical and Biomolecular Engineering,
Faculty of Engineering, University of Sydney, Sydney, Australia
2
Department of Chemical Engineering, Sahand University of Technology, Tabriz, Iran
Received 2012-09-24, Revised 2012-10-04; Accepted 2012-10-17
ABSTRACT
Volatile Organic Compounds (VOCs) are among the most toxic chemicals which are detrimental to humans
and environment. There is a significant need of fully satisfactory method for removal of VOCs. There are
several methods including physical, chemical and biological treatments available to remove VOCs by either
recovery or destruction. The aim of the present study is to summarize the available methods for VOC
removal; trying to find a promising method among the available techniques. A wide range of VOCs can be
treated biologically in which it offers advantages over more traditional processes including lower operating
and capital costs and a smaller carbon footprint. However, due to a complex nature and diversity of VOCs it
is hard to find a simple and promising method. Treatment still requires more research to solve the associate
problems with available VOC elimination techniques.
Keywords: Volatile Organic Compounds (VOCs), Removal Techniques, Bio-Treatment
2012). Most VOCs are photo-chemically sensitive and
when exposed to oxides of nitrogen and sunlight, would
form ozone and other products as represented in Equation 1:
1. INTRODUCTION
Volatile Organic Compounds (VOCs) are man-made
and/or naturally occurring highly reactive hydrocarbons.
World Health Organisation defined VOCs as any organic
compound whose boiling point is in the range from (50100°C) to (240-260°C), corresponding to having
saturation vapour pressures greater than 102 kPa at 25°C
(ISO16000-6, 1989). Many types of VOCs are toxic or
even deadly to humans and can be detrimental to the
environment. Therefore a multitude of definitions exist
globally depending on the context frame used by
different organisations such as United Nations Economic
Commission for Europe (UNECE) and U.S.
Environmental Protection Agency (EPA).
Natural origins of VOCs include wetlands, forests,
oceans and volcanoes with the estimated global VOCs
biogenic emission rate at about 1150Tg/yr (Guenther et al.,
1995). A majority of VOCs are created from anthropogenic
activities consisting of manufacturing industries,
petrochemical industries and vehicular emissions (EPA,
NO x + VOC + Sunlight → O3 + NO x + other products
(1)
The reactions represented by Equation 1, involve VOCs
oxidation with NOx, hydroxyl catalysing some of the key
reactions including other chemical compounds. Ozone
formation is thus primarily driven by available nitrogen
oxides and VOCs. Resulting ground level ozone formation
and carcinogenic smog is the main cause of concern. As the
wide range of VOCs implies a broad range of reaction rates,
VOCs are capable of long range distribution and
accumulation in components of environment. Efficiency of
ozone formation varies with the ratio of nitrogen oxides to
VOCs with higher NOX -FQUOTE NOx emissions resulting
in reduced ozone production efficiency (Finlayson-Pitts and
Pitts, 1986; 1999; Chanin, 1993). The isopleth plot in Fig. 1
illustrates the relation between O3, NOx and VOC with clear
regimes
of
different
O3-NOx-VOC-sensitivity.
Corresponding Author: Aydin Berenjian, School of Chemical and Biomolecular Engineering, Faculty of Engineering,
University of Sydney, Sydney, Australia
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Dependencies vary on assumptions and conditions used
during plot generation. When NOx is limited, O3
increases with increasing NOx with relatively minimal
response to increasing VOC. Thus VOC control could
be less effective in reducing O3 in NOx sensitive areas
e.g., rural areas and downwind from the source
(Finlayson-Pitts and Pitts, 1986).
In the limited VOC regime to the left of the ridge
line, lowering VOC concentrations result in lower peak
ozone concentrations. Contradictory lowering NOx at
constant VOC concentrations predicts increased peak
ozone concentrations to the ridge line. This is due to
more hydroxyl radicals available to react with VOCs due
to reduced NOx-hydroxyl reactions. Only past the ridge
line does the ozone concentration begin to decease. Ozone
concentrations are also lowered by proportionately
decreasing both VOCs and NOx concentrations
simultaneously (Dodge, 1977; Finlayson-Pitts and Pitts,
1986). Thus, Fig. 1 emphasises there is greater benefit in
reducing VOC concentrations over NOx for optimal ozone
control strategies closer to the source of emission.
