11

Literature Review Chapter 2

2.1 Numerous studies on degradation of textile dyes and polycyclic aromatic

hydrocarbons with the use of ultrasound have been reported in the literature. However,

only the important ones, which had direct relevance to the present study, are being

reported here briefly;

Gayathri et al (2010) investigated [73] sonochemical degradation of two basic dyes

(Rhodamine B, Methylene Blue) and two acid dyes (Acid Orange II, Acid Scarlet Red

3R) in aqueous solution using sulphate radicals activated by immobilised cobalt ions. The

decolorisation efficiency for all four dyes solutions were in the order: persulphate

(PS) < cobalt activated persulphate (PS + Co) < persulphate + ultrasonication (PS + US)

< cobalt activated persulphate + ultrasonication (PS + US + Co).

Okitsu et al (2005) studied

[74] sonochemical degradation of two azo dyes

(C.I. Reactive Red 22, Methyl Orange). Ultrasound favoured the decolorization of azo

dyes while addition of t-butyl alcohol radical scavenger suppressed the sonochemical

decolorization. Attack of high concentration of OH radicals at the interface region of

cavitation bubbles facilitated the decomposition of azo dyes.

Petrier et al (2010) reported [75] sonochemical degradation of bisphenol–A through

bicarbonate ion in water which increased by a factor of 3.2 at low concentration of

bisphenol A (0.022 µ mol l-1

). This increase was due to the OH radicals which reacted

with bicarbonate ions to produce carbonate radical, which then migrated towards the bulk

of solution and helped in the degradation of bisphenol-A.

12

Rehorek et al (2004) used [76] ultrasound of 850 kHz at 60, 90 and 120 W for the

degradation of different industrial azo dyes (Acid Orange 5 and 52, Direct Blue 71,

Reactive Black 5 and Reactive Orange 16 and 107) to mineralize to non–toxic end

products. All investigated dyes had been decolorized and degraded within 3–15 hours at

90 W and within 1–4 hours at 120 W, respectively. They reported simultaneous azo bond

scission, oxidation of nitrogen atoms and hydroxylation of aromatic ring structures

through hydroxyl radicals attack on azo dyes. Degradation measured in terms of absolute

quantities (µmol/h) implied a linear correlation between the radical formation and the

amount of pollutants.

Wang et al (2008) investigated [77] sonochemical degradation rate of Acid Red B and

Rhodamine B and found that the degradation was much higher in the presence of

nanosized ZnO powders than with ultrasonic irradiation alone. They found that the rate

of degradation of Acid Red B was about two times higher than that of Rhodamine B at

the initial concentration of 10.0 mg/L, 1.0 g/L nano-sized ZnO powder, pH 7.0 and 60

min of irradiation and reported the difference of chemical forms of both dyes and surface

properties of ZnO powders. The kinetics of sonocatalytic reactions of Acid Red B and

Rhodamine B had pseudo first-order kinetics. The optimal conditions for high

degradation ratios of Acid Red B and Rhodamine B were found when 1.5 and 2.0 g/L

nanosized ZnO powders were added at pH 7.0 and 11.0, respectively, for the initial

concentration of 10 mg/L.

Priya et al (2006) studied [78] ultrasonic degradation of two dyes Rhodamine B and

Rhodamine Blue in the absence and presence of anatase TiO2 and Degussa P-25 TiO2.

The degradation rate was more in the presence of both catalysts but a higher rate of

degradation was found with anatase than Degussa P-25. The degradation rate increased

with decreasing pH, increasing temperature and higher intensity.

13

Wang et al (2008) reported [79] sonocatalytic degradation of Orange II, Ethyl Orange and

Acid Red G in the presence and absence of Au/TiO2. Discoloration and total organic

carbon removal of these dyes increased in the presence of catalyst compared to its

absence. Au/TiO2 led to the formation of active OH and H

radicals which enhanced both

oxidation and reduction mechanisms.

