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Microbubble-aided water and wastewater purification: A review
Article in Reviews in Chemical Engineering · December 2012
DOI: 10.1515/revce-2012-0007
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DOI 10.1515/revce-2012-0007 Rev Chem Eng 2012; 28(4-6): 191–221
Snigdha Khuntia , Subrata Kumar Majumder * and Pallab Ghosh
Microbubble-aided water and wastewater purification: a review Abstract: Microbubble-based methods, in recent times,
have been widely used for purification of water and
wastewater. Microbubbles have several physicochemi-
cal properties, which make them eminently suitable for
wastewater treatment. In this review, these properties
have been analyzed in detail from the perspective of appli-
cation. Various types of microbubble generators and their
operation principles have been discussed. The transport
of gas into the aqueous phase has been explained, and the
correlations to predict the volumetric mass transfer coef-
ficient have been presented. Many practical applications
using ozone, oxygen and air microbubbles, some of which
are currently at various stages of commercialization, have
been presented. Other important uses of microbubbles
for wastewater treatment, namely, removal of fine solid
particulate matter and oil, have also been discussed. In
addition, directions for future research of microbubble
technology and their potential applications have been
identified.
Keywords: microbubble; microbubble generator; ozona-
tion; wastewater treatment.
*Corresponding author: Subrata Kumar Majumder , Department of
Chemical Engineering , Indian Institute of Technology Guwahati,
Guwahati-781039, Assam , India ,
e-mail: [email protected]; [email protected]
Snigdha Khuntia : Department of Chemical Engineering , Indian
Institute of Technology Guwahati, Guwahati-781039, Assam , India
Pallab Ghosh : Department of Chemical Engineering , Indian Institute
of Technology Guwahati, Guwahati-781039, Assam , India
1 Introduction The harmful effects of toxic chemicals in civilization are
well-known. It is not possible to eliminate all such chem-
icals due to several reasons, such as yield of secondary
compounds and an economic rate of production. Pres-
ervation of a safe and clean environment has become a
social problem, and water is particularly important from
this perspective (Hiroshi 2006 ). Amid deepening global
environmental problems, various efforts have been
taken for improving water quality (Akiko et al. 2005 ). In
recent years, microbubbles have received some impor-
tance among water purification technologies (Burns et al.
1997 , Nakano et al. 2005 , Matsuo et al. 2006 , Usui 2006 ,
Yamasaki et al. 2009 , 2010, Agarwal et al. 2010 , Wen et al.
2011 ). Microbubbles of air, oxygen and ozone are being
extensively used in various water treatment applications.
Apart from water treatment, microbubbles have
found use in various other fields. In medical therapeutic
applications, microbubbles have been used for scanning
body organs, and also as a drug or gene carrier (Lindner
2004 , Tsutsui et al. 2004 , Matsumoto et al. 2005 , Hernot
and Klibanov 2008 , Maliwal and Patidar 2008 , Kurup
and Naik 2010 , Dicker et al. 2011 ). Microbubbles have
been used for antibacterial activities under aerobic and
anaerobic conditions (Himuro et al. 2009 ). Microbubbles,
stabilized by various surfactants, are used for adsorp-
tion of protein from aqueous solution (Jauregi and Varley
1998 ). Petroleum-based as well as biological surfactants
have been used for stabilizing microbubbles (Kukizaki
and Baba 2008 , Xu et al. 2011 ). Microbubbles have been
functionalized by surfactants, nanoparticles, pharmaceu-
ticals and bioactive molecules. Owing to their small size,
microbubbles are highly effective in industrial separation
processes, such as removal of volatile contaminants and
particulate matters present in the aqueous phase (Ahmed
and Jameson 1985 ). Flotation processes employing micro-
bubbles are useful for the removal of low-density particu-
late matter present in water (Terasaka and Shinpo 2007 ).
Microbubbles act as carriers of fine particles, which are
lifted up from the bottom of the column. The flotation
process can be dissolved-air flotation, dispersed-air flo-
tation or electroflotation, depending on the microbubble
generation technique (Ketkar et al. 1991 , Liu et al. 2010a ).
Microbubbles have also found use in cleaning (Himuro
2007 , Akuzawa et al. 2010 ), soil washing (Roy et al. 1992 ),
removal of oil from soil (Gotoh et al. 2006 ) and water
(Xiaohui et al. 2011 ), fermentation (Kaster et al. 1990 , Ago
et al. 2005 , Xu et al. 2011 ), marine fish farming (Tsutsumi
2010 ), horticulture (Park and Kurata 2009 ), food tech-
nology (Shen et al. 2008 , Xu et al. 2008 , Soli et al. 2010 ),
absorption of acid gas (e.g., CO 2 ) by alkali (Akimov et al.
2011 ), and many more applications, which have been
summarized by Li (2006) . Microbubbles can enhance the
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192 S. Khuntia et al.: Microbubble-aided water purification
growth rate of marine creatures, such as oysters and scal-
lops, by providing a large gas-liquid interfacial area that
facilitates interphase transfer of air.
Water treatment using microbubbles has recently
become a well-known technology for many industrial
applications due to its superior efficiency, compared
with conventional methods (Ohnari 1997 , Jyoti and Pandit
2001 ). Ozone microbubbles have been used for oxida-
tion, disinfection (Sumikura et al. 2007 ), decolorization
and deodorization (Shin et al. 1999 ). Ozone has a strong
oxidizing ability, and by utilizing this capability its use
in water purification or sewage water treatment is antici-
pated for sterilization, removal of color and odor, and
degradation of organic substances (Camel and Bermond
1998 ). In this review, various properties of microbubbles,
available technologies for generation of microbubbles and
the measurement of their size are discussed. In addition,
the major applications of microbubbles in the treatment of
water and wastewater are discussed from the perspective
of future applications and research directions.
2 Structure and properties of microbubbles
Microbubbles are tiny spherical bubbles with a diameter
of ≤ 50 μ m. They are different from ubiquitous common
bubbles (also known as ‘ macrobubbles ’ ), not only in terms
of size but also in terms of their physicochemical proper-
ties. These special physicochemical properties have made
microbubbles particularly useful in various water and
wastewater treatment applications (Ohnari et al. 2002 ,
Takahashi 2004 , 2010, Tsuge 2010 ). These features are:
low rising velocity through water, surface having high
curvature, large gas-liquid interfacial area and electrically
charged gas-liquid interface. Many of the useful features
of microbubbles for wastewater treatment are associated
with these properties.
2.1 Size and shape
Microbubbles are spherical in shape. For a three-dimen-
sional object with a given volume, the area is minimal
when the object has a spherical shape, which is mathemat-
ically expressed by ‘ isoperimetric inequality ’ (Osserman
1978 ). Therefore, gas-liquid interfacial energy is minimal
when the bubble is spherical. ‘ Macrobubbles ’ , which are
commonly used in fermentors, gas-liquid reactors and
ore flotation equipment, have a diameter in the range of
2 – 5 mm. These bubbles are also called ‘ millibubbles ’ . The
‘ microbubbles ’ have a diameter of ≤ 50 μ m (Takahashi
et al. 2007a ). Bubbles, which have a diameter of < 200 nm,
are known as ‘ nanobubbles ’ (Agarwal et al. 2010 ). The
bubbles, whose diameters lie between 200 nm and 10 μ m,
are called ‘ micro-nanobubbles ’ (MNBs) (Tsuge 2010 ). Cur-
rently, there is no universally accepted classification of
the various types of bubbles in terms of their size.
A photomicrograph of microbubbles is shown in
Figure 1 A (Sumikura et al. 2007 ). The photograph was
taken 20 s after the generation of microbubbles. In water,
microbubbles form a milky dispersion, as shown in Figure
1B. The dispersion containing nanobubbles is, however,
transparent (Tsuge 2010 ). According to Sebba (1988), a
surfactant-stabilized microbubble is composed of two
concentric gas spheres, inside which a layer of water is
sandwiched. This structure is shown in Figure 1C. The
diameter of the inner gas core is approximately 50 – 60
μ m, and the thickness of the internal water film is ∼ 1 μ m
(Bredwell and Worden 1998 ). The surfactant molecules
impart a stabilizing influence to the bubble against
Bulk liquid
100 μm
A
B C
Aqueous film
Inner gas core
Gas shell
Figure 1 The size and shape of microbubbles: (A) photomicrograph of microbubbles generated in tap water and collected on a cover-glass
(Adapted from Sumikura et al. 2007 , with permission from the copyright holder, IWA Publishing), (B) microbubble dispersion in distilled
water, and (C) the concentric-gas-sphere model of surfactant-stabilized microbubble (the drawing is not to scale; the regions are exagger-
ated for illustration) (Sebba 1988 ).
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S. Khuntia et al.: Microbubble-aided water purification 193
coalescence by various repulsive interfacial forces (e.g.,
electrostatic double layer, steric and hydration forces).
The stability of the surfactant-stabilized microbubbles
allows them to be pumped without collapse. Therefore,
such microbubbles can be produced in a small vessel and
then pumped to a much larger reactor.
Microbubbles, formed ultrasonically or by special
generators, have a wide size distribution. It is a challeng-
ing task to produce monodispersed microbubbles (Zhang
and Li 2010 ), which are important in medical applications
(Feshitan et al. 2009 ). Size distribution can be unimodal
or bimodal. A typical size distribution (Tsuge et al. 2009 ) is
shown in Figure 2 . The microbubble generation technique
has a significant effect on the size of the microbubbles (Li
2006 ). Tsuge et al. (2009) have studied single pore nozzle
and rotating flow microbubble generators and compared
the size of the bubbles formed by them. With the rotat-
ing flow microbubble generator, the bubbles are sheared
and crushed by the rotating flow, which results in smaller
microbubbles. In addition, at large outlet pressure of the
pump bubbles of smaller size are produced (with both
generators). This is due to the change in pressure at the
time of bubble formation, which is greater and results in
smaller bubbles. The presence of surface active impurities
and electrolytes can have a significant effect on microbub-
ble size. Walker et al. (2001) have reported that the size of
microbubbles decreased with increasing NaCl concentra-
tion up to 1 mol/dm 3 . It has been reported (Hofmeier et al.
1995 ) that surface elasticity plays an important role in the
formation of smaller bubbles in the presence of NaCl. Pure
liquids, which have small surface elasticity, give rise to
large bubbles through coalescence. In aqueous solutions,
12
9
Single pore nozzleRotating flow microbubble generator
6
Freq
uenc
y (n
umbe
r %)
3
00 50 100
Bubble diameter (mm)
150 200
Figure 2 Size distributions of microbubbles generated by a single
pore nozzle and a rotating flow microbubble generator at 0.55 MPa
pressure (Tsuge et al. 2009 ).
where surface elasticity is larger, bubble size is reduced
through inhibition of coalescence.
2.1.1 Measurement of size of microbubbles
Several methods are available for measuring the size
of microbubbles. Bubble diameter down to ∼ 1 μ m can
be measured by a photographic technique (Zhang et al.
2000 , Kawahara et al. 2009 ). Laser light scattering is a
more advanced method for measuring the size of micro-
bubbles. In this method, a photomultiplier tube detects
the scattered light. The relationship between the intensity
of the scattered light and the bubble diameter varies with
the size of the bubbles. For nanobubbles, the Rayleigh
scattering theory may be applicable (if the diameter of
the bubbles is much smaller than the wavelength of the
light), and the scattered intensity is proportional to the
sixth power of the bubble diameter (Pelssers et al. 1990 ).
For larger bubbles, for which the diameter is close to the
wavelength of light, the Mie scattering theory is used to
measure the bubble diameter (Kerker 1969 ). The scattered
light intensity varies in a complex manner with the dia-
meter of the bubble. For large microbubbles, the scattered
intensity varies with the square of the bubble diameter
(Glantschnig and Chen 1981 ). The laser scattering tech-
nique is used to measure bubble size distribution in the
range of 0.02 – 2000 μ m. The size and scattered intensity is
calibrated using standard particles (e.g., polystyrene latex
particles). When spherical bubbles are irradiated by laser,
a scatter pattern is obtained. However, bands of interfer-
ence can also be seen on a plane away from the focus. The
bubble diameter is proportional to the number of interfer-
ence bands (Tsuge 2010 ). Bubble diameters ranging from
10 μ m to 1 mm can be measured by this technique.
Another method for the measurement of size of large
microbubbles is the ‘ pore electrical resistance method ’
(Tsuge 2010 ). When electrodes are placed in an electro-
lyte solution, which is partitioned with a separator having
small pores, and a voltage applied to the electrodes, elec-
trical resistance is determined by the resistance of the part
with the micropores. When microbubbles pass through
these pores, electrical resistance increases, and the
number of bubbles and their size can be determined from
the resistance. Bubble diameter in the range of 0.2 – 600
μ m can be measured by this method. However, if the dis-
tribution of bubble diameter has a range wider than this,
the pore size needs to be adjusted.
Guidi et al. (2010) have presented a technique com-
prising acoustic and optical methods to investigate
the response of isolated lipid-shelled microbubbles to
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194 S. Khuntia et al.: Microbubble-aided water purification
low-pressure ultrasound tone bursts. These bursts induced
slow deflation of the microbubbles. Their experimental
set-up had a microscope connected to a camera acquiring
one frame per pulse transmitted by a single-element trans-
ducer. The behavior of each bubble was measured at mul-
tiple frequencies by cyclically changing the transmission
frequency over the range of 2 – 4 MHz. The bubble echoes
were captured by a second transducer and coherently
recorded. Microbubbles with radii larger than 3 μ m did
not experience any size reduction. Smaller bubbles slowly
deflated, generally until the radius reached a value around
1.4 μ m, independent of the initial microbubble size. The
resonant radius was evaluated from the echo amplitude
normalized with respect to the geometrical cross-section.