VOCs are not only outdoor pollutants as high
concentrations have been recorded indoors as well.
Indoor sources include solvents used in the production
and maintenance of building materials, furnishings or
equipment e.g., paint, carpets, plastics and photocopying
machines. The National Health and Medical Research
Council (NHMRC) interim national indoor air quality
recommends a maximum hourly average total VOC level
of 500 µg m−3, where each compound should not
contribute more than 50% of the total (NHMRC, 2002).
In a study investigating VOC emissions from office
furniture, a range of major VOCs emitted over time was
found to be in excess of the NHMRC recommendation.
Formaldehyde was found to be emitted in high
concentrations from furnishings using reconstituted
wood-based panels e.g., particleboards, medium-density
fibreboard (NHMRC, 2002). Table 1 details the results
for major VOCs emitted by new office furniture.
Indoor levels of VOC can be regulated by selecting
low-emission materials for furnishing, cleaning and
construction. This approach is impractical as many of
the materials emitting VOCs are considered standard
fixtures and attempting to replace them could prove
costly. Furthermore, proper ventilation with outdoor air
can help in managing air pollutants indoors by means of
dilution. However with the increasing number of highrise buildings developing in major cities, the problem of
VOC pollution becomes both an indoors and outdoors
issue that requires a permanent solution (Guo et al.,
2004; Alvarez-Hornos et al., 2008).
Various traditional VOCs removal and elimination
technologies exist. In the present study we summarize
the available methods for VOCs removal; try to find a
promissing technique.
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1.1. Common VOCs Removal Methods
There are physical, chemical and biological
treatments available to remove VOCs from air by either
recovery or destruction. The contaminated liquid can be
aerated either via packed tower aeration or mechanical
surface aeration with VOC emissions captured and
treated. Traditional primary removal solutions for VOCcontaminated liquids are liquid-phase activated carbon
adsorption or air stripping. For soil VOC contamination
remediation, Soil Vapour Extraction (SVE) is used (EPA,
1991; Tchobanoglous and Kreith, 2002). These methods
allow for the extraction of VOCs into another phase.
In liquid-phase activated carbon adsorption, the
treated liquid is put in physical contact with activated
carbon to allow dissolved organic contamination to bind
to it. The activated carbon can be either regenerated or
removed after treatment (DOE, 1994). Reactors
commonly used for this process are the fixed bed and the
pulsed/moving bed. When dealing with halogenated
VOCs and pesticides, this carbon method has limited
effectiveness. Economical and logistical issues would
arise from disposing or decontaminating the spent
carbon, hence carbon adsorption is applied more
effectively for “polishing” post-treated liquid discharges
with low VOC concentrations (EPA, 1990; 1993; 1995).
Similarly VOCs in air emission can be treated with
activated carbon by pumping it through activated carbon
packed bed reactors. However, problems with spent
carbon are the same as the liquid-phase carbon method.
Common permanent solutions to treat extracted VOC
emissions would be through oxidation by thermal,
internal combustion engine, catalytic or UV. Essentially
VOCs are broken down into less harmful compounds
such as carbon dioxide, water and hydrochloric gas.
Thermal oxidation units are generally single
chambers with ceramic blanket refractory lining the
oxidisers, equipped with a propane or natural gas burner
and a stack. In the chamber, burner capacities range
between 0.4 to 2 (mil BTUs)/hr with operating
temperatures from 760 to 870°C and a maximum gas
residence time of 1 sec. The internal combustion engine
works in the same way; however, it is adapted to high
VOC concentrations to allow the organic compounds to
be used as fuel. Auxiliary fuel is only added to enhance
oxidation. To reduce the need for auxiliary fuel, air to
air heat exchangers can be used to transfer heat from the
exhaust gases to the incoming feed.
In treating
halogenated VOCs, the exhaust stream would require a
gas scrubber to control the acid vapour (Benitez, 1993).
Oxidation is deemed inefficient for low VOCs
concentrations in the range of between 0.1 to 10g Mm−3
as continuous oxidation is difficult to maintain unless
with
more
supporting
fuel
(Rao,
2007).