Vinu et al (2009) examined [80] sonophotocatalytic degradation and reduction of total

organic carbon of anionic dyes (Orange G, Methyl Blue, Indigo, Carmine) using

combustion synthesized TiO2 (CS TiO2) and Degussa P-25 TiO2. The degradation rate

and the reduction of total organic carbon with combined sonophotocatalytic method were

found higher compared to photocatalyst and ultrasound alone.

Stock et al (2000) evaluated [81] degradation of Napthol Blue Black azo dye and

observed an additive effect of sonolysis and photocatalysis process. Sonolysis enhanced

the degradation rate while photolysis promoted mineralisation of dyes.

Madhavan et al (2010) investigated [82] the degradation and mineralisation of Orange G

by ultrasound and in combination with TiO2. Low pH favoured sonolytic degradation of

Orange G while alkaline pH favoured photocatalytic degradation. Total organic carbon

(TOC) measurement suggested that the mineralisation efficiency was higher with

sonophotocatalyst compared to ultrasound and photocatalyst alone.

Perez et al (2008) compared [83] the degradation process of Malachite Green in the

presence of ultrasound with or without photocatalyst in CCl4. Ultrasound alone improved

the degradation rate in CCl4 and not with sonophotocatalyst.

Wang et al (2007) investigated [84] sonocatalytic degradation of Acid red B in the

presence of anatase- and rutile- TiO2 and found that the degradation rate of dye was faster

in the presence of anatase TiO2 powder compared to rutile TiO2.

14

Wang et al (2008) investigated [85] sonophotocatalytic degradation of Methyl Orange

using Degussa P-25, Yili TiO2 and Ag/TiO2. Ag/TiO2 was found to be the most effective

photocatalyst for dye degradation. In this investigation, mannitol and dimethyl sulfoxide

behaved as scavengers, and decreased the degradation rate.

Abdullah et al (2010) reported [86] sonocatalytic degradation of Congo Red,

Methyl Orange and Methylene Blue by TiO2 catalyst and H2O2. Ultrasound increased the

degradation rate. Small amount of rutile phase of photocatalyst showed better

sonocatalytic activity while excessive rutile phase showed poor activity. The highest

degradation rate of Congo red was due to multiple labile azo bonds which caused highest

reactivity with free radicals.

Abbasi et al (2008) performed [87] sonochemical degradation of Basic Blue 41 dye with

TiO2 as a catalyst in aqueous solution. Increase in H2O2 concentration and lowering the

initial dye concentration increased, but lower pH decreased the dye removal rate.

Ultrasound was found to be efficient for the degradation of azo dyes to non-toxic end

products.

Wang et al (2009) studied [88] sonocatalytic degradation of azo fuchsine in the presence

of cobalt and chromium-doped mixed crystal of TiO2. The sonocatalytic activity of

Cr-doped TiO2 powder was higher than that of Co-doped TiO2 and undoped TiO2.

Zhang et al (2009) studied [89] degradation of C.I. Acid orange 7 by ultrasound in

combination with Fenton process. Experimental results showed that increase of H2O2,

ultrasonic power and dissolved O2 increased the decolorization rate of dye. Increase in

initial pH decreased the decolorization rate. Decolorization rate varied a little at different

power or iron powder additions at the fixed H2O2 concentration. High COD removal

could be achieved by more H2O2 dosage and increased duration of sonic radiation.

15

Song et al (2009) investigated [90] the combination of ultrasound with fly ash/H2O2

(Fenton like reagent) for the degradation of C.I. Direct Black 168. Fly ash was the most

efficient heterogeneous catalyst. Concentration of the dye, H2O2, dosage of fly ash and

pH of solution affect the removal of dye considerably.

Ai et al (2010) studied [91] the degradation of several azo dyes such as

Rhodamine B, Methylene Blue, Reactive Brilliant Red X-3B and Methyl Orange by

ultrasound assisted electrocatalytic oxidation (US-EO) process and found effective in the

degradation. Ultrasound and electrocatalytic oxidation showed synergistic effect.