Hosokawa et al. (2009) have developed a technique
to measure the diameter of microbubbles by using phase
Doppler anemometry and an image processing method
(Burger and Burge 2008), which is based on the Sobel
filter (Schau 1980) and Hough transform (Yu et al. 2007 ).
The size distribution and the mean diameter of the micro-
bubbles were evaluated based on the diameters measured
by both methods. They have claimed that size distribution
can be accurately determined for a wide range of diam-
eters by these methods.
Dynamic light scattering (DLS) (Berne and Pecora
2000 ) is a sophisticated technique for the measurement
of diameter of MNBs and nanobubbles. This method
measures Brownian motion and relates this to the size of
the bubbles. The Brownian motion becomes slower with
the increase in the size of the bubble. The velocity of the
Brownian motion is defined by the translational diffusion
coefficient. The size of a bubble is calculated from the
translational diffusion coefficient by using the Stokes-
Einstein equation (Ghosh 2009a ).
3 A b
RTd
N Dπμ=
(1)
The diameter measured in DLS is a value that refers to
how a particle diffuses within a liquid. Thus, it is referred
to as the ‘ hydrodynamic diameter ’ . The diameter that is
obtained by this technique is the diameter of a sphere
that has the same translational diffusion coefficient as
the bubble. The translational diffusion coefficient will
depend not only on the size of the bubble but also on any
surface structure, as well as the concentration and type of
ions in the medium. Factors that affect the diffusion speed
of particles are ionic strength of the medium (which influ-
ences the thickness of the electrostatic double layer) and
the surface structure (e.g., adsorbed long chain molecules
projected out into the liquid). Variation in temperature can
cause variation in viscosity and, hence, induce convection
currents, which would alter the random movement of the
molecules. Therefore, temperature must be very stable in
the DLS experiments.
An X-ray particle tracking velocimetry (PTV) tech-
nique has been developed by Lee and Kim (2005) to
simultaneously measure the size and velocity of micro-
bubbles in a liquid without optical aberration. This
technique is based on a combination of in-line X-ray
holography and PTV. The X-ray PTV technique uses a con-
figuration similar to that of conventional optical imaging
techniques. These researchers generated microbubbles
with diameters in the range of 10 – 60 μ m from a fine wire
by electrical heating and used them as tracer particles.
The X-ray PTV technique simultaneously recorded the
size and velocity of the microbubbles moving upward in
an opaque tube of 2.7 mm inner diameter. Owing to the
different refractive indices of water and air, phase con-
trast X-ray images showed the exact size and shape of
the microbubbles.
As microbubbles generally have a broad size distribu-
tion, the average diameter of the microbubbles is often
expressed in terms of the Sauter mean diameter (SMD), d 32
(Kawahara et al. 2009 , Liu et al. 2010b , Maeda et al. 2010 ),
which is defined as:
3
32 2
i i
i i
n dd
n d=∑∑
(2)
The bubble size distribution is often plotted as a
cumulative distribution in which a percentage of the total
volume of bubbles below a given size is plotted vs. bubble
size. 90 d , 50 d and 10 d correspond to 90, 50 and 10 volume
percents on this cumulative size distribution curve. The
bubble size dispersion coefficient ( δ ) is defined as (Kuki-
zaki and Goto 2006 ):
90 10
50
-d d
dδ=
(3)
DLS provides the z -average diameter of the bubbles.
This is the mean diameter of the distribution based on the
intensity of scattered light, and this diameter is obtained
by the cumulants analysis of the correlation function. It
is not a mass or number average because it is calculated
from signal intensity. In DLS, this is the most important
and stable value that is produced. This is the size to be
used if a value is required for quality control purposes.
It will be comparable with other techniques if the dis-
tribution is unimodal, the bubbles are spherical and
monodispersed.
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S. Khuntia et al.: Microbubble-aided water purification 195
2.2 Hydrodynamic properties
The motion of microbubbles in water under gravity is
similar to that of a hard colloid particle. The rising velo-
city of a microbubble through a liquid is determined by
the balance of the buoyancy and drag forces acting on it.
In water, the Reynolds number of the microbubbles is well
below unity, and the rising velocity ( u ) can be expressed
by Stokes ’ law as (Ghosh 2009a ):
2
18lgd
uρμ
= (4)
Experimental data reported in the literature (Takahashi
2005a ) closely obey Stokes ’ law, as shown in Figure 3 .
The average bubble rising velocity (u ̅) in a bubble
column can be expressed in terms of fractional gas
hold-up ( ε g ) and superficial gas velocity ( u
s ) as (Kawahara
et al. 2009 ):
u̅=us/ε
g (5)
Fractional gas hold-up ( ε g ) is defined as the ratio of
the volume occupied by the gas phase and the combined
volume of the gas and liquid phases, i.e.:
g
g
g l
V
V Vε =
+
(6)
Superficial gas velocity in the bubble column is
defined as:
u s = Q
g /A
c (7)
The coefficient of friction of two-phase gas-liquid
flow decreases with the increase in the volume fraction
10,000
1000
100
Ris
ing
velo
city
(mm
s-1
)
1010 100
Microbubble diameter (μm)
Experimental data
Prediction from Stokes law
Figure 3 Rising velocity of microbubbles in distilled water. Experi-
mental data reported by Takahashi (2005a) are compared with the
prediction from Stokes ’ law.
of microbubbles (Cui et al. 2003 ). This effect is utilized
by blowing-in microbubbles at the bottom of large ships
(Kodama et al. 2000 , Tsuge 2010 ).
2.3 Thermodynamic properties
One of the most important characteristics of microbubbles
is that they gradually decrease in size and collapse under
water, whereas macrobubbles rise rapidly towards the
surface of the water and burst. By contrast, nanobubbles
are stable for a much longer time (sometimes for months)
(Takahashi 2005b ). The general difference between mac-
robubbles, microbubbles and nanobubbles is depicted in
Figure 4 . The decrease in size of microbubbles is due to the
dissolution of the gas present inside the bubble. The inte-
rior pressure varies inversely with the bubble diameter.
The pressure-diameter relationship can be expressed by
the Young-Laplace equation (Hiemenz and Rajagopalan
1997 ):
4-g lp p p
d
γ=Δ = (8)
Therefore, the excess pressure inside a 1- μ m dia meter
microbubble in water at 298 K (γ = 72.5 mN m−1) would be
290 kPa. A gas dissolves in water according to Henry ’ s
law: p̃ =Hc. Therefore, pressurized gas becomes efficiently
dissolved in surrounding water. When the gas dissolves
in the liquid, then the rate of shrinkage increases with
time (Takahashi 2010 ). The pressure inside the microbub-
ble goes on increasing with decreasing size, which leads
to the collapse of the microbubble. This is coupled with
the fact that microbubbles have a very slow rate of ascent
through water, and the gas inside them has a reasonably
good solubility in water. The time required for collapse of
microbubbles is of the order of several minutes (Takahashi
2010 ), although some studies have predicted their very
small lifetime in water (Katiyar and Sarkar 2010 ). Kwan
and Borden (2010) have reported that microbubbles stabi-
lized by surfactant can remain stable for several minutes,
and that dissolution time increases with the increasing
diameter of the microbubble. The characteristic of col-
lapsing in water is the source of the microbubbles ’ poten-
tial in water treatment.
From Eq. (8), nanobubbles are expected to have very
high internal gas pressures. For example, the excess pres-
sure inside a 20-nm diameter nanobubble would be 14.5
MPa. Consequently, gas in nanobubbles should rapidly
dissolve in ambient liquid. It has been theoretically shown
that the lifetime of a nanobubble of 20-nm diameter
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196 S. Khuntia et al.: Microbubble-aided water purification
would be only ∼ 1 μ s (Ljunggren and Eriksson 1997 ). Some
scientists have raised questions on the applicability of the
Young-Laplace equation in the nanometer dimensions,
due to the departure of surface tension from the macro-
scopic value, because of the variation of thermodynamic
quantities as a consequence of the large curvature asso-
ciated with nanobubbles (Li Juan et al. 2008 ). The appli-
cability of the Young-Laplace equation to nanobubbles
has been validated in a few studies (Matsumoto 2008 ,
Matsumoto and Tanaka 2008 ). Molecular dynamics simu-
lations have shown that the density of a nanoscale nitro-
gen bubble in water can be as large as 270 kg m −3 (Fang
and Hu 2006 ). The following equation has been developed
to predict the lifetime of nanobubbles (Li Juan et al. 2008 ):
( )
3
0
4
48 1-
g
g l
Hd
D
ρτ
γ ρ ρ=
(9)
Therefore, as the density of gas in the nanobubble
increases, the lifetime of the nanobubble can increase
dramatically. The gas in the nanobubbles might form a
new phase if its density is so large. Another important
characteristic of shrinking microbubbles is the hydrate
formation, which makes them useful in gas storage and
transportation (Takahashi et al. 2003 ). Gas hydrates
(also termed ‘ clathrates ’ ) are crystalline compounds
which occur when water forms a cage-like structure
around smaller guest molecules (Sloan 1998 ). According
to Henry ’ s law, the amount of dissolved gas around the
shrinking bubble increases with increasing gas pressure
inside the bubble. The area surrounding a microbub-
ble changes its state to favor the nucleation of hydrate.
Some of the nuclei grow to hydrate films surrounding the
bubble, and the entire bubble could change to a hydrate
particle stabilized by the thick film. It has been reported
that the formation of hydrate may impart stability to the
nanobubbles for a prolonged time (of the order of months)
(Takahashi 2009 ).
2.4 Electrical properties
The surface of microbubbles in water is charged. Being
charged, microbubbles move towards the oppositely
charged electrode when an electric field is applied. This
surface charge can be measured in terms of ζ potential,
which is the potential at the ‘ plane of shear ’ (Hiemenz and
Rajagopalan 1997 , Ghosh 2009a ). Although the location of
the plane of shear is not precisely known (and therefore
the value of ζ potential may be different from the actual
potential at the surface), it is the only quantity that can be
experimentally measured. The ζ potential is determined
Bursts (coalesces) atthe air-water interface
Rapidly rises towardthe water surface
Ultimately disappearsin water by dissolution
Graduallyshrinks in water
Stable formonths
MicrobubbleMacrobubble
Water surface (air-water interface)
Nanobubble
Figure 4 Difference between macrobubble, microbubble and nanobubble in terms of their behavior in water.
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S. Khuntia et al.: Microbubble-aided water purification 197
by measuring the electrophoretic mobility of the micro-
bubble and then applying the Smoluchowski equation
(Ghosh 2009a ):
0
Eμζ
εε=
(10)
The charge at the gas-liquid interface plays a very
crucial role in the stability of the microbubbles against
coalescence with neighbor bubbles in the dispersion
(Ghosh 2009b , Srinivas and Ghosh 2011 ), and in the gen-
eration of free radicals (such as hydroxyl radicals, · OH),
which are immensely important in the oxidation of unde-
sirable inorganic and organic compounds present in
wastewater (Takahashi et al. 2007a,b , Bando et al. 2008a ,
Li et al. 2009a,b ). The ζ potential of air microbubbles in
distilled water is negative. The value reported by Taka-
hashi (2005a) and Qu et al. (2009) are -35 mV and -57.9 ± 3.9
mV, respectively at pH = 5.8. The latter authors had added
a very small amount of NaCl (i.e., 0.05 mol m −3 ) to the dis-
tilled water. Hasegawa et al. (2009) have generated ozone
microbubbles using two microbubble generators with
π /3 and π /6 radian slit angles. They observed very small
variations in the value of ζ potential (i.e., -43 and -38 mV,
respectively) for these two slit angles.
The ζ potential does not significantly vary with bubble
diameter (Takahashi 2005a , Hasegawa et al. 2009 ), which
is expected from the theory of electrostatic double layer.
However, when microbubbles shrink in water, the tran-
sient value of ζ potential shows a significant variation
with time and, hence, with bubble diameter (Takahashi
2010 ). When the bubble shrinks, the ions at the air-
water interface are concentrated into a narrower region.
Therefore, they tend to diffuse into the bulk water from
the interface. However, the rate of diffusion of the ions is
slower than the rate of shrinkage of the bubble. Thus, the
ions cannot escape fast enough and become concentrated
near the interface, which is manifested by the increase in ζ
potential. It has been reported that the concentrated ions
at the bubble boundary prevent escape of the internal gas
and stabilize nanobubbles (Takahashi 2009 , Tsuge 2010 ).
When the microbubble collapses, an ionic field of very
high ion concentration is created that helps the forma-
tion of free radicals, which are the main reactive species
in the oxidative treatment of wastewater (Takahashi et al.
2007a ). When an external stimulus, such as an ultrasonic
wave, is applied to water, microbubbles form by acoustic
cavitation. As the pressure inside the bubble is inversely
proportional to the bubble diameter, rapid shrinking
(pressure collapse) leads to a sharp increase in pressure.
When the rate of such pressure increase is sufficiently
40
0
-40
-80
Zeta
pot
entia
l (m
V)
-1202 7
pH
Experimental data
12
Figure 5 Variation of ζ potentials of microbubbles with pH
(Takahashi 2005a ).
high, the temperature within the microbubble also rises
sharply because of adiabatic compression. As a result,
at the time of microbubble collapse, an area with a pres-
sure of a few thousand atmospheres and a temperature
of a few thousand degrees is created. This extreme reac-
tion site (termed ‘ hot spot ’ ) is a very small area but has
enough potential to force the decomposition of gas mole-
cules in the microbubble, which creates the free radicals
(Takahashi 2010 ).