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Fig. 1. General ozone isopleth used in EPA’s, Empirical Kinetic Modelling Approach (EKMA) modelled against ratios found in
cities (Dodge, 1977)
Fig. 2. Schematic diagram of a biofilteration unit (Delhomenie and Heitz, 2005)
Catalytic oxidation involves the addition of a
catalyst to a thermal oxidiser to accelerate oxidation
via the adsorption and reaction of oxygen and organic
compounds on the catalyst surface. This lowers the
reaction temperature to a range between 320 to 540°C
compared to that required in a conventional thermal
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oxidiser. The direct pre-heating of feed to the
reaction temperature is required to initiate catalytic
oxidation when passed through a solid catalyst bed.
Catalysts may contain various metal oxides such as
chromium oxide or nickel oxide and may even contain
noble
metals
e.g.,
palladium.
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Table 1. Formaldehyde and VOC concentrations and emission factors from new office furniture (Brown, 1999)
Concentration (µg/m3)
Emission Factor (µg/m2.h)
-------------------------------------------------------------------------------------------VOC
4h
24 h
4h
24 h
Formaldehyde
190
230
240
290
Freon 22
42
700
530
890
Methanol
4000
2700
5000
3400
Acetone
450
254
560
320
Butanal
290
220
360
280
Ethylacetate
660
580
830
730
Methyl isobutyl ketone
200
210
250
260
Hexanal
28
23
35
29
m- and p-Xylene
59
48
74
60
Cyclohexanone
55
54
69
68
Phenol
180
90
230
110
Total VOC
970
820
200
1030
organisms matched to the type of VOC to be destroyed.
Some easily biodegradable VOCs are esters, benzene,
toluene and phenols. Due to future growth potentials for
the use of biofiltration methods, these methods are
detailed and expanded upon in the following sections.
Thermal oxidisers can be used as pre-treatment to
catalytic units for high VOC concentrations (EPA, 1987;
Benitez, 1993; Rao, 2007; Berenjian and Khodiev, 2009;
Berenjian et al., 2011).
UV oxidation method has been used to oxidise
organic and explosive compounds in wastewater. Strong
chemical oxidisers directly react to the contaminants
with UV photolysis achieved through synergistic UV
irradiation combined with ozone and/or hydrogen
peroxide (Christman and Collins, 1990; EPA, 1990).
Generally low pressure 65W lamps are used for ozone
treatment systems and 15-60W lamps are used in
hydrogen peroxide treatment. Final products of UV
oxidation are carbon dioxide, water and salts. Although,
able to treat a wide range of VOCs, certain contaminants
may be volatilised rather than destroyed e.g., TCA. In
such a case, post treatment by activated carbon
adsorption may be required (Zappi et al., 1992).
Bio-treatment of VOCs is a relatively less established
VOC removal method; however, it offers possible
advantages over more traditional processes with lower
operating and capital costs and a smaller carbon
footprint. This is due to lower energy requirements as
microorganisms are used to metabolise organic
compounds at ambient temperature instead being
dependent on heat or radiation. Since 1923, gas biofiltration has been applied to treat various VOC exhaust
contaminants (Leson and Smith, 1997; Wieczorek,
2005). Biological methods include bio-filters, biotrickling filters, bio-scrubber, suspended growth reactors
and membrane bio-reactors (Doble and Kumar, 2005).
Key parameters across bio-systems are identified as
moisture content in the medium, temperature, pH and
availability of essential and non-carbon nutrients (Leson
and Winer, 1991; Swanson and Loehr, 1997; Sidkar and
Irvine, 1998). As a wide range of VOCs are
biodegradable, they can be treated biologically with
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1.2. Bio-Treatment
Bio-technologies effectively and economically treat
contaminated air flows from 60 to 150,000m3/h with
lower VOC concentrations than most other available
technologies. This advantage enables it to be applied to
smaller scale projects such as possibly retro-fitting
building ventilations with a compact VOC treatment
system. Existing bio-treatment of VOCs used in gas
treatment revolves around the concept of bio-filtration.
There are various other bio-filtration processes available
such as Bio-Trickling Filters (BTFs), bio-scrubbers and
newer technology e.g., bio-membrane bioreactors (Cox
and Deshusses, 1998; Doble and Kumar, 2005). Biofiltration is generally based on two principals: firstly, the
transfer of pollutants from the air feed to the support
medium. Secondly, the contaminants are bio-catalysed to
biomass, carbon dioxide, water and other by-products
(Miller and Allen, 2004). Schematic diagram of a
biofilteration unit is shown in Fig. 2.