Wang et al (2010) studied [92] the degradation of Reactive dye and Brilliant Red X 3B by

coupling of electrolysis with ultrasound. The experiment was carried out in aqueous

solution with Ti-IrO2 as anode and graphite as cathode. The azo bond of the dye molecule

and the naphthalene ring were broken, facilitating the degradation of dyes, compared to

ultrasound or electrolysis alone. Degradation rate increased with increase in dye

concentration and addition of Cl- ion into the solution. Decolorization of reactive dye was

favoured by acidic condition.

Meshram et al (2010) investigated [93] the degradation of Congo red by ultrasound at

various parameters. It was found that the degradation of Congo red followed pseudo-first

order reaction kinetics and the degradation rate constant was found to be

1.13 x 10-3

min-1

for 20 mg/L dye at temperature 25º C and pH 5.9. The effects of

different parameters on the degradation were studied. It was found that with increasing

initial dye concentration and temperature, the rate of degradation was decreased. Effect of

pH indicated that both high acidic and high basic conditions were favourable to ultrasonic

degradation of dye. The effects of Fe2+

and Fenton reagent addition on the sonochemical

degradation were also investigated and the results indicated that the degradation rate of

Congo red was accelerated by Fe2+

, NaCl or Fenton reagent addition.

16

Tauber et al (2008) investigated [94] degradation of azo dyes by oxidative processes,

such as; Laccase and ultrasound treatments. Laccase treatment degraded Acid Orange and

Direct Blue dyes but none of the Reactive dyes. Ultrasound and Laccase showed

synergistic effect.

Laughrey et al (2001) studied [95] sonochemical degradation of aqueous polycyclic

aromatic hydrocarbon (PAH). Oxygen was an important precursor in degradation. It was

proved by increase in degradation rate constant when oxygen was bubbled into PAH

solution. Addition of organic compound decreased degradation rate due to scavenging of

oxygen.

Psillakis et al (2004) studied [96] the degradation of naphthalene, acenaphthylene and

phenanthrene. Degradation decreased with increasing initial concentration and

temperature, decreasing applied power, ultrasound frequency and in the presence of

excess dissolved salts. Addition of 1-butanol decreased the degradation due to scavenging

of hydroxyl radical while addition of Fe2+

ions increased degradation by Fenton like

reaction.

Kim et al (2003) studied [97] sonochemical decomposition of benzothiophene following

pseudo first order kinetics. The rate constant increased with increasing ultrasound energy

intensity and showed 77% decomposition of benzothiophene in presence of OH radicals.

Psillakis et al (2003) found [98] that the sonochemical treatment was capable of

destroying the lower molecular weight polycyclic aromatic hydrocarbons

(naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene,

fluoranthene and pyrene) completely but the higher molecular weight polycyclic aromatic

hydrocarbons, such as, benzo[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene,

chrysene, benzo[k]fluoranthene, indeno[1,2,3-cd]pyrene, dibenzo[ah]anthracene,

benzo[ghi]perylene were recalcitrant to ultrasound treatment.

17

2.2 Some important studies on degradation of textile dyes and polycyclic aromatic

hydrocarbons without the use of ultrasound have also been reported, as under;

Meric et al (2005) evaluated [99] the effectiveness of three processes, Fenton’s oxidation,

ozone and coagulation–flocculation to remove toxicity, colour and COD from textile

industry waste water. Daphnia magna was used to test acute toxicity. A colour range of

150-250 platin-cobalt (Pt-Co) units was assessed for toxicity. All three processes had

same rate for colour removal but COD removal rate was same only with Fenton and

coagulation-flocculation method.

Rauf et al (2009) discussed [100] decolorisation and degradation of dyes with high energy

radiation such as gamma radiation and pulsed electron beam. The transient species e-aq

found to be effective in decolorization but less effective in degradation. Degradation of

dye was initiated by attack of OH on electron rich site of dye molecule. The effect of

radiation dose, oxygen, pH, H2O2, added ions and types of dyes induced dye degradation.

Liu et al (2007) investigated [101] the decolorization of anthraquinone dyes by ozone.