The ζ potential strongly depends on the pH of the
medium. Takahashi (2005a) has reported that the ζ
potential of microbubbles in distilled water is negative
above pH = 4.5. In highly alkaline solutions (pH > 10), the
ζ potential exceeds -100 mV. However, below pH = 4.5,
the ζ potential is positive. These results are shown in
Figure 5 . The charge on the gas-water interface is devel-
oped due to the adsorption of OH − and H + ions. The
counterions are attracted towards the interface and
form an electrostatic double layer. At high pH, where the
solution is alkaline, the OH − ions adsorb at the air-water
interface rendering the microbubble negatively charged.
In acidic solution, there is an excess of the H + ions over
the OH − ions at the interface. The presence of inorganic
salts (e.g., NaCl and MgCl 2 ) makes the ζ potential less
negative, depending on their concentration (Takahashi
2005a , Srinivas and Ghosh 2011 ).
The ζ potential of a microbubble significantly depends
on the presence of surface active compounds (e.g., sur-
factants and alcohols) in water. These compounds have
a high concentration at the air-water interface even when
they are present in small quantities. Therefore, they easily
displace the OH − and H + ions from the interface, because
the surface activity of these ions is much smaller. The
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198 S. Khuntia et al.: Microbubble-aided water purification
cationic surfactants present in water render the ζ poten-
tial positive, whereas the anionic surfactants give rise to
negative ζ potential (Yoon and Yordan 1986 ). Aliphatic
alcohols, when present in small amounts, do not signifi-
cantly alter the ζ potential of microbubbles (Elmahdy
et al. 2008 ). Wastewater contains various surface active
compounds released from animals and plants. Therefore,
the electric charge of microbubbles in industrial wastewa-
ter can vary widely.
It has been reported that free radicals are also gener-
ated from collapse of air, oxygen and ozone microbubbles
(Takahashi et al. 2007b ). The shrinking rate of collapsing
microbubbles is much slower than that of ultrasound-
induced cavitation. Therefore, the shrinking speed of
collapsing microbubbles is not sufficient for generating
adiabatic compression. There is lack of sufficient data for
a detailed explanation of the appropriate mechanism of
generation of hydroxyl radicals from collapsing micro-
bubbles. However, from many experiments, for example,
electron spin resonance spectrum and poly vinyl alcohol
(PVA) decomposition, Takahashi et al. (2007b) have given
a probable theory of generation of free radicals from col-
lapsing microbubbles. They have observed that there is
an increase in ζ -potential with shrinkage of microbub-
bles under highly acidic conditions. The low pH caused
a change in the ζ -potential of microbubbles, which accel-
erated the collapse speed due to the reduction in electro-
static repulsion between opposite sides of the bubble wall.
If the speed of collapse is high enough, due to adiabatic
compression, a hot spot will cause the generation of high
temperature and thus the free radical may be generated.
Takahashi et al. (2007b) observed that the acoustic cavita-
tion had a negligible effect on the decomposition of PVA,
whereas ozone microbubbles effectively decomposed
PVA. It implies that high temperature is not responsible
for increased generation of OH radicals by ozone micro-
bubbles under strongly acidic conditions. Therefore, they
concluded that during the collapse of the microbubble,
some excess ions become trapped at the air-water inter-
face. This high ion concentration at the surface of micro-
bubbles increases ζ -potential. This extreme accumulation
of ions at the site of the collapsed microbubbles is suffi-
cient for the generation of free radicals from air, oxygen or
ozone microbubbles.
3 Microbubble generators (MBGs) Microbubbles have been generated by a wide variety
of methods (Sebba 1985 , Michelsen and Sebba 1994 ,
Ohnari et al. 1999 , Ohnari 2000, 2002 , Sadatomi 2003 ,
Matsuyama et al. 2006 , Sadatomi et al. 2007 , Shakouchi
et al. 2007 , Takahashi 2009 , Ikeura et al. 2011 , Terasaka
et al. 2011 ). The method of generation of microbubbles
has an effect on their properties and consequently on
the effectiveness of wastewater treatment. To illus-
trate, generation of fine MNBs of high number density
is required in wastewater treatment. Basically, there are
two types of MBGs, that is, gas-water circulation and
pressurization-followed-by-decompression. In the gas-
water circulation type of MBG, the gas is introduced into
the water vortex, and the gas bubbles thus formed are
converted into microbubbles by breaking the vortex. In
the pressurization-decompression type of MBG, a suffi-
cient amount of gas is dissolved in water under moder-
ately high pressure to form a supersaturated solution.
The solution is unstable and the gas escapes from water
generating a large number of microbubbles.
The spiral-liquid-flow MBG (Ohnari et al. 1999 ) is
shown in Figure 6 A. The pumped water is tangentially
introduced from a side hole into a cylinder. The spiral
flow of liquid forms a maelstrom-like cavity in the cylin-
der (Terasaka et al. 2011 ). The gas is sucked from an orifice
on the bottom and then spouts out with the liquid from
a hole situated at the top of the cylinder. The microbub-
bles are generated by the centrifugal force imparted by
the high-speed rotating liquid. The rotational speed lies
in the range of 1885 – 3770 rad s−1. The gas flow rate/liquid
flow rate ratio is in the range of 1/7 to 1/15. One of the early
MBGs of this category, M2-LM, has been used for oyster
culture in Japan. The details of this generator have been
described in the work of Li (2006) . By supplying the micro-
bubbles to the seawater around the oysters, their growth
rate was significantly increased. The range of bubble
diameter obtained from the M2-LM generator was 10 – 50
μ m. However, the concentration of the microbubbles was
lower than the pressurization-decompression type of gen-
erators. Li and Tsuge (2006a) connected this MBG to the
outlet of a centrifugal pump that can stably operate even
when the gas content is 7 – 10 % . Because of the whirlpool
mixing effect and increased pressure, the air was pres-
surized and dissolved in water. The air that could not dis-
solve was converted into microbubbles in the MBG. As the
microbubbles were formed in a spiral pattern, coalescence
of the bubbles after their formation could be prevented,
and the concentration of the bubbles increased.
In the venturi-type MBG (Figure 6B), a liquid stream
containing macrobubbles flows from the inlet of a venturi
tube. The two-phase flow is accelerated through the throat
of the venturi tube. The pressure changes rapidly and the
microbubbles are formed by reducing the macrobubbles
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S. Khuntia et al.: Microbubble-aided water purification 199
and/or by cavitation. The ejector-type MBG is shown in
Figure 6C. The liquid flow channels in the cylindrical gen-
erator are designed to shrink and stepwise enlarge. The
gas is self-sucked from the most reduced pressure point
and reduced to a number of microbubbles by cavitation.
In the MBG developed by Sadatomi et al. (2005) , pressur-
ized water is introduced into a pipe with a spherical body
in the core. The water velocity around the body, especially
in a downstream region, becomes higher than the inlet
velocity, and thus the pressure there becomes lower. If the
pressure becomes less than the atmospheric pressure, air
is automatically sucked into the water stream through a
number of small holes drilled on the pipe wall in the lower
pressure region. Because the water flow there is highly
turbulent, and as a result of the shear flow the air sucked
is broken into a large number of microbubbles.
A MBG method similar to those described above
is hydrodynamic cavitation (Senthilkumar and Pandit
1999 , Gogate and Pandit 2000a, 2004a , Senthilkumar
et al. 2000 , Sivakumar and Pandit 2002 , Ambulgekar et
al. 2005 , Wang et al. 2008 , Wang and Zhang 2009 , Gogate
2011 , Saharan et al. 2012 ). With this method, cavitation is
generated by the flow of liquid through a simple geometry
(e.g., a venturi tube or an orifice plate) under controlled
conditions. When the pressure at the throat ( ‘ vena con-
tracta ’ ) falls below the vapor pressure of the liquid, the
liquid flashes, generating a number of cavities. These
cavities subsequently collapse when the pressure recovers
downstream of the mechanical constriction. The collapse
of the cavitation bubbles initiates some physicochemical
effects (e.g., production of shock waves, shear forces and
chemical reactions), resulting in the intensification of the
dispersion processes. Free radicals are generated by these
processes, which are utilized in the treatment of waste-
water. Several studies (Ambulgekar et al. 2005 , Wang
et al. 2008 , Wang and Zhang 2009 , Saharan et al. 2012 )
have reported the decomposition of pesticides and dyes
by using the hydrodynamic cavitation method.
The pressurization-decompression type of MBG is
shown in Figure 6D. The gas is dissolved in the liquid in
Microbubbles
Microbubbles
Throat
Gas water
Gas
Gas
PumpVent
Pressurized section
Decompressed section
Liquid
Liquid
Pump
AB
C
D
Spiralliquid flowGas pillar
Gas
Microbubbles
Figure 6 Microbubble generators: (A) spiral liquid flow, (B) venturi, (C) ejector, and (D) pressurization-decompression (adapted from
Terasaka et al. 2011 , with permission from the copyright holder, Elsevier Publishing).
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200 S. Khuntia et al.: Microbubble-aided water purification
a tank by pressurizing the gas-liquid mixture. When this
supersaturated liquid is flashed using a reducing valve,
microbubbles are generated. The size and number of the
microbubbles depend on the pressure and decompression
process. Terasaka et al. (2011) have made a comparative
study on the performance of the four types of MBGs shown
in Figure 6, in terms of the gas hold-up and the mass trans-
fer of gas into the liquid phase. They reported that the spi-
ral-liquid-flow and pressurization-decompression types
of MBGs showed high gas hold-up. The former type also
showed high mass transfer coefficient of oxygen. MBGs
are much more efficient for transferring gas into the liquid
phase than the conventional gas distributors (e.g., perfo-
rated plate and constant-flow nozzle). However, the power
requirement of MBGs is higher because of the requirement
of the pump, which is not required by conventional gas
distributors. Ikeura et al. (2011) have reported that the
pressurization-decompression type of MBG is more effec-
tive for the decomposition of fenitrothion (a pesticide)
by ozone. The main factor that contributed to this differ-
ence in performance was the smaller size of microbubbles
generated by this type of MBG, which could come into
contact with the pesticide (which had infiltrated into vege-
tables) more easily. The variation of the fluid flow pattern
from one MBG to another can lead to the difference in
microbubble size, broadness of the size distribution, and
also the shape of the microbubble (Nouri et al. 2008 ).
Some researchers have reported that electrical properties
of microbubbles can vary with the method of their genera-
tion (Hasegawa et al. 2008 ), which can have an important
influence on the treatment of wastewater.
Maeda et al. (2010) have studied the effect of liquid
volumetric flux and flow pattern in the decompression
nozzle on the mean diameter and number density of
microbubbles generated by a pressurization-decompres-
sion MBG. They observed that very few microbubbles were
generated when there was no cavitation. When the bubble
cavitation took place, the mean bubble diameter was low,
which did not depend on the liquid flux, and the bubble
number density increased with the flux. For sheet cavita-
tion, the mean bubble diameter increased with the liquid
flux, whereas the number density remained invariant.
Bredwell and Worden (1998) have used a spinning-
disk MBG employing a high-speed electric motor. A disk
of 5 cm diameter and 1 cm thickness was spun at speeds
above 419 rad s−1. Stationary baffles located within 5 mm of
the spinning disk created a localized high-shear zone. The
stainless steel disk and baffles were mounted in a fermen-
tation vessel of 6 dm 3 capacity. A nonionic surfactant (i.e.,
Tween 20) was used to stabilize the microbubbles. The
surfactant molecules adsorbed at the air-water interface
formed a monolayer. This layer played an important role
in the stability of the microbubbles ( D ’ Arrigo 1984 ).
Several other methods have also been employed to
generate microbubbles. For example, porous glass mem-
branes have been used for the generation of microbubbles
(Fujikawa et al. 2003 , Kukizaki et al. 2004 , Kukizaki 2006 ,
Kukizaki and Goto 2006, 2007 , Kukizaki 2009 , Trushin
et al. 2011 ). Kukizaki et al. (2004) have used a sintered
porous glass membrane with a mean pore size of 84 nm
to produce MNBs of 720 nm mean diameter. They added
a surfactant to prevent coalescence of the bubbles. They
observed that the mean diameter of the bubbles was pro-
portional to the mean pore diameter of the membrane.
The type of surfactant was found to be important in the
size of the microbubbles (Kukizaki and Baba 2008 ). Mitani
et al. (2005) have used a microporous diffuser system that
produced sub-micron ozone MNBs. The diffuser extended
concentrically throughout the length of a cylindrical
reactor in order to increase the contact time between dis-
solved ozone and the pollutants present in water, as well
as to increase the interfacial area between the gas and
the liquid. Coalescence occurred in the top portion of the
reactor. Fine bubbles appeared throughout the length of
the reactor, thus creating a very large gas-liquid interfa-
cial area, which enhanced the mass transfer rate of ozone.
Zhang et al. (2001) have used a rotating porous plate at
the air intake unit of the device. Compressed air was taken
into the device through the rotating porous plate. The size
of the bubbles decreased with increasing rotational speed,
which increased the dissolution of air into the water.