The traditional bio-filteration system is essentially a
fixed-bed bioreactor where the bio-catalyst or
microorganism is immobilised to an inert supporting
medium to develop biofilms. Ideal packing bed materials
have high void fraction, light weight, low pressure drop,
hydrophilic and low bulk density. Occasional irrigation of
nutrient solution onto the packing bed and humidity
control maintains the biofilms. Polluted air is fed from the
bottom through the porous biologically-active media,
where pollutants diffuse into the aqueous-phase or are
absorbed directly by the biofilm with treated air emitted
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contaminants are pressurised to induce diffusion over the
partition into the aqueous solution. The membranes can be
further treated to have various advantageous properties e.g.,
hydrophobic and microporous. Membrane bioreactors for
VOC treatment still require more research to minimise
fouling and high cost (Artiga et al., 2005).
from the top. Unfortunately, VOCs containing sulphates,
chlorides and nitrates have acidic by-products that can
adversely affect the ecosystem and reduce the
effectiveness of the bioreactor (Garner and Barton, 2002).
The biotrickling filter has gas flowing through a fixed
bed with microorganisms immobilised through it. Ideal
packing material is similar to the traditional set up.
Nutrient solution is continuously irrigated from the top,
collected at the bottom and recycled back up. Co-current
or counter-current flow between the gaseous and
aqueous phases was found to have no influence on the
degradation performance (Cox and Deshusses, 1998).
The biotrickling system allows for more efficient
removal of soluble VOCs from the air stream as well as
enabling feedback control for nutrient levels and pH. As
there is a constant feed of nutrients to the system, excess
biomass builds up quickly and can lead to performance
loss and pressure drops from uneven biomass
distribution (Cohen, 2001).
Bioscrubbers consist of an absorption tower and a
bioreactor.
In the absorption tower, gas phase
contaminants are diffused into an aqueous solution via
means of counter-current gas-liquid flows through inert
packing. The washed gas is emitted from the top and the
contaminated liquid pumped towards an aerated
bioreactor. The microorganism or activated sludge in the
reactor is suspended in a nutrient-rich media and
residence time for treatment varies according to the type
and concentration of VOCs in the feed. After completed
degradation of contaminants, the medium is filtered and
biomass left to sediment with portions being recycled
through the bioscrubbing process again. An aqueous
system with no high pressure drops allows for more
evenly distributed temperature, nutrient and pH controls.
System limitations are due to the narrow band of VOCs
treatable. This range is limited by water solubility of
contaminants thus only applying to organics with low
Henry’s coefficient (<0.01) at low concentrations
(<5g/m3). Compared to other bioreactors considered, the
bioscrubber requiring two sub-units could be a space
limitation; especially when multiple systems in series or
parallel are required to process higher VOC volumes.
Bioscrubbers, however, are not as well-utilised within
the biotechnology market compared to biofilters and
biotricklers (Swanson and Loehr, 1997). To treat lesssoluble pollutants, silicon oil could be added to aqueous
systems as an emulsifying agent to favour the VOC mass
transfer from gas to liquid phases.
Finally, membrane bioreactors involve a membrane
partitioning the liquid and gas phases. The nutrient and
biomass growth is within the liquid side whilst the gaseous
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1.3. Operating Parameters
The operations of biofilters involve a series of steps
beginning with the transfer of the pollutant air to the
aqueous phase, followed by its adsorption onto the medium
or absorption into the biofilm and then biodegradation of
VOCs within the biofilm (Kumar et al., 2011). The
biodegradation of pollutants by biofilms in biofiltration
sytem is a combination of physicochemical and
biological phenomena. Basically these three mechanisms
are responsible for the transfer and subsequent
biodegradation within the bed (Doble and Kumar, 2005;
Mathur et al., 2006).
Governing control parameters across all bioreactors includes packing material, air flow rate,
temperature (Darlington et al., 2001), pH, humidity
and non-carbon nutrients e.g., nitrogen, phosphate and
potassium (VanLith et al., 1997; Devinny et al.,
1999). However, as the type of microorganism used is
dependent on the type of VOC to be destroyed,
treating a wide range of VOCs in indoor air would result
in a diverse ecology of mixed fungal and bacterial
populations (Khammar et al., 2005). This diversity
creates complex interactions within the bioreactors
which require research into the ecology of the biomass to
optimise overall management of the ecosystem.