The dyes were decolorized to 96% in 40 minutes converting complicated dye molecule

with the transformation of their –SO3H, Cl, nitrogen groups into SO4-2

, Cl-, NO3

-

respectively.

Barragan et al (2007) examined [102] the biodegradation of an azo dye (Acid orange 7)

by Enterobacter, Pseudomonas and Morganella species of bacteria inoculated on solid

media such as Kaolin, bentonite and powdered activated carbon (PAC). Solid media

selected must have certain special characteristics such as adequate particle size,

adsorption capacity, and surface texture. PAC with 0.490 nm particle size showed

favourable condition for bacterial degradation of dyes.

18

Swati et al (2012) proposed [103] the photocatalytic degradation of azo dye, Acid Red 73

in presence of Methylene Blue Immobilized Resin (MBIR) Dowex11 with solar light in

aqueous solution. The operational parameters such as dye concentration, catalyst loading,

pH and light intensity were investigated during the process. The results showed that the

optimum degradation larger than 92% of Acid Red 73 was observed in around 160 min at

room temperature at 7.5 pH and catalyst loading in around 2.0 gm.

Chatterjee et al (2010) reported [104] the reductive degradation of Reactive Black 5

(RB5) dye by zero-valent iron (ZVI) particles which were synthesized by the aqueous

phase borohydride reduction method. Three different surfactants, triton X-100 (TX100,

non-ionic surfactant), cetyl trimethyl ammonium bromide (CTAB, cationic surfactant),

and sodium dodecyl sulfate (SDS, anionic surfactant) were selected for the treatment of

ZVI in the decolourisation of RB5. The normalized residual concentration after

decolorization of 500 mg/L RB5 by ZVI for 3 h was 0.236, while ZVI particles treated

with TX100 (0.5 g/L), CTAB (1.0 g/L), and SDS (2.5 g/L) exhibited normalized residual

concentration of 0.172, 0.154, and 0.393, respectively, after 3 h. ZVI exhibited good

color removal efficiency at acidic pH. Decolorization kinetics by pseudo-first-order rate

equation showed that removal rate was increased after treatment with TX100 as well as

CTAB, while that was reduced after SDS treatment.

Rahmani et al (2010) reported [105] the removal of two azo dyes, Acid Orange 7 (AO7)

and Reactive Black 5 (RB5) by Fenton-like reaction. Removal of dye was increased by

increasing the iron mass and contact time. It was observed that high removal of dyes for

UV system was obtained at pH =11, while at pH=3 in the Feº and Feº/UV system.

Removal of AO7 in Feº/UV and Feº was increased by increasing the initial dye

concentration, while in the UV system it was decreased. Removal of RB 5 in Feº/UV and

Feº system was decreased by increasing the initial dye concentration while increased in

the UV system.

19

Abadulla et al (2000) reported [106] the degradation of triarylmethane, indigoid, azo, and

anthraquinonic dyes by Trametes hirsuta and a purified laccase. It was found that the

decolourisation velocities of dyes were dependent on the substituents present on the

phenolic rings of dyes. The laccase lost 50% of its activity at 50 mM NaCl while the 50%

inhibitory concentration (IC50) of the immobilized enzyme was 85 mM. Treatment of

dyes with the immobilized laccase reduced their toxicities (based on the oxygen

consumption rate of Pseudomonas putida) by up to 80% (anthraquinonic dyes).

Taheri et al (2008) employed [107] photocatalytic degradation of Acid Red 114 by

titanium dioxide nanoparticles. UV-Vis, Ion Chromatography (IC) and Chemical Oxygen

Demand (COD) analysis were employed to obtain the details of the photocatalytic

degradation of the AR 114. The effects of different experimental parameters such as dye

concentration, anions (NO3-, Cl

-, SO4

2-, HCO3

-) and pH were investigated.

Nikazar et al (2008) investigated [108] the photocatalytic degradation of azo dye Acid

Red 114 in water with TiO2 supported on clinoptilolite (CP) using solid-state dispersion

(SSD) method. The effects of experimental parameters such as pH, amount of

photocatalyst, and initial dye concentration were examined. The maximum effect of

photodegradation was observed at 10 wt % TiO2, 90 wt % clinoptilolite. A first order

reaction with k = 0.0127 min−1

was observed for the photocatalytic degradation reaction.

Chen (2009) carried [109] out the photocatalytic degradation of azo dye Reactive Orange

16 by TiO2. The effects of various parameters, such as photocatalyst amount, dye

concentration, light intensity, and temperature on photocatalytic degradation were

investigated. It was found that the decolourization efficiency was 87 % after 20 min

reaction and 100 % after 80 min reaction. The total mineralisation was 70 % after 20 min

and 100 % after 120 min, respectively. The results indicated that color degradation was

faster than the decrease of total organic carbon. The photocatalytic degradation process

was well described by first-order kinetics.

20

Plata et al (2008) studied [110] photochemical degradation of polycyclic aromatic

hydrocarbons. Disappearance rate of one set of polycyclic aromatic hydrocarbon isomer

benzo[a]pyrene and benzo[e]pyrene (kBAP/kBEP = 2.2 approximately) is different from

other set of isomer benzo[a]anthracene and chrysene (kBAA/kCHR = 2 approximately).

Inspite of their close structural similarity the difference in rate was presumed to be due to

differing capacity for direct photoreaction in oil film. However, the PAH

photodegradation was later discovered to be due to other compounds in the oil mixture.

Tran et al (2009) carried [111] out electrochemical degradation of polycyclic aromatic

hydrocarbons in a parallelepipedic electrolytic cell containing five anodes (Ti / RuO2)

and five cathodes (stainless steel). Current density of 9.23 mA cm-2

was beneficial for

polycyclic aromatic hydrocarbon (PAH) oxidation. (500-4000 mg) electrolyte Na2SO4

and initial PAH concentration had no effect on oxidation efficiency of hydrocarbon.

Alkaline media was not favourable for PAH oxidation.

Zhang et al (2008) investigated [112] photocatalytic degradation of phenanthrene, pyrene

and benzopyrene in presence of TiO2 using U.V. light source in a photochamber

maintaining the temperature of 30ºC. Degradation rate of these hydrocarbons increased

by TiO2 which reduced the half life of phenanthrene, pyrene, benzopyrene from 533.15 to

130.72 h, 630.09 to 192.53 h and 363.22 to 103.26 h respectively. Acidic or alkaline

condition favoured photocatalytic degradation than neutral condition. The degradation

was increased by combination of U.V. irradiation and TiO2 catalysis.

Bertrand et al (1990) investigated [113] biodegradation of hydrocarbon by an extremely

halophilic archaebacterium (strain EH4). Maximum growth of it on eicosane was

observed for 3.5 mol/L NaCl concentration while no growth was observed for less than

1.8 mol/L NaCl concentrations.

21

Lal et al (1996) tested [114] biodegradability crude oil from the different sources viz:

Bombay High and Gujarat crude oil using two bacteria strains, Acinetobacter

calcoaceticus (S30) and Alcaligenes odorans (P20). Oil samples from Bombay and

Gujarat were degraded to the extent of 50% and 29% respectively with S30 while 45%

and 32% respectively with P20 and in combination with both bacteria by 58% and 40%

respectively. S30 degraded more of alkanes fraction than aromatic fraction and could not

grow on pure polycyclic aromatic hydrocarbon compound except naphthalene. P20

degraded alkane and aromatic fraction equally and could grow on anthracene,

phenanthrene, dibenzothiophene, fluorene, fluoranthene, pyrene and chrysene.

Jonsson et al (2006) investigated [115] the capacity of Fenton’s reagent and ozone to

degrade polycyclic aromatic hydrocarbons (PAH) in soil. The degradation efficiency for

both methods was dependent on initial PAH concentration in soil. It was observed that

low molecular weight PAHs were more susceptible for degradation whereas high

molecular weight PAHs appeared to be strongly sorbed to soils and therefore less

chemically available for oxidation.

Tran et al (2009) investigated [116] the electro-oxidation of polynuclear aromatic

hydrocarbons (PAH) from creosote solution using Ti/IrO2 and Ti/SnO2 circular mesh

electrodes. Circular Ti/SnO2 electrode was found more effective in removing PAHs.

Degradation efficiency was greatly influenced by current density and retention time while

circulation flow rate and oxygen injection have very little influence.

Chung et al (2008) evaluated [117] the degradation efficiency of polycyclic aromatic

hydrocarbon in sewage sludge by electron beam irradiation and found to be effective in

removing PAHs. The degradation of PAHs was of first order with respect to adsorbed

dose.

22

Oleszczuk et al (2003) analysed [118] the degradation of polycyclic aromatic

hydrocarbons (PAH) such as phenanthrene, anthracene, fluoranthene, pyrene,

benzo[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[e]pyrene,

benzo[k]fluoranthene and benzo[a]pyrene in the surface layer (5 cm) of soil polluted with

aircraft fuel in an amount of 10 mL/kg was analysed. Group of microorganism occurring

naturally in the soil were able to degrade pollutants. Favourable conditions for the

development of microorganisms and the high content of nutrients played an important

role in the case of degradation of 3- rings PAHs, and the degradation of PAHs with a high

number of rings (> 4) was determined by their bioavailability.

Gonzalez et al (2010) reported [119] the biological degradation of anthracene. Earthworm

has been used to remediate anthracene-contaminated soil. About 41% removal of

anthracene was found for treatments without earthworms and 93% for those with

earthworms. Four degradation products were detected in soil of which

9,10-anthraquinone was most abundant. Thus the removal of anthracene and

9,10-anthraquinone was accelerated by E. Fetida.

Hammel et al (1995) reported [120] the degradation of polycyclic aromatic hydrocarbons

(PAHs) by ligninolytic fungi. Extracellular peroxidises by these fungi were responsible

for the initial oxidation of PAHs. Fungal lignin peroxidises oxidized certain PAHs

directly, whereas fungal manganese peroxidases cooxidise them indirectly during

enzyme-mediated lipid peroxidation.

Eibes et al (2006) reported [121] the enzymatic degradation of anthracene,

dibenzothiophene and pyrene by manganese peroxidase in media containing acetone.

These compounds were degraded to a large extent after a short period of time (7, 24 and

24 h, respectively). The order of degradability, in terms of degradation rates was as

follows: anthracene > dibenzothiophene > pyrene. Anthracene was degraded to phthalic

acid. A ring cleavage product of the oxidation of dibenzothiophene, 4-methoxy benzoic

acid, was also observed. Thus it was concluded that the addition of acetone increased the

solubility of PAHs and it did not hamper significantly the MnP activity.

23

Bisht et al (2010) reported [122] biodegradation of naphthalene and anthracene by

chemo-tactically active Rhizobacteria of populous deltoides. Kurthia sp and B.Circulans

showed positive chemotactic response for naphthalene and anthracene. B.Circulans SBA

12 and Kurthia SBA4 degraded 87.5% and 86.6% of anthracene while Kurtia sp. SBA 4,

B.Circulans SBA 12 and M.Varians SBA 8 degraded 85.3%, 95.8% and 86.8% of

naphthalene respectively after 6 days of incubation.

Kim et al (2001) studied [123] the effect of non-ionic surfactants on biodegradation of

polycyclic aromatic hydrocarbons (PAHs) in the aqueous phase and in the soil slurry.

Brij 30 surfactant was utilised for exsitu remediation, while Tween 80 and Triton X-100

for in-situ remediation because they have a higher dispersion in the aqueous phase. The

desorption experiments suggested that the surfactant concentration of 0.5, 1 or 2 g/L, with

a soil to water ratio of 1:10 (g/mL), would transfer substantial amount of the

phenanthrene from the soil to liquid phase. Brij 30 was the most biodegradable surfactant

tested, showed no substrate inