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Microbubble-aided water and wastewater purification: A review

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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: skmaju@iitg.ernet.in; rsmaju@yahoo.com

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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