1.4. Packing/Filter Bed
As the filter bed physically supports the entire
biosystem, some of the criteria required to achieve
optimal conditions are high-specific surface area which
is excellent for microbial activity, decent water retention
capacity and high porosity. Basic traditional organic
materials used for packing are peats, composts and soils
because of their stability, low cost and effectiveness
(Doble and Kumar, 2005). Modern approaches enable
beds to have evenly-distributed rigid structures
preventing bed compaction. Furthermore, biocatalysts
can be immobilised within a porous structure (polymer
beads) to allow for easier maintenance and preservation
of specific microorganisms (Lu et al., 2002). Pressure
drop of the filter bed is an important biofiltration
parameter which drastically effects on operating cost.
Filter beds including small particles have high specific
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surface area which in turn make them suitable for
microbial activity and increase their resistance to air
flow (Kumar et al., 2011).
medium can have direct effect on microorganisms and
microbial enzymes and also indirect influence on the
availability of required nutrients. In fact, depending on
the type of microorganism the pH has significant effect
on the biofiltration efficiency (Clark et al., 2004;
Delhomenie and Heitz, 2005). Fungi are able to grow
under both neutral as well as acidic conditions and are
metabolically active over a wide pH range between
approximately 2 and 7. Whereas bacteria are usually
considered to be less tolerant to pH fluctuations and
require a near neutral environment for their activity
(Kennes and Veiga, 2004). Due to neutrophilic behaviour
of the most of microorganisms in biofilters, maximum
degradation of VOCs would usually be achieved at pH
around 7 (Kumar et al., 2001). Veiga et al. (1999) studied
the effect of pH on alkyl benzene degradation (between
pH 3.5 and 7.0) and found that alkyl benzene degradation
increases with pH increase. Lu et al. (2002) reported the
maximum degradation for BTEX between pH values of
7.5 and 8.0. However, the ideal pH of the biofilter medium
depend on the pollutant being treated and the
characteristics of the microbial ecosystem.
1.5. Flow Rate
In general, overall efficiency of biofilters is related to
two main microscopic processes. The first process is
transfer rate of the VOCs from gas phase to biofilm and
the second one is biodegradation rate of VOCs
(Delhomenie and Heitz, 2005). The air flow rate is an
important parameter which can affect both of these
processes. In fact, by increasing the flow rate, the rates
of VOCs transfer to the biofilm and degradations
decrease. This can be explained by the fact that in
extremely high air flow rates, contact times between
microorganisms and gases are too short and the
biodegradation reaction cannot be completed (Cox and
Deshusses, 1997). The Empty Bed Residence Time
(EBRT) is defined as a time that parcel of air will
remain in the biofilter and calculated as the empty
bed filter volume divide by the air flow rate. Several
studies have shown that to improve biofiltration
performance and better VOCs removal efficiency, the
EBRT should be greater than the time required for
diffusion process (Doble and Kumar, 2005; Kumar et al.,
2011). It has been reported that by increasing the size of
particles in the filter beds, the resistance of filter beds
decrease towards gas flow (Clark et al., 2004).
1.8. Moisture
The moisture content of the filter bed plays an
important role in biofilter performance because
microorganisms need water to attain their metabolic
activity (Kumar et al., 2011). Availability of excess
water hinders the transfer of oxygen and hydrophobic
pollutants to the biofilm, leading to the development of
anaerobic zones within the bed. It also leads to the
reduction of the specific surface available for gas/biofilm
exchanges, which in turn, causes bed compaction and
increasing the pressure drop. Too low bed moisture
content leads to bed desiccation and gas flow channeling,
which particularly affects the microflora (Delhomenie
and Heitz, 2005). The support of microbial populations
sufficient to reduce VOCs requires that moisture levels
in the filter medium be maintained between 40% and
70% (w.b.) (Clark et al., 2004). The main factors
affecting the bed moisture content are gas flow rate
through the bed, water holding capacity of the filtering
materials, reaction exothermicity and inlet gas relative
humidity (Devinny et al., 1999).
1.6. Temperature
Operating temperature is one important factor that
affects biofilter performance. The microorganisms that
most effectively degrade VOC compounds in biofilters
are mesothermic and their optimum activity temperatures
are between 30 and 40°C (Clark et al., 2004). In fact, the
proper temperature increases the rate of biofilm
development and biomass accumulation (Yang et al.,
2010). Deeb and Alvarez-Cohen (1999) indicated that in
the optimum range of operating temperature, the
degradation performance on biofilter can increase 2fold by 10°C increase in temperature. This can be
explained by the fact that by rising the temperature,
VOCs and O2 solubilise in water decrease and
diffusion transfer increases. Due to exothermic nature
of the biodegradation reactions, the variation of
temperature in the filter bed is a consequence of
microbial activity (Hwang et al., 2002).
1.9. Microorganism
Microorganisms including bacteria and fungi are used
in VOCs biodegradation methods. Researchers have
indicated that heterotrophic bacteria and fungi are the
main microorganisms used in elimination of VOCs
(Delhomenie and Heitz, 2005). However, the main
1.7. pH
For most bioreactors, pH has a significant impact on
biofiltration efficiency. The pH of the biofilteration
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found to be significantly removed by biofilm
(VanGroenestijn and Liu, 2002). Upon comparison,
the α-pinene partitioning of biofilm containing
biomass and organics was also found to be greater
than biofilm in water (Miller and Allen, 2004).
As the dominant carbon source, the VOC gas supply
for the biotrickling filter creates a gradient of high
specific biomass growth rates starting from the feed inlet
and tapering off towards the outlet (Song and Kinney,
1999). In a gas-phase biofilter, biomass accumulation
rate is equal to the biofilm growth rate minus the biofilm
decay rate resulting from hydraulic scouring or
detachment. Thus, allowing for general modelling of
biofilm thickness over time to be calculated from
bacterial growth and decay as shown in Equation 3 and 4
(Alonso et al., 1997; 1998):
advantage of fomenting fungal rather than bacterial
growth for the biofiltration of hydrophobic pollutants is
their ability to degrade the substrates under extreme
environmental conditions regarding pH, low water
content and limited nutrient concentrations (Kennes and
Veiga, 2004). Generally, heterotrophic microbial strains
used in VOC bioreactors, can bio-catalyse VOCs via two
pathways: (1) consuming organic compounds in the
course of catabolic pathway for energy or (2) using
VOCs as a carbon source for anabolic processes. Species
of Pseudomonas, Candida, Mycobacterium, Alcaligenes,
Exophiala, Acetinobacter, Fusarium, Cladosporium,
Rhodococus, Aspergillus and mucor are some of
microorganisms which identified and used for
degradation of VOCs by biofiltration (Marek et al.,
1999; Diehl et al., 2000; Christen et al., 2002; Qi et al.,
2002; Woertz et al., 2002). The filter bed inoculation
depends on both the nature of the filtering materials and
the VOC biodegradability level. Generally, a biofilter
contains between 106 and 1010 cfu of bacteria and
actinomycetes per gram of bed, respectively. It can
also contains 103 to 106 spores of fungi per gram of
bed (Delhomenie and Heitz, 2005).
Biofilm is a complex aggregation of microorganisms
(aerobic, anaerobic and facultative type bacteria, fungi,
algae and protozoa) which attach themselves on the
surface of the packing media and forms a biological
film or slim layer with a viscous slimy structure
(Kumar et al., 2011). There are three main biological
processes that occur in the biofiltration systems,
namely: attachment of microorganisms, growth of
microorganisms and decay and detachment of
microorganisms (Yang et al., 2010).