Microbubbles of uniform diameter have been pro-
duced by exposing the bubbles formed by a needle-like
narrow tube to ultrasonic waves (Tsuge 2010 ). Micro-
bubbles were produced in a high-viscous liquid (e.g.,
silicone oil), and had a diameter of ∼ 10 μ m. Microbub-
bles have also been produced by condensation of mixed
vapors (Terasaka et al. 2009 ). The advantage of this
method is that the power requirement is less, compared
with mechanical MBGs. Nitrogen microbubbles (with the
peak of the diameter distribution in the range of 20 – 40
μ m) have been generated by blowing a mixture of nitro-
gen and steam into water through a fine nozzle (Tsuge
2010 ). The steam condensed, but the nitrogen did not
condense. The size of the microbubbles changed depend-
ing on the gas composition, inside diameter of the nozzle
and vapor velocity. Tsuge et al. (2008) have performed
electrolysis of water and demonstrated the effect of stir-
ring (at low frequency) on the formation of microbubbles
of hydrogen and oxygen from the electrodes. The range of
diameter of the microbubbles was 15 – 100 μ m. Shin et al.
(1997, 1999) have generated microbubbles by electrostatic
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S. Khuntia et al.: Microbubble-aided water purification 201
spraying. They observed that the bubble size decreased
with decreasing Reynolds number and increasing Weber
number. Three different modes of bubble formation were
observed: a spraying mode at low airflow rate and high
applied voltage, a dripping mode at high flow rate and low
applied voltage, and a mixed spraying-dripping mode.
Walker et al. (2001) have reported that the electrostatic
spraying method is suitable for generating microbubbles
in low-conductivity solutions, but not suitable at high-salt
concentrations due to the high electric current. Small-
pore gas diffusers are more suitable in these solutions.
The acoustic cavitation method is another technique
that generates MNBs. Although widely used in various
laboratory experiments, this method has hardly been
used in large-scale water treatment, due to high cost
and expertise required in diverse areas such as chemical
engineering, materials science and acoustics (Gogate and
Pandit 2000a ). With this method, cavities are produced
by passing sound waves, usually ultrasound ( > 16 kHz),
through the liquid medium. The passage of ultrasound
through the aqueous medium generates cavities, pro-
motes their growth and their collapse. The entire process
of cavity generation, growth and collapse occurs over
a period of microseconds (Gogate and Pandit 2000b ).
During the collapse of the bubble, a high pressure (e.g.,
several hundred times that of the atmospheric pressure)
and high temperature (e.g., several thousand Kelvin) are
generated (Leong et al. 2011 ). These extreme reaction
sites force decomposition of the gas molecules in the
bubble, which generates free radicals (e.g., ⋅OH). It has
been reported in the literature that ultrasound sonication
generates microbubbles that are smaller in diameter than
those generated mechanically, and the size distribution is
narrower. Several studies have reported the use of acoustic
methods for purification of wastewater (Dahl 1976 , Suslick
et al. 1986 , Teo et al. 2001 , Sivakumar and Pandit 2002 ,
Sivakumar et al. , 2002 , Goel et al. 2004 , Gogate and Pandit
2004a,b , Adewuji 2005 , Sivasankar and Moholkar 2009 ,
Laxmi et al. 2010 ). In-depth reviews of wastewater treat-
ment by acoustic methods have been presented in these
studies. In this review article, details of the ultrasonic
methods of wastewater treatment have not been covered.
Riverforest Corporation (USA) has manufactured
different types of MBGs ranging from small-, medium-
and large-scale applications. The MBG model ASMK-III
is a pressurized system which can be used for lab-scale
study (microbubble output = 6.7 × 10 −4 m 3 s−1). Other models,
SMX-115 and SMX-155, are medium capacity MBGs with
microbubble flow rates of 8.3 × 10 −4 m 3 s−1 and 0.001 m 3 s−1,
respectively. The swirling type Tornado MB nozzle is ope-
rated at a minimum pressure of 0.04 MPa and flow rates
in the range of 5 × 10 −4 m 3 s−1 and 0.001 m 3 s−1, which are
used efficiently for small-scale applications. Different
types of MBGs with nozzles at high air-water ratio are also
available for large-scale applications. For special uses of
ozone, oxygen, CO 2 and hydrogen microbubbles, the MB
nozzle with the MB400 pressure chamber may be suit-
able to produce very thick and uniform microbubbles.
Negatron Co. Ltd. (Korea) also manufactures many MBGs
for various applications with a wide range of micro-
bubble size (0.01 < 300 μ m) at flow rates of 4.17 × 10 −4 –
0.033 m 3 s−1.
4 Transport of gas into the liquid phase
One of the most important aspects of the use of micro-
bubbles in water treatment is the transfer of gas from the
microbubble into the liquid. In oxidation applications
(such as ozonation), the gas present in the microbubble
must dissolve in the surrounding aqueous phase before
reacting with inorganic and organic contaminants. By
contrast, in applications such as the growth of aquatic
life, or in the fermentation reactors, the transfer of oxygen
into water is essential. According to the ‘ two-film (or two-
resistance) theory ’ of mass transfer (Lewis and Whitman
1924 ), the gas-liquid interface lies between a gas film and
a liquid film, as shown in Figure 7 . The overall mass trans-
fer coefficients in the liquid and gas phases, i.e., K l and K
g
are given by (Cussler 1997 ):
1 1 1 1
l g l gK K H k k H= = +
(11)
The methods of determination of Henry ’ s law constant
( H ) for gases such as ozone have been outlined in the lite-
rature (Sotelo et al. 1989 , Kuosa et al. 2004 , Bi ń 2006 ). For
microbubbles, the gas molecules have to diffuse through
a small distance and, therefore, gas-phase mass trans-
fer resistance, i.e., 1/( k g H ), is negligible (Motarjemi and
Jameson 1978 ). Also, for the sparingly soluble gases such
as oxygen and ozone, mass transfer resistance mainly lies
in the liquid phase (Johnson and Davis 1996 ). Thus, it can
be assumed that K l ≈ k
l .
For a microbubble rising in water following Stokes ’
law, that is, Eq. (4), the mass transfer coefficient can be
calculated from the following correlation (Clift et al. 1978 ):
{ }1 3
1 1l
D duk
d D
⎡ ⎤= + +⎢ ⎥
⎣ ⎦ (12)
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202 S. Khuntia et al.: Microbubble-aided water purification
Eq. (12) shows that the mass transfer coefficient
increases with decreasing size of the microbubble. Waste-
water often contains surface active impurities. These
compounds quickly adsorb on the surface of the micro-
bubbles and immobilize the surface. The mass transfer
rate is reduced by this process. Apart from the reduction
in surface flow, surfactant molecules act as a physical
barrier for the gas molecules to pass through the inter-
face (Koide et al. 1976 ). In addition, the surfactant con-
centration may alter the thickness of the liquid shell of
the microbubbles, which reduces mass transfer rate. The
Fr ö ssling equation is valid for ‘ solid ’ microbubbles (i.e.,
microbubbles whose surface is immobile and behave like
solid particles) (Motarjemi and Jameson 1978 ):
1 31 2
0.6l l
l
k d du
D D
ρ μμ ρ
⎛ ⎞⎛ ⎞= ⎜ ⎟ ⎜ ⎟⎝ ⎠ ⎝ ⎠
(13)
For circulating microbubbles, the Higbie equation is
applicable (Motarjemi and Jameson 1978 ):
1 2
2l
d
Dk
tπ
⎛ ⎞= ⎜ ⎟⎝ ⎠
(14)
The mass transfer coefficients predicted by Eqs. (13)
and (14) differ considerably. Kawahara et al. (2009) have
presented a comparison of k l obtained by these two equa-
tions. Experimental data on oxygen transfer in tap water
and saline water show an increase in the mass trans-
fer coefficient with increasing bubble diameter, which
is opposite to that predicted by these equations. They
attributed this to the bubble-induced turbulence, which
is proportional to the bubble diameter (Sato et al. 1981 ).
Kawahara et al. (2009) have presented modifications
of Eqs. (13) and (14). They have reported that a unique
relationship exists between k l and the product of Sauter
mean diameter ( d 32
) and bubble rising velocity ( u ), which
is given by:
0.756
-10 327.46 10lk ud
D
μγ
⎛ ⎞= × ⎜ ⎟⎝ ⎠
(15)
The following equation, given by Calderbank and
Moo -Young (1961) , is applicable to the bubbles, which
have a diameter < 100 μ m (Motarjemi and Jameson 1978 ):
1 3
2 30.31l
gk D
ρ
μ
Δ⎛ ⎞= ⎜ ⎟⎝ ⎠
(16)
Another equation, similar to Eq. (16), has been pro-
posed by Waslo and Gal -Or (1971) for industrial disper-
sions containing small bubbles:
1 3 1 35 3
2 3
5 3
1-0.55
3 2
g
l
g
gk D
ρε
με
⎛ ⎞ Δ⎛ ⎞= ⎜ ⎟ ⎜ ⎟⎝ ⎠+⎝ ⎠
(17)
Both Eqs. (16) and (17) are applicable for small rigid-
surface microbubbles.
Bredwell and Worden (1998) have observed that the
experimentally determined values of k l were similar in
magnitude to those predicted by the correlations. The
average value of the quantity,
1 35 3
5 3
1-0.55 ,
3 2
g
g
ε
ε
⎛ ⎞⎜ ⎟+⎝ ⎠
over the
gas hold-up range upto 0.8, is approximately 0.31. Several
correlations for determination of ε g have been reported by
Kawahara et al. (2009) . A comparison of the gas hold-up
obtained by various methods has been given by Li (2006) .
As the gas continues to dissolve in the liquid phase,
its concentration in the liquid phase increases. The rate of
increase in concentration of the gas in the liquid phase is
given by (ASCE 2007 ):
( )* -l
dck a c c
dt=
(18)
In most gas-liquid contacting equipment, the interfa-
cial area is not precisely known. Therefore, the product of
Liquid
BubbleEnlargement of gas-liquid
interface
Gas-liquidinterface
pg
pi
ci
cl
Bulkgas
Bulkliquid
Gasfilm
Liquidfilm
Figure 7 Mass transfer of gas from the bubble into the surrounding liquid according to the two-film theory.
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S. Khuntia et al.: Microbubble-aided water purification 203
mass transfer coefficient ( k l ) and interfacial area per unit
volume ( a ) is used. This product, k l a , is termed ‘ volumet-
ric mass transfer coefficient ’ . The interfacial area per unit
volume is related to the fractional gas hold-up ( ε g ) and
Sauter mean diameter ( d 32
) as (Kawahara et al. 2009 ):
32
6 ga
d
ε=
(19)
The quantities, k l a and c *, are sometimes called ‘ clean
water parameters ’ . If the initial concentration of the gas in
water (i.e., at t = 0) is c 0 , then integration of Eq. (18) gives
the variation of concentration of the gas in the aqueous
phase:
*
*
0
-ln -
-l
c ck at
c c
⎛ ⎞=⎜ ⎟⎝ ⎠
(20)
The volumetric mass transfer coefficient, k l a , can be
determined from the slope of the plot of
*
*
0
-ln
-
c c
c c
⎛ ⎞⎜ ⎟⎝ ⎠
vs.
time (Ago et al. 2005 ), or by a non-linear regression analy-
sis, as described by ASCE (2007) . Several correlations for
the computation of k l a have been reported in the literature
(Akita and Yoshida 1973 , Van ’ t Riet 1979 , Koide et al. 1983 ,
Bredwell and Worden 1998 , Nedeltchev et al. 2006a,b ,
Kawahara et al. 2009 ). However, only a few studies have
reported the applicability of these correlations for predict-
ing the volumetric mass transfer coefficient in microbub-
ble systems. The available experimental data indicate
that these correlations are not accurate for the predictive
purpose.
The transfer of gas is expressed as the volumetric
transfer rate (VGTR) of gas (Chu et al. 2008a ):
VGTR = k l a ( c * - c ) (21)
Experimental data reported in the literature on the
microbubble systems indicate that high values of volu-
metric mass transfer coefficient can be obtained even
in the (mechanically) unagitated systems. For example,
Bredwell and Worden (1998) have reported k l a values for
oxygen microbubbles (which had an average initial dia-
meter of 60 μ m) in the range of 0.06 – 0.5 s −1 . These values
are considerably higher than those for mechanically
agitated, commercial-scale fermentors. They have demon-
strated the enhancement of mass transfer using micro-
bubbles for synthesis-gas fermentation. Butyribacterium
methylotrophicum was grown in a continuous, stirred-tank
reactor. The k l a value for microbubble sparging was six
times larger than the conventional gas sparging through
a 10- μ m stainless steel frit, which produced 0.5 – 1 mm
0.009
0.007
0.005
k la (s
-1)
0.003
0.001 0.001
0.004
0.007
VO
TR (kg m
-3 s-1)
Air flow rate (dm3 s-1)
0.013
0.010
0 0.005 0.010 0.015 0.020 0.0300.025
Figure 8 Variation of k l a and volumetric oxygen transfer rate (VOTR)
with air flow rate (Chu et al. 2008a ).
diameter bubbles. The gas flow rate for the microbubbles
was only approximately one-half of that for conventional
sparging. Thus, the large specific interfacial area for the
microbubbles resulted in high values of k l a . Similar results
have been reported by Kaster et al. (1990) . Ago et al. (2005)
have measured k l a for carbon dioxide microbubbles. They
observed an increase in the volumetric mass transfer coef-
ficient by several times compared with the conventional
bubbling technique. Chu et al. (2007a) have studied mass
transfer of ozone in water by microbubbles. Compared
with an ordinary bubble contactor, the mass transfer effi-
ciency in the microbubble system was 1.6 – 2.7 times higher.
Increase in the flow of gas led to the increase in the volu-
metric mass transfer coefficient and the gas transfer rate,
as illustrated in Figure 8 (Chu et al. 2008a ).
Nakano et al. (2005) have compared the volumetric
mass transfer coefficients of oxygen in water by the micro-
bubble and air-stone dissolution methods. The MBG pro-
duced a much higher k l a than the air-stone method. As the
gas flow rate was increased, the volumetric mass transfer
coefficient increased considerably. Li and Tsuge (2006b)
have reported that the volumetric mass transfer coefficient
for ozone in water increased with the increasing induced-
gas and water flow rates. Yasuda et al. (2010) have studied
mass transfer of ozone in water in an airlift bubble column
reactor with a draft tube. They observed that k l a increased
with increasing diameter of the draft tube.
5 Ozonation using microbubbles Oxidation processes constitute a major step in the treat-
ment of wastewater. Ozone is applied to remove the
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204 S. Khuntia et al.: Microbubble-aided water purification
organic and inorganic compounds present in the waste-
water. Despite several advantages of using ozone, it has
a few disadvantages, which limit its application in water
treatment. The main drawbacks of ozone are its relatively
low solubility and stability in water. In addition, other
factors that have limited the use of ozone are the high cost
of production and only a partial oxidation of the organic
compounds present in water. The partial oxidation can
lead to the generation of carcinogenic products. However,
often, partial oxidation of the organic compounds leads to
biodegradable products (e.g., carboxylic acids, aldehydes
and ketones) (Glaze and Weinberg 1993 ), which can be
removed by adsorption on activated carbon. The reaction
of ozone with the pollutants in water is rather slow (Taki ć
et al. 2008 ), and the overall reaction rate can be affected
by both the reaction kinetics and mass transfer (Zhou and
Smith 2000 ). Several advanced ozonation and catalytic
ozonation processes have been attempted to render the
use of ozone commercially viable (Gunten 2003 , Kasprzyk -
Hordern et al. 2003 , Ikehata et al. 2008 ). The mechanism
of decomposition of ozone in water is presented in Table 1
(Kasprzyk -Hordern et al. 2003 , Beltr á n 2004 ).
To alleviate some of the drawbacks of conventional
ozonation processes mentioned above, microbubble-aided
ozonation has been successfully attempted by several sci-
entists. Many organic and inorganic compounds react with
ozone or hydroxyl radicals directly or indirectly (Gunten
2003 ). Catalysts are also used to enhance the free radical
generation from ozone. Free hydroxyl radicals are more
powerful than molecular ozone for oxidation. In addi-
tion, ozone microbubbles are also capable of generating
hydroxyl radicals under certain conditions as discussed in
Section 2.2. The significant increase in ion concentration
Reaction Rate constant
Initiation
O3+OH−→HO
2⋅+O−
2⋅ 70 (dm 3 mol −1 s −1 )
Propagation
HO2⋅→O−
2⋅+H+ 7.9 × 10 5 s −1
O−2⋅+H+→HO
2⋅ 5 × 10 10 (dm 3 mol −1 s −1 )
O3+O−
2⋅→O−
3⋅+O
21.6 × 10 9 (dm 3 mol −1 s −1 )
O−3⋅+H+→HO
3⋅ 5.2 × 10 10 (dm 3 mol −1 s −1 )
HO3⋅→O−
3⋅+H+ 3.3 × 10 2 s −1
HO3⋅→HO⋅+O
21.1 × 10 5 s −1
O3+HO⋅→HO
42 × 10 9 (dm 3 mol −1 s −1 )
HO4⋅→HO
2⋅+O
22.8 × 10 4 s −1
Termination
HO4⋅+HO
4⋅→H
2O
2+2O
35 × 10 9 (dm 3 mol −1 s −1 )
HO4⋅+HO
3⋅→H
2O
2+O
2+O
35 × 10 9 (dm 3 mol −1 s −1 )
Table 1 Mechanism of decomposition of ozone in pure water
(Kasprzyk-Hordern et al. 2003, Beltr á n 2004).
around the shrinking gas-water interface helps generation
of free radicals (Takahashi et al. 2007a ). In fact, in recent
times, a major application of microbubbles in wastewater
treatment involves ozonation. This involves decoloriza-
tion (e.g., removal of dyestuff), degradation of pesticides
and other harmful organic compounds, and removal of
odor (e.g., residual ammonia) from water/wastewater.
Some of the applications of ozone microbubbles are speci-
fied in Table 2 .
Air, oxygen and nitrogen microbubbles are also used
for decomposition of various compounds. Some ozone
resistant compounds, for example, PVA, dimethyl sul-
foxide (DMSO), can be decomposed using hydroxyl radi-
cals. Ozone microbubbles under strong acidic conditions
generate hydroxyl radicals, which can be effectively used
for decomposition of PVA and DMSO (Li et al. 2009c ).
Takahashi et al. (2007a) have reported that air microbub-
bles also generate hydroxyl radicals. Li et al. (2009c) have
used air and ozone microbubbles for the study of decom-
position of DMSO. They concluded that the amount of free
radicals generated from air microbubbles is less than that
of ozone microbubbles. Li et al. (2009b) have reported
that phenol degradation can be effectively done by free
radicals generated from air, oxygen and nitrogen micro-
bubbles. They observed that for 2 h duration of phenol-
decomposition, the rate of decomposition by free radicals
generated from nitrogen, air and oxygen microbubbles
was increased by 36 % , 59 % and 83 % , respectively.
5.1 Removal of color
Wastewaters released from the textile and dye manu-
facturing industries are heavily colored, a major part of
which is non-biodegradable in nature. Ozonation has
been used by various researchers for the removal of color
(Shu and Huang 1995 , Wang et al. 2003 ). Microbubbles
enhance the efficiency of ozonation. Chu et al. (2007b)
have ozonated aqueous solutions of CI Reactive Black 5
(an azo dye of molecular formula C 26
H 21
N 5 O
19 S
6 · 4Na) using
microbubbles. The average diameter of the microbubbles
was 58 μ m. They have compared the performance of the
microbubble system with a conventional bubble contac-
tor. The reaction rate was much higher in the microbubble
system. They obtained good efficiency of removal of color
and observed that all color was removed within 1.8 ks. A
test using terephthalic acid as the chemical probe implied
that more hydroxyl radicals were produced in the micro-
bubble system, which contributed to the degradation of
dye molecules. Some of the results from their study are
presented in Figure 9 .
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S. Khuntia et al.: Microbubble-aided water purification 205
Cate
gory
of i
mpu
rity
Type
of e
fflue
ntIm
purit
ies f
or re
mov
alM
etho
d of
mic
robu
bble
ge
nera
tion
Size
of
mic
robu
bble
( μ m
)Re
mov
al ( %
)Ti
me
(ks)
Refe
renc
es
So
lub
le o
rga
nic
sP
etr
ole
um
in
du
str
ial
wa
ste
wa
ter
BTE
XE
lect
ros
tati
c s
pra
yin
g1
0 –
80
83
0.6
Wa
lke
r e
t a
l. (
20
01
)
Ch
lori
na
ted
org
an
ic c
om
po
un
dU
nd
erg
rou
nd
wa
ter
TCE
Air
sh
ea
rin
g M
BG
10
– 3
01
00
7.5
Na
ka
no
et
al.
(2
00
5)
Co
lor
Tex
tile
wa
ste
wa
ter
CO
DS
pir
al
liq
uid
flo
w M
BG
< 5
87
01
2C
hu
et
al.
(2
00
7a
)
CI
rea
ctiv
e b
lack
Sp
ira
l li
qu
id f
low
MB
G <
58
99
1.8
Ch
u e
t a
l. (
20
07
b)
De
com
po
sit
ion
of
slu
dg
eS
ew
ag
e d
isp
os
al
pla
nt
Ph
os
ph
oru
s a
nd
nit
rog
en
Ca
sca
de
typ
e
Pre
ss
uri
zati
on
pu
mp
50
– 7
05
03
.6B
an
do
et
al.
(2
00
8b
)
Wa
ste
wa
ter
tre
atm
en
t p
lan
tS
lud
ge
so
lub
iliz
ati
on
Sp
ira
l li
qu
id f
low
MB
G <
58
80
1.2
Ch
u e
t a
l. (
20
08
b)
De
terg
en
t/p
ho
tore
sis
t
str
ipp
ing
so
lve
nt
Se
mic
on
du
cto
r
ma
nu
fact
uri
ng
in
du
str
y
DM
SO
Ro
tati
ng
flo
w M
BG
50
a
70
0.6
Li e
t a
l. (
20
09
c)
Pe
sti
cid
eV
eg
eta
ble
s (
lett
uce
)R
es
idu
al
fen
itro
thio
nD
eco
mp
res
sio
n1
0 a
55
0.6
Ike
ura
et
al.
(2
01
1)
Ga
s-w
ate
r ci
rcu
lati
on
40
a
45
0.6
Tabl
e 2
Us
e o
f o
zon
e m
icro
bu
bb
les
in
wa
ste
wa
ter
tre
atm
en
t.
a M
ea
n b
ub
ble
dia
me
ter.
00 1000 2000 3000 4000 5000
Time (s)
With microbubble generatorWith bubble contactor
20
40
60
Col
or re
mov
al e
ffici
ency
(%)
80
100
Figure 9 A comparison of the performance of a microbubble gen-
erator and a porous gas diffuser system in the removal of color from
wastewater (Chu et al. 2007b ).
Chu et al. (2007a, 2008a) have treated textile waste-
water collected from a plant that contained residual azo
dyes, alkali and surfactants. They treated this wastewa-
ter using ozone microbubbles. They have reported that
the input ozone could be almost completely utilized by
the microbubble system, and the rate of decolorization
and organic reduction were much faster than those of a
conventional bubble contactor. For the textile wastewater
tested by them, 80 % of color was removed in 84 ks by the
ozone microbubbles, whereas the conventional bubbles
took double this time. The chemical oxygen demand
(COD) reduction efficiency in the microbubble system was
higher by 20 % .
5.2 Removal of pesticides
Residues from the pesticides used in agriculture persist
in soil for extended periods before they are completely
degraded by the natural processes. These compounds
migrate into plants and animals, and into water. The pro-
duction and use of such pesticides are extensive and inev-
itable. At the same time, health hazards caused by them
are considerable. Soil and water are the ultimate reposi-
tories of most of the residue pesticides. The common
methods of removal of pesticide residues from water are
adsorption on the sediments and activated charcoal, bio-
logical trapping by algae in sewage lagoons, oxidation by
ozone and permanganate, and separation by nanofiltra-
tion membrane and ion exchange (Street 1969 ).
In recent years, several researchers have used ozone
microbubbles for the degradation of residual pesticide
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206 S. Khuntia et al.: Microbubble-aided water purification
in wastewater. Ikeura et al. (2011) have investigated the
removal of residual fenitrothion. Lettuce leaves, cherry
tomatoes and strawberries were immersed in an aqueous
solution containing the pesticide. The pesticide was
allowed to infiltrate into the vegetables for a day. There-
after, they were washed in tap water for a minute and
treated with ozone. Ozone microbubbles were formed with
dechlorinated tap water. Ozone generation was stopped
when the concentration of dissolved ozone reached
2 mg dm −3 . The vegetables were immersed in the solutions
and readings were taken at different times for 600 s. The
amount of pesticide removed by this process differed from
one vegetable to another, as illustrated in Figure 10 . For
lettuce, the residual pesticide was removed efficiently.
By contrast, the dissolved ozone and the hydroxyl radi-
cals could not penetrate through the thick pericarp of
the cherry tomatoes and, therefore, could not reach the
sarcocarp. They were inactivated by coming into contact
with the pericarp. This led to a small removal in the case
of the tomatoes. The strawberries had a rougher surface
and hence a larger surface area, which could easily come
into contact with ozone, and hence the pesticide removal
was better for the strawberries than the tomatoes. The
pesticide removal efficiency also depended on the micro-
bubble generation technique, which has been discussed
in Section 3.
5.3 Removal of other compounds
Chemicals discharged into water from petroleum indus-
tries contain various dissolved organic compounds and
120
100
LettuceStrawberryCherry tomato
80
60
Res
idua
l fen
itrot
hion
(%)
40
20
00 200 400
Time (s)
600 800
Figure 10 Residual percentages of fenitrothion for lettuce, cherry
tomato and strawberry after immersion in solutions containing
ozone microbubbles (Ikeura et al. 2011 ).
high concentrations of salt. Walker et al. (2001) have
studied the degradation of benzene, toluene, ethylbenzene
and xylenes (BTEX) by ozone microbubbles. They have
reported that the percentage removal of BTEX increased
with increasing ozone input concentration. For every mole
of hydrocarbon oxidized, at least 3.8 mol of ozone was
required. The ozone-to-BTEX molar ratio increased with
increasing gas flow rate, indicating that multipoint injec-
tion might be beneficial for the degradation of hydrocar-
bons. The ratio of ozone input/BTEX removed increased
with increasing salt concentration, which suggests that
the salt probably catalyzed the destruction of ozone.
Boncz et al. (2005) have reported that in some cases (e.g.,
at low pH) inorganic anions may stimulate ozone disso-
ciation and radical formation. In such cases, a decrease
of oxidation selectivity may occur. In more alkaline solu-
tions, inorganic anions may act as radical scavengers and
thus may cause an increase in oxidation selectivity. The
increase in selectivity, however, will be accompanied by a
loss of efficiency, because the radicals removed from the
system by these scavenging reactions do not contribute to
the oxidation process.
Chlorinated organic compounds, e.g., trichloroethyl-
ene (TCE) present in wastewater, do not easily decompose
when ozone is supplied to it. In a procedure developed
by Nakano et al. (2005) , these compounds were adsorbed
on activated charcoal. After that, they were extracted
and concentrated from activated carbon by using a suit-
able organic solvent. Ozone microbubbles were then
passed through the solution. The treatment efficiency
was improved, because the organic compound (TCE)
was concentrated in the organic solvent. The amount of
waste due to processing can be decreased by recycling the
reproduced activated carbon and organic solvent. Acetic
acid and an acetate solution were found to be effective
solvents for activated carbon regeneration and ozone
treatment (Nakano et al. 2005 ). The solubility of ozone in
acetic acid is much higher than that in water. Therefore, a
larger amount of ozone was transferred to the acetic acid
medium, which could react with TCE to oxidize chloride
into chlorine. During the reaction of TCE with ozone in
water, carbonyl chloride (which is a harmful compound)
is produced as an intermediate. By contrast, TCE in acetic
acid was first transformed to chlorinated intermediates
and chloride ions and, finally, the chlorinated intermedi-
ates were dechlorinated almost completely and oxidized
to chlorine gas (Tsai et al. 2004 ). It has also been reported
in the literature that TCE was completely dechlorinated
in acetic acid and 60 % acetate solution, and the solvent
could be recycled without accumulating harmful chemi-
cals such as carbonyl chloride (Nakano et al. 2005 ).
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S. Khuntia et al.: Microbubble-aided water purification 207
Ozone is a strong cell-lytic agent, which can kill
microorganisms present in activated sludge and further
oxidize organic substances released from the cell. Sludge
disintegration by ozone can be described as a sequence
of decomposition reactions of cell destruction, solubiliza-
tion and subsequent oxidation of released organics into
carbon dioxide (Chu et al. 2008b ). Bando et al. (2008b)
have investigated the decomposition of sludge (obtained
from a sewage disposal plant) by ozone microbubbles.
They have studied the influence of ozone microbubbles
on foam behavior and decomposition performance. A
concurrent upflow bubble column with a MBG was used
and the sludge was decomposed under various condi-
tions. Decomposition was enhanced in the presence of
ozone, that is, the decomposition time was shorter and the
required dose of ozone was also lower.
Chu et al. (2008b) have studied ozone-decomposition of
sludge collected from a wastewater treatment plant, which
adopts an anoxic-anaerobic-oxic process. They observed
that the microbubble ozonation system was effective in
increasing the ozone utilization and improving sludge solu-
bilization. An ozone utilization efficiency of more than 99 %
was obtained using the microbubble system for a contact
time of 4.8 ks. By contrast, efficiency gradually decreased
from 94 % to 72 % for a conventional bubble contactor. The
rate of microbial inactivation was faster in the microbubble
system. Compared with the bubble contactor, the micro-
bubble system released more than two times COD and total
nitrogen, and eight times total phosphorus content from
the sludge, by using the same ozone dosage.
The semiconductor manufacturing industry releases
wastewater containing a large quantity of DMSO, which
originates from the washing and rinsing processes. Li et
al. (2009c) have studied the oxidation of DMSO by ozone
microbubbles in a bubble column reactor. They have
studied the effects of gas and liquid flow rates on DMSO
degradation. It was found that ozonation of DMSO is a
first-order reaction, which is controlled by mass transfer.
The reaction rate constant increased with increasing gas
velocity. The variation of DMSO concentration with time
and ozone flow rate is shown in Figure 11 . The results
clearly show a more rapid decrease in the concentration
of DMSO with increasing ozone flow rate.
Liu et al. (2011) have compared efficiencies of air,
oxygen and ozone microbubbles in the treatment of coke
wastewater. Microbubble flotation with ozone showed the
best performance. The ozone microbubbles exhibited high
absolute values of ζ potential, creating electrostatic repul-
sive forces which prevented coalescence of bubbles. This
also created attractive interaction between the bubbles and
the particles present in the wastewater. The fluorescence
1.0
0.8
0.6
0.44.17 cm3 s-1
8.33 cm3 s-1
16.67 cm3 s-1
25 cm3 s-1
[DM
SO
]/[D
MS
O] in
itial
0.2
00 300 600 900
Time (s)
1200 1500
Figure 11 Concentration profiles of DMSO at different ozone flow
rates (Li et al. 2009c ).
intensity of the microbubble samples showed that the
ozone microbubbles produced the most hydroxyl radicals,
which contributed to the degradation of the organic materi-
als present in the coke wastewater. The removal efficiency
of pyridine in the ozone microbubble flotation process
was, respectively, 4.5 and 1.7 times higher, and the benzene
removal efficiency was 3.6 and 1.5 times higher, compared
with the air and oxygen microbubble flotation processes.
5.4 Disinfection
Ozone, by virtue of its strong oxidative power, is often
used in disinfection (Khadre et al. 2009 ). Ozone is effec-
tive in inactivating bacteria, viruses and certain algae.
The resistance of microorganisms follows the increasing
order: bacteria, viruses and cysts (Camel and Bermond
1998 ). Ozone microbubbles have been used for disinfec-
tion against fungi and bacteria. Kobayashi et al. (2011)
have investigated the disinfectant ability of ozone micro-
bubbles against Fusarium xysporum f. sp. melonis and
Pectobacterium carotovorum subsp . carotovorum in
infected plant roots. The microbubbles remained in the
water for a longer period than the millibubbles, resulting
in extremely high disinfecting activity against both phy-
topathogens. The disinfectant activity and durability of
the water treated with the ozone microbubbles increased
with an increase in the initial concentration of dissolved
ozone. Kobayashi et al. (2009) have also investigated the
ability of CO 2 microbubbles to inactivate Escherichia coli
suspended in a saline solution at a pressure lower than
2 MPa and 313 K. A significant reduction in the bacterial
population occurred with microbubble-aided CO 2 treat-
ment. The dissolved CO 2 concentration in the solution was
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208 S. Khuntia et al.: Microbubble-aided water purification
fairly high in the presence of the microbubbles. However,
bacteria could not be inactivated with nitrogen micro-
bubbles under the same conditions. The increase in the
CO 2 feed rate led to higher disinfection.
Tsuge et al. (2009) have studied the bactericidal
effect of ozone microbubbles on Bacillus subtilis . They
have studied the effects of parameters such as the outlet
pressure of pump, the concentration of ozone fed and
dissolved, and the microbubble formation methods. By
using ozone microbubbles, a high bactericidal effect was
obtained even with a small amount of ozone. Some of the
applications of microbubbles in water disinfection are
listed in Table 3 .
5.5 Impact of bromate formation during ozonation of water
During the disinfection of bromide-containing water (e.g.,
seawater) using ozone, a byproduct in the form of bromate
is formed, which is carcinogenic in nature (Gunten and
Hoign é 1994 ). The World Health Organization (WHO) has
recommended a permissible limit of 25 mg m −3 of bromate
in drinking water. Ozone readily oxidizes bromide ion
to bromate as per the following reactions (Gunten and
Hoign é 1994 ):
O 3 + Br − → O
2 + OBr − (22)
O 3 + OBr − → 2O
2 + Br − (23)
O 3 + OBr − → BrO−
2 + O
2 (24)
O 3 + HOBr → BrO−
2 + O
2 + H + (25)
BrO− 2 +O
3 → BrO−
3 (26)
Bromide ion directly reacts with molecular ozone
through a chain of first-order reactions with respect to
ozone concentration yielding hypobromous acid and
hypobromite (Gunten and Hoign é 1994 , Beltr á n 2004 ).
Hypobromite further reacts with ozone to form bromate.
Oxidation of hypobromous acid by ozone is very slow and
has a very negligible role in bromate formation. However,
the formation of bromate can be masked if ammonia is
present, as shown by the following reactions (Gunten and
Hoign é 1994 ):
HOBr → H + + OBr − (27)
HOBr + NH 3 → NH
2 Br + H
2 O (28)
HOBr + NH 2 Br → NHBr
2 + H
2 O (29)
HOBr + NHBr 2 → NBr
3 + H
2 O (30)
2H 2 O + NHBr
2 + NBr
3 → N
2 + 3Br − + 3H + + 2HOBr (31)
At the normal pH range of drinking water, the rate of
bromate formation is slow. Ammonia reacts with ozone in
the presence of bromide and forms monobromamine with
a faster reaction rate than that of hypobromite (Gunten and
Hoign é 1994 ). Ozone slowly reacts with monobromamine
to form nitrate and bromide. After most of the ammonia
is depleted, bromate is formed according to reactions
(22) – (26). The masking effect of ammonia on bromate form-
ation can be effectively used for ozonation of ammonia to
nitrate or nitrogen (Tanaka and Matsumura 2002 ).
6 Studies on microbubble-aided water purification
6.1 Use of air and oxygen microbubbles for water treatment
Increased transfer of oxygen to the liquid phase
can enhance the biodegradation capabilities of
Microbe type Microbubble type
Contact time (ks)
Temperature (K)
Disinfection ( % )
Other parameters References
Escherichia Coli CO 2 3.6 313 > 99 Pump outlet pressure = 0.5 – 2 MPa, CO
2 flow
rate = 1.7 – 33 cm 3 s−1
Kobayashi
et al. (2009)
E. Coli O 2 3.6 293 No effect Pump outlet pressure = 0.55 MPa Tsuge et al.
Bacillus subtilis O 3 3.6 293 > 99 Pump outlet pressure = 0.18 – 0.55 MPa,
Dissolved O 3 concentration = 0 – 0.5 g m −3
(2009)
Fusarium oxysporum f. sp. Melonis
O 3 1.2 288 – 303 > 99 [O
3 ] = 1.58 – 0.82 g m −3 , O
3 microbubble flow
rate = 41.7 cm 3 s−1
Kobayashi
et al. (2011)
Pectobacterium carotovorum subsp. carotovorum
O 3 1.8 288 – 303 > 99
Table 3 Use of microbubbles in water disinfection.
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S. Khuntia et al.: Microbubble-aided water purification 209
microorganisms. Gas-liquid dispersions, such as foams,
have been used to increase the biodegradation of hydro-
carbons (Ripley et al. 2002 ). Air microbubbles are often
used to supply oxygen in aerobic biodegradation pro-
cesses. Microbubbles can deliver oxygen to rather inac-
cessible regions and are more efficient than conventional
millibubbles (Kutty et al. 2010 ). Fresh microbubbles
replace the oxygen-depleted bubbles and biodegrada-
tion continues effectively. Choi et al. (2009) and Park
et al. (2009) have employed a saponin-based microbubble
suspension to enhance aerobic biodegradation of phen-
anthrene. As microbubble suspension flowed through the
sand column, oxygen was delivered to the less permeable
regions. Burkholderia cepacia RPH1, a phenanthrene-
degrading bacterium, was transported in a suspended
form in the microbubble suspension. The bacterial cells
partially attached to the gas-water interface. A significant
increase in biodegradation efficiency was obtained by
introducing the microbubbles.
Some studies have been reported in the literature on
the enhancement in aerobic biodegradation of phenol
(Michelsen et al. 1984 ), p -xylene (Jenkins et al. 1993 ),
penta chlorophenol (Mulligan and Eftekhari 2003 ) and
TCE (Rothmel et al. 1998 ) using microbubbles. Michelsen
et al. (1988) have developed a microbubble-based biode-
gradation system for treating hazardous wastes. They have
built a test cell for in situ biodegradation of flowing ground
water by the air microbubbles in 50 – 65 % dispersion. The
microbubbles were stabilized by surfactants. Microbubble
dispersion proved to be superior to the sparged air-water
injections. Approximately 25 % of oxygen present in the
microbubbles increased dissolved oxygen, as anaerobic
groundwater flowed through the treatment zone. Approxi-
mately one-third of this oxygen was required to biodegrade
the surfactant used to stabilize the microbubbles. The use
of surfactant in preparing and stabilizing the microbub-
bles has a side effect that the surfactant itself can pollute
water. The modern commercial microbubble generators,
e.g., the ASMK-III of the Riverforest Corporation, do not
require a surfactant for generating microbubbles. Some of
the applications of air and oxygen microbubbles in water
disinfection are listed in Table 4 .
SHARP Corporation (Japan) has developed a non-
dilution technology to remove 90 % of nitrogen contained
in wastewater discharged from its semiconductor facilities
(SHARP Corporation 2006). Their process involves the use
of MNBs of air (or oxygen-enriched air) to activate micro-
organisms. An improved microorganism culture increased
their concentration. The process developed by them is
schematically shown in Figure 12 . By using activated
microorganisms, they have treated nitrogen contained in
wastewater without dilution.
Compound Microbe type Microbubble type
Contact time (ks)
Removal ( % )
References
Biodegradation p -Xylene Pseudomonas putida O 2 2.7 99 Jenkins et al. (1993)
Trichloroethylene ENV 435 bacteria Air 86.4 95 – 99 Rothmel et al. (1998)
n -Hexadecane Acinetobacter junii O 2 1815 60 Ripley et al. (2002)
Pentachlorophenol Rhamnolipid Air – 84 Mulligan and Eftekhari (2003)
Phenanthrene Burkholderia cepacia RPH1 O 2 691 34 Choi et al. (2009)
Municipal water
treatment
TOC – Air 36 – Yamashita et al. (2010)
SS – Air 36 –
TN – Air 5.4 70
TP – Air 173 50
DOC – Air 43.2 59
Sulfate Desulfotomaculum nigrificans Air 43.2 80
Nitrification TN – Air 173 90 Winarto et al. (2010)
Coagulation-
flotation
COD – Air – 97 a Liu et al. (2010a,b)
Color – Air – 20 a
Nitrate – Air – 47 a
Ammonia – Air – 163 a
Table 4 Use of air/oxygen microbubbles in water treatment.
a Removal efficiencies (in % ) for the coagulation-microbubble-floatation process are higher than those of the coagulation-sedimentation
process.
COD, chemical oxygen demand; DOC, dissolved organic carbon; TN, total nitrogen; TOC, total organic carbon; TP, total phosphorous;
SS, suspended solid.
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210 S. Khuntia et al.: Microbubble-aided water purification
Tasaki et al. (2009a) studied the degradation of methyl
orange using oxygen microbubbles ( d = 5.8 μ m). They per-
formed photodegradation experiments with a BLB black
light blue lamp (365 nm), a UV-C germicidal lamp (254 nm)
and an ozone lamp (185 + 254 nm) both with and without
microbubbles. The oxygen micro bubbles accelerated the
decolorization rate of methyl orange under 185 + 254 nm
irradiation. By contrast, the microbubbles under 365 and
254 nm irradiation were ineffective for decolorization.
They found that the pseudo-zero order decolorization
reaction constant in the microbubble system was almost
twice that in the conventional millibubble system. Total
organic carbon (TOC) reduction rate of methyl orange was
greatly enhanced by microbubbles under 185 + 254 nm irra-
diation. However, TOC reduction rate by nitrogen micro-
bubbles was much slower than that with 185 + 254 nm irra-
diation only.
Takahashi et al. (2007a) have investigated the decom-
position of phenol in aqueous solution using air micro-
bubbles, without using any dynamic stimulus such as
UV irradiation or ultrasonic wave. They added a small
amount of nitric acid to generate the hydroxyl radical
from collapsing microbubbles. Under acidic conditions,
approximately 30 % of phenol was decomposed in 10.8 ks.
The intermediate products of phenol decomposition, that
is, hydroquinone, benzoquinone, and formic and oxalic
acids, were detected. They have reported that hydro-
chloric and sulfuric acids were also effective in generating
hydroxyl radicals.
Li et al. (2009b) have reported a strong effect of pH on
phenol degradation. Using air microbubbles, 59 % phenol
degraded after 7.2 ks at pH = 2.3, whereas only 3 % phenol
degraded at pH = 4.5, after the same amount of time. The
Before treatment
1. Microorganisms activated
2. Microorganisms concentrated
Optimize the qualityand conditions of
microorganism culture~5 g dm-3
≥10 g dm-3
MNB
MNB
After microorganisms areincreased and activated
Microorganisms
Ammonia, etc.
Figure 12 Activation and increase in microorganism population by using air micro-nanobubbles for treatment of wastewater
(adapted from SHARP Corporation Japan 2006 , with permission from the copyright holder, SHARP Corporation).
1.2
1.0
0.8
0.6
0.4
Phe
nol c
once
ntra
tion
(mol
m-3
)
0.2
0 0
5
10
15
TOC
removal efficiency (%
)
20
25
0 2000 4000Time (s)
6000
Phenol concentration profileTOC removal efficiency profile
8000
Figure 13 Variation of phenol concentration and TOC removal
efficiency with time (Li et al. 2009b ).
concentration profile of phenol and total organic carbon
removal efficiency are shown in Figure 13 . With oxygen
microbubbles, phenol degradation efficiency was much
higher and a similar effect of pH was also observed with
oxygen microbubbles. With nitrogen microbubbles,
degradation efficiency was 36 % only, at pH = 2.3.
Petroleum-based surfactants (e.g., alkylbenzene
sulfonates) are widely used as industrial detergents, emul-
sifiers and dispersing agents. Consequently, it is common to
find high concentrations of these surfactants in municipal
and industrial wastewaters, particularly from the washing
processes. Tasaki et al. (2009b) have employed a vacuum
UV method using MNBs (average diameter = 720 nm)
for the degradation of these surfactants. Sodium dodecyl
benzene sulfonates (SDBS) were used as the model
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S. Khuntia et al.: Microbubble-aided water purification 211
surfactant. Degradation experiments were conducted with
an ozone lamp, both with and without the MNBs. Their
results show that the oxidation rate of SDBS was signifi-
cantly enhanced by the oxygen MNBs. After 1 day of opera-
tion, 99.8 % of SDBS was oxidized and 76.8 % total organic
compound was removed in the integrated nanobubbles/
vacuum ultraviolet (VUV) system. The MNBs were found
to be more effective than the microbubbles (of 75.8 μ m
average diameter), due to the larger gas-liquid interfacial
area of the former system.
The effectiveness of air microbubbles in applications
such as decolorization can be enhanced by the use of UV
irradiation. The decolorization of methylene blue solution
has been studied by Shibata et al. (2011) . UV irradiation
improved decolorization at all pH values of the solution.
The effect of UV irradiation on decolorization was most
intensive for the neutral solution due to the enhanced
radical generation through photochemical reaction.
A high decomposition performance was obtained with the
venturi-type MBG in the presence of UV irradiation.
6.2 Removal of fine particles from wastewater
Wastewater often contains a significant amount of finely
dispersed solid particles. Removal of these particles
can be effectively done by microbubble flotation. The
microbubble-based flotation method has been practiced
in the processing of fine minerals (Yoon et al. 1992 , Yoon
1993 ). Air flotation is frequently used in water treatment
to remove algae from nutrient-rich water, or to treat water
of low turbidity at low temperature. Two types of air flota-
tion methods are used, that is, dissolved air flotation (DAF)
and induced air flotation (IAF). The DAF method is fairly
common for removing pollutants such as colloidal matters,
fine and ultrafine particles, precipitates, ions, microorgan-
isms, proteins, and dispersed and emulsified oils present
in water. Reviews on air flotation methods have been pre-
sented in the literature (Edzwald 1995 , Rubio et al. 2002 ,
Rodrigues and Rubio 2007 ). Details of commercially avail-
able microbubble-based flotation technologies for waste-
water treatment (e.g., those provided by Sionex, WesTech,
Aeromax Systems and Purac Engineering) have been pre-
sented by Rubio et al. (2002) .
In the DAF method, wastewater is introduced into the
water treatment unit. The fine particles in wastewater are
flocculated in the pretreatment process. There are two
steps in the pretreatment process, that is, particle desta-
bilization and particle flocculation. Particle destabili-
zation takes place by adding a chemical coagulant in a
flash mixer, where the coagulant is dispersed uniformly
and quickly by a high-speed stirrer. The particle floccu-
lation and growth of aggregates occur during the slow
mixing stage in the flocculator. A part of the purified water
is recycled and saturated up to 80 % with air at an elevated
pressure. The pressurized water is then decompressed via
injection nozzles in the flotation cell. Microbubbles, typi-
cally of the size range of 40 – 80 μ m, are formed in this
process. The bubbles readily adhere to the surface of the
floc particles forming bubble-particle agglomerates, and
rise to the liquid surface where a layer of sludge is formed.
The use of fine microbubbles improves the bubble-
particle collision efficiency, which is one of the most
important factors in the removal of particles. The amount
of air dissolved in water at a given pressure and tempera-
ture is rather low. The maximum value of air/water ratio is
approximately 1/17 to 1/13 (Li 2006 ).
The DAF method is more popular than the IAF method.
However, the DAF method has a few disadvantages, such
as high electrical power requirement, complexity in the
system and higher service cost. In the IAF method, rela-
tively large bubbles with diameters of several hundreds
of micrometers are formed by mechanical agitation or
sparger air injection. This method has been occasionally
used in wastewater treatment. For example, oils and fats
from dairy and abattoir wastes and blue-green algae from
natural wastewaters and maturation ponds have been
removed by the IAF method using a confined plunging jet
flotation cell (Yan and Jameson 2004 ). Large bubbles lead
to poor collision efficiency, and high shear in the conven-
tional mechanical flotation cells can lead to the breakage
of fragile particles, thereby generating finer particles. The
conventional technique is to add a surfactant to reduce
the bubble size. However, addition of a surfactant leads to
various pollution problems and an additional cost for the
treatment of sludge. Li (2006) has developed a separated
IAF system in which the bubble generation zone was sepa-
rated from the flotation. The microbubbles generated in
this IAF system had diameters in the range of 20 – 200 μ m,
and the average diameter was ∼ 70 μ m. Li studied the
removal of fine kaolin dispersions in water.
The surface charge of the bubbles and particles plays
an important role in separation. The ζ potential, therefore,
is an important parameter in DAF or IAF. Coagulants, e.g.,
alum [Al 2 (SO4)
3 ], are added to reduce electrostatic repul-
sion between the particles so that they can stick together
and create flocs. The microbubbles are negatively charged
(see Section 2.4) and thus also are the kaolin particles dis-
persed in water. The ζ potential of kaolin becomes less
negative with the addition of alum, approaches zero and
even becomes positive as the alum dosage is increased.
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212 S. Khuntia et al.: Microbubble-aided water purification
Therefore, the collision between the negatively charged
microbubble and the particle becomes more fruitful as the
alum dosage is increased. Consequently, good removal
efficiency (80 – 90 % ) can be achieved at an optimal alum
dosage. Too much addition of alum decreases the ζ poten-
tial of the floc particles (Han and Dockko 1999 ), which
reduces the particle removal efficiency.
Terasaka and Shinpo (2007) have developed a micro-
bubble-based floatation (MBF) system for removal of
carbon particles (mean diameter = 1 μ m) from wastewater.
The spiral liquid flow MBG (M2-LM type) was attached
at the bottom of the floatation column. The mean diam-
eter of the microbubbles was ∼ 63 μ m. Various cationic,
anionic and nonionic surfactants (i.e., cetyltrimethyl-
ammonium bromide, sodium dodecyl sulfate and Tween
20) were added, depending on charge requirements. The
carbon particles concentrated in the foam layer at the top
of the flotation column. The concentration of the carbon
particles in the foam reached a maximum value when the
concentration of surfactant was ∼ 5 % of its critical micelle
concentration (CMC). By contrast, the concentration of
the carbon particles in liquid phase decreased exponen-
tially within an hour of operation. They also proposed a
model to design the microbubble flotation column for the
removal of suspended fine particles.
Terasaka et al. (2008) have used MBF to recover the
iron oxide fine particles (mean diameter = 4.5 μ m) from a
suspension containing dilute surfactants (i.e., Tween 20,
sodium dodecyl sulfate and cetyltrimethylammonium
bromide). The suspended metal oxide particles were
hardly removed from the suspension using other aerators
such as a single orifice and a glass ball filter. However,
the MBF technique achieved 90 % recovery of the parti-
cles in just 60 min. The surfaces of the microbubbles and
the metal oxide particles were electrically charged. When
pH was controlled between 4.4 and 7.8, the surface of the
microbubbles was negatively charged and the surface of
the metal oxide particles was charged positively. There-
fore, the microbubbles and particles were attracted to
each other. At pH ∼ 5, the microbubbles were adsorbed
well on the metal oxide particles. In the foam layer at the
top of the flotation column, the particles were trapped
and then removed from the bulk liquid. The recoverabil-
ity depended on the degree of ionization and the concen-
tration of added surfactant. In their system, the suitable
concentration of the surfactant was only 1 – 2 % of the CMC
to achieve the highest recoverability of the suspended par-
ticles. To understand the mechanism of the separation of
the metal oxide particles using MBF, a kinetic model was
proposed, which takes into account the adsorption and
release of the particles.
The size of the microbubbles plays an important
role in MBF efficiency. Small microbubbles are required
for efficiently removing colloidal impurities from waste-
water. Adsorption on the surface of the microbubbles is an
important parameter (Yoshida et al. 2008 ). Cassell et al.
(1975) have investigated the effects of the concentration of
frother (e.g., ethanol) and collector (e.g., lauric acid) upon
bubble size, and upon the removal efficiency of humic
acid. Aluminum sulfate was employed as a coagulant.
All experiments were conducted at a pH of 7.5 in a 0.4 dm 3
batch flotation cell. Ethanol produced a greater effect on
bubble size than lauric acid. A small addition of ethanol
reduced the bubble size drastically. They have reported
that microflotation removal efficiency rapidly decreased
when the bubble diameter exceeded 55 – 60 μ m.
6.3 Removal of oil from wastewater
Produced water generated at the exploration and produc-
tion sites contains a large amount of oil. The oil compo-
nents may be either dissolved in the water or present as
a dispersed phase. The oil may have aliphatic, aromatic,
phenolic and fatty acid components. Several methods,
e.g., gravity separation (with or without corrugated-plate
interceptor), induced gas flotation (IGF), induced static
flotation (ISF), hydrocyclone and centrifuge, are used for
removing oil from wastewater. In the microbubble-based
flotation of oil from wastewater, the oil droplets adhere
to the surface of the bubbles, rise upward and collect at
the surface of water. On the surface, a frothy layer of oil
and gas is formed, which is skimmed off. Smaller micro-
bubbles are more effective in separating oil from water,
which results in a drier froth and a very low skim volume.
The water leaving the skim tank can have as little as
5 g m −3 oil (which typically correlates to a separation effi-
ciency > 95 % ). A large number of microbubbles per unit
volume of the dispersion create a higher probability for the
oil droplets to contact the bubbles. The microbubbles also
provide a large total surface area for attachment. The MBF
process reduces the use of flocculants. The electrostatic
charge on the bubble and oil surfaces plays an important
role in the efficiency of the removal of oil. Flotation of the
oil is promoted by decreasing the electrostatic repulsion
between oil flocs and air bubbles (Gray et al. 1997 ). Rubio
et al. (2002) have presented a review of the existing tech-
nologies for separating oil from wastewater.
Leung (1988) has developed a technique for the use of
microbubbles in the conventional ‘ hot water process ’ in
the primary flotation and settling step to recover bitumen.
In this process, streams of steam and air in admixture
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S. Khuntia et al.: Microbubble-aided water purification 213
were injected via a submerged nozzle into a flowing
aqueous stream. A fine dispersion of microbubbles (with a
dia meter < 100 μ m) was formed. The stream of microbub-
bles was injected into the aqueous slurry formed in the hot
water process. This injection was followed after the condi-
tioning step and prior to the introduction of the slurry into
the flotation/settling step.
The use of the DAF method in the removal of oil
from wastewater has been practiced over the past several
decades. Typically, the major portion of free and emulsi-
fied oil is removed by using the American Petroleum Insti-
tute (API) separator, followed by chemical clarifiers, DAF
or filters (Luthy et al. 1978 ). Thoma et al. (1999) have used
the DAF method to treat simulated produced waters that
contained paraffins (e.g., octane and decane) and aromatic
compounds (i.e., benzene, toluene and ethylbenzene). The
diameter of the microbubbles was in the range of 60 – 100
μ m. Using a gas flow rate of 1.7 cm 3 s−1, 95 % of the dissolved
octane was removed in 25 min and 75 % of micro-dispersed
decane was removed in 1.8 ks. For dissolved ethylbenzene
and toluene, 40 % removal was achieved in 21.6 ks at an air
deli very rate of 1 cm 3 s−1, and 70 % removal was achieved in
the same period with an air flow rate of 3.33 cm 3 s−1. They
have employed a solvent sublation technique in which
there is an immiscible layer of oil at the top of the water
column. The oil layer acts as a reservoir for the collection
of the organic contaminants. Various volatile and semi-
volatile materials dissolved in wastewater can be removed
by this method (Valsaraj et al. 1991 ).
Al -Shamrani et al. (2002) have used the DAF method
for separating oil from synthetic oily wastewater, which
was produced by emulsifying low concentrations of
oil in water with a nonionic surfactant (i.e., Span 20).
The operating parameters were the saturator pressure,
recycle ratio and air-to-oil ratio. They measured the
ζ potentials of the droplets of emulsion and flocculated oil.
The oil droplets were negatively charged between pH = 5
and pH = 10. Aluminum sulfate and four different cationic
polyelectrolytes were used to destabilize the system. The
ζ potential of flocculated oil droplets strongly depended
on the concentration of the polyelectrolyte. Although an
inversion of the charge of the oil droplets was observed
even at very low polyelectrolyte concentrations, the
polyelectrolytes were ineffective in enhancing the sepa-
ration. This effect was related to the structure of the poly-
electrolytes and to their adsorption mechanism. By con-
trast, for aluminum sulfate, it was important to decrease
the magnitude of the ζ potential to decrease electrostatic
repulsion so that the emulsion was destabilized prior to
flotation. A saturator efficiency of approximately 90 %
was achieved. Optimum conditions for separation were
obtained with an air-to-oil ratio of 0.0075, corresponding
to a recycle ratio of 10 % .
Gotoh et al. (2006) have investigated the removal of
oil from oil-polluted soil by air microbubbles. The separa-
tion of oil from high-concentration oil-in-water emulsion
was significantly enhanced by microbubble injection.
Approximately 70 – 80 % of the oil droplets were success-
fully removed by flotation using microbubbles. Oil flota-
tion was enhanced by physical adsorption characteristics
and very low slip velocity of microbubbles.
Deng et al. (2011) have developed a T-tube dynamic
state flotation device in which microbubbles were gener-
ated by using a porous polymeric membrane. The device
has several bubble entry points situated at the lower
part of the T-tube. Nitrogen microbubbles were produced
by applying high pressure. The bubbles, carrying the
oil droplets, flew upward through the T-tube, and then
passed through the upper exit. The pressure difference at
the equipment inlet controlled the elimination of the oil.
They have reported that this device can reduce oil concen-
tration from 38 to 12 g m −3 .
7 Conclusions and future prospects During the past few years, microbubbles have been com-
mercially used for various applications. This technology
has big prospects, although it is a rather new technology
for wastewater treatment. It is fairly certain that micro-
bubbles will be at the center of various technologies to be
used in the improvement of the aquatic environment. One
major challenge in microbubble technology is to reduce
power consumption and use low-power methodologies
without sacrificing desired microbubble properties. The
compression-decompression and acoustic cavitation
methods usually require high power consumption. New
methods, such as fluidic oscillation, are promising in
terms of lower power consumption.
Environmental hypoxia can be dramatically improved
by air and oxygen microbubbles to the semi-closed water
area. There is a great possibility of the use of nanobubbles
generated by the crushing of microbubbles (Takahashi
2005b ). Complete decomposition of an organic compound
may be effectively done by the nanobubbles consuming
a small amount of ozone. Moreover, the ozone nano-
bubbles are stable in water for a longer duration. In addi-
tion, microbubble technology can be used as an effective
disinfection technology for protection against infectious
diseases and sterilization of water. It can also be used
in curbing infectious diseases where an antibiotic is not
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214 S. Khuntia et al.: Microbubble-aided water purification
used. This can greatly contribute to the achievement of a
secure and safe society.
The use of hydroxyl radicals generated from collaps-
ing microbubbles is a novel method for the treatment of
wastewater. Only a few studies have been reported to date
regarding the treatment of wastewater using hydroxyl
radicals generated from microbubbles. The mechanism
of free radical generation is yet to be studied for ozone,
air, oxygen and nitrogen microbubbles. The use of ozone
in the water treatment process may yield some second-
ary and harmful compounds (Gunten and Hoign é 1994 ).
These surplus problems in water treatment can be solved
by the use of hydroxyl radicals generated from collaps-
ing microbubbles. Use of catalysts for the generation of
hydroxyl radicals from microbubbles can also be a tool for
research in the water treatment process.
The effect of the container can be important on the
efficiency of the microbubble-aided processes because the
microbubbles adhere to the surface of the container (Yang
et al. 2003 ). When the bubble is far away from the wall,
its transport in the bulk liquid is affected by the external
forces such as buoyancy and fluid drag. As the bubble
approaches the wall within a distance comparable to its
size, displacement of the fluid between the bubble and the
collector leads to additional hydrodynamic drag on the
bubble. At even closer distances ( < 100 nm), the bubble
motion is also affected by the surface forces. The material
of construction of the container (e.g., glass or plastic) can
significantly influence the bubble attachment. Further
studies are necessary for the design of the microbubble
reactor.
Residual ammonia in sewage from households is a
major source of pungent odor. Ammonia can be removed
from water by physical, biological and chemical methods.
Chemical oxidation is effective when the amount of
ammonia is low. The well-known oxidation by break-
point chlorination may result in the formation of toxic
chlorinated compounds in treated water. Microbubble-
aided ozonation can be a promising method for oxida-
tion of ammonia by ozone. It has been reported by Kuo
et al. (1997) that the reaction rate of ozone with ammonia
increases as the pH is increased from 8 to 10. At pH = 10,
addition of hydrogen peroxide (i.e., peroxone oxidation)
tremendously increases the rate constant. For waste-
waters containing high amounts of ammonia, the per-
oxone process can be effective and economical.
A promising use of microbubbles is in the treat-
ment of wastewater from seafood processing plants. This
wastewater contains blood, offal products, viscera, fins,
fish heads, shells, skins and meat fines. Fats, oil and
grease (FOG) and microorganisms are among the most
objectionable components from seafood processing waste-
water. Removal of these contaminants can significantly
reduce the adverse impact of seafood industries on adja-
cent water bodies such as rivers, lakes and oceans. Micro-
bubbles can effectively remove these substances from
wastewater. Studies reported in the literature suggest that
more than 80 % of FOG can be removed by microbubbles,
and separation efficiency can increase up to 95 % by the
use of additional coagulants/flocculants. Microbubble-
aided ozonation can be a very effective method for the
treatment of wastewater from seafood industries.
The disposal of untreated waste from wine industries
causes salination and eutrophication of water resources,
waterlogging and anaerobiosis, and loss of soil structure.
Winery wastewater is seasonally produced. It is generated
mainly as a result of cleaning in wineries, such as washing
operations during crushing and pressing of grapes,
rinsing of fermentations tanks, washing of barrels, and
bottling and purges from the cooling process. The volume
of wastes and the level of pollution greatly vary over
the year. Consequently, a reasonably versatile process is
required to face stream fluctuations. During peak seasons,
winery wastewater has a very high loading of solids and
soluble organic contaminants, but after this period con-
taminant load decreases substantially. As the wastewater
contains large amounts of ethanol and sugars, often a bio-
logical treatment process is used. A bioreactor using air
microbubbles may be suitable for treating winery waste-
water, and the treated water may be utilized in irrigation.
Microbubbles may be effective in oxidizing inorganic
matters present in wastewater (e.g., by ozonation). One of
the target applications is the removal of arsenic from water.
Arsenic is found at low concentration in natural water. The
maximum permissible concentration of arsenic in drink-
ing water has been set at 10 mg m −3 by the US Environmen-
tal Protection Agency (EPA) and the WHO. Considerable
amounts of arsenic are found in water and soil in many
countries (Mandal and Suzuki 2002 ). In water, the most
common valence states of arsenic are As (III) and As (V).
In the pH range of 4–10, predominant As (III) is neutral in
charge, whereas As (V) is negatively charged. As (V) is gen-
erally more efficiently removed than As (III) in commonly
practiced water treatment processes such as ion exchange,
iron coagulation followed by microfiltration, and activated
alumina adsorption. Hence, for drinking water supplies con-
taining significant concentrations of As (III), pre-oxidation
of As (III) to As (V) is mandatory for high arsenic removal.
Chlorine, ferric chloride, potassium perman ganate, ozone
and hydrogen peroxide can perform oxidation effectively.
Therefore, oxidation using ozone microbubbles can be an
effective route for converting As (III) to As (V).
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S. Khuntia et al.: Microbubble-aided water purification 215
Nomenclature a interfacial area per unit volume, m −1
A c cross-sectional area of the column, m 2
c concentration, mol m −3
c * saturation concentration of the gas in water, mol m −3
d diameter of microbubble, m
d 0 initial diameter of nanobubble, m
d 32
Sauter mean diameter, m
D diff usivity of gas in liquid, m 2 s −1
D b translational diff usion coeffi cient of microbubble, m 2 s −1
E electrophoretic mobility, m 2 V −1 s −1
g acceleration due to gravity, m s −2
H Henry ’ s law constant, Pa m 3 mol −1
k g gas phase mass transfer coeffi cient, mol N −1 s −1
K g overall gas-side mass transfer coeffi cient, mol N −1 s −1
k l liquid phase mass transfer coeffi cient, m s −1
K l overall liquid-side mass transfer coeffi cient, m s −1
n number of bubbles
NA Avogadro ’ s number, mol −1
p̃ partial pressure of gas, Pa
p g gas pressure, Pa
p l liquid pressure, Pa
Q g volumetric gas fl ow rate, m 3 s −1
R gas constant, J mol −1 K −1
t time, s
t d time taken for the bubble to travel one bubble diameter, s
T temperature, K
u rising velocity of bubble, m s −1
u̅ average rising velocity of bubbles, m s −1
u s superfi cial gas velocity, m s −1
V g volume of the gas phase, m3
V l volume of the liquid phase, m3
Greek letters γ surface tension of liquid, Nm−1
δ microbubble size dispersal coeffi cient
Δ p pressure diff erence between gas and liquid phases, Pa
Δ ρ density diff erence between the gas and liquid phases, kg m−3
ε dielectric constant of water
ε 0 permittivity of free space, C2 J −1 m −1
ε g fractional gas hold-up
μ iscosity of liquid, Pa s
ρ g density of gas, kg m−3
ρ l density of liquid, kg m−3
τ lifetime of a nanobubble, s
ζ zeta potential, V
Acknowledgments: The authors thank the Department
of Science and Technology (Water Technology Initiative),
Government of India, for financial support of this work,
through grant number: DST/TM/WTI/2k10/266/(G), dated
5 September, 2011.
Received June 6, 2012; accepted September 10, 2012
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From left to right: Pallab Ghosh, Snigdha Khuntia and Subrata
Kumar Majumder.
Dr. Pallab Ghosh is an Associate Professor in the Department
of Chemical Engineering at the Indian Institute of Technology
Guwahati, India. His research interests are interfacial
phenomena at gas-liquid and liquid-liquid interfaces,
nanofiltration and reverse osmosis membranes, gas-liquid and
liquid-liquid reactions, vapor-liquid and liquid-liquid equilibria
and design of chemical process networks using randomized
algorithms. He is a member of the American Chemical Society
(ACS), a Life Member of the Indian Institute of Chemical
Engineers (IIChE) and a Life Member of the Indian Society for
Technical Education (ISTE).
Ms. Snigdha Khuntia is a PhD student in the Department of
Chemical Engineering at the Indian Institute of Technology
Guwahati, India under the joint supervision of Dr. Pallab Ghosh
and Dr. Subrata Kumar Majumder. She is currently working on
wastewater treatment using ozone microbubbles.
Dr. Subrata Kumar Majumder is an Associate Professor in the
Department of Chemical Engineering at the Indian Institute of
Technology Guwahati, India. His research interests are multiphase
flow and reactor development, inverse fluidization, mineral
processing, process intensifications and microbubble technology.
He is a Life Member of the Indian Institute of Chemical Engineers
(IIChE) and the Indian Institute of Mineral Engineers (IIME).
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