Assuming the biofilm is built on water, Henry’s Law
represented in Equation 2 roughly estimates maximum
concentrations of aqueous pollutant ( C*aq ) available to be
(2)
where, pi represents the partial pressure of a specific
pollutant in gaseous phase and Hi is Henry’s constant
coefficient of the pollutant. This model of estimation ,
however, does not in fact express the transport and
reaction processes of the biomass itself. Waterinsoluble compounds such as α-pinene have been
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b = bs + b d
(4)
Other factors such as oxygen availability, carbon
source and biofilm activity all interact to affect the
biofilm decay rate. Factors contributing to shear force
could result from calculating in fluid flow or particle
attrition as well. This general model of prediction would
enable biomass controls to be set in place to prevent a
drop in efficiency from excess biomass. There are five
methods of control including physical, chemical,
biological, improved biofilter designs and improved
operation modes. Most physical treatments involve bed
stirring or backwashing to drain excess accumulated
biomass. Chemical controls can be put on the carbon and
nutrient sources to starve the biofilm. Periodical
starvation does not affect removal performance in
biotrickle reactors as microorganisms degrade their own
extracellular polymers for energy (Zhang and Bishop,
2003). Chemical washing of the biofilter to control
bio-catalysed:
pi
Hi
(3)
Where:
Lf = Biofilm thickness
rd = Ratio between VOC diffusivities in biofilm and
water
Dw = Contaminant diffusivity in water
Cf = VOC concentration in biofilm
Y = Yield coefficient
Xf = Film bacterial density
b = Specific shear/decay coefficient
bd = Specific decay coefficient
bs = Specific shear rate
1.10. Biomass
C*aq =
∂Lf
∂C
Y
= (rd D w f | x = L )
− Lf b
∂t
∂x
Xf
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biomass could have a detrimental effect on performance
and requires time to reacclimatise. It must be noted that
to achieve higher bio-conversion efficiency, the bio-filter
should be exposed to low VOC concentrations initially to
allow culture adaption to utilising VOCs as a carbon
source. Following which, the organic loading can be
increased over time.
2005). Some disadvantages of using bioscrubbers are
treatment of only water soluble compounds, generation
of liquid waste, need of extra air supply, need of excess
sludge disposal and operation/maintenance complicities
(Kumar et al., 2011).
2. CONCLUSION
1.11. Overview of Using Biological Treatments
Based on the present study, biological methods
(biofilter, biotrickling filter, membrane bioreactor and
bioscrubber) can be used as potential methods for VOCs
removal as compared to the other available elimination
techniques. However, use of biological systems to remove
VOCs still have limitations and challenges. These systems
do not response well to sudden loading stresses and can
fail. Due to detrimental nature of VOCs to humans and
environment, more practical and more efficient VOC
removal techniques are in demand. A combination of
existing treatments can be considered as another approach
which may help to increase VOCs treatment effectiveness
at higher concentrations, whilst reducing costs e.g.,
biofiltration as a post-treatment to adsorption.
As previously detailed, the four most common types
of biological treatment units are biofilter, biotrickling
filter, membrane bioreactor and bioscrubber. These
systems have differences in their complexity, process
design, equipment dimensions and working parameters,
but they have unique advantages which make them
proper techniques to remove wide range of compounds.
Biofilters have advantages such as low initial investment
and subsequently minimized operating cost, degradation
a wide range of components, easy to operate and
maintain, no production of unnecessary waste streams
and low pressure drop (Yang et al., 2010). Biotrickling
filters have less operating and capital constraints, less
relation time/high volume through put and capability to
treat acid degradation product of VOCs (Cox and
Deshusses, 1998). In membrane bioreactors there are no
moving parts, process scale up is more easy and flow of
gas and liquid can be varied independently, without the
problems of flooding, loading or foaming (Artiga et al.,
2005). Bioscrubbers need relatively smaller space
requirements and able to deal with high flow rates and
severe fluctuations. Other advantages of using this
system can be mentioned as operational stability, better
control of operating parameters and relatively low
pressure drop (Swanson and Loehr, 1997).
Applications of biological systems to remove VOCs
have some limitations. For example, limitations of using
biofilters are less treatment efficiency at high
concentrations of pollutants, extremely large size of
bioreactor challenges space constraints to require close
control of operating conditions and limited life of
packing and clogging of the medium due to particulate
medium (Delhomenie and Heitz, 2005). Accumulation of
excess biomass in the filter bed, requirement of design
for fluctuating concentration, complexity in construct
and operation and secondary waste stream are among the
disadvantages of using biotrickling filters (Cox and
Deshusses, 1997; Doble and Kumar, 2005). Limitations
of application of membrane bioreactors to removal
VOCs are high construction costs, long-term operational
stability and possible clogging of the liquid channels due
the formation of excess biomass (Doble and Kumar,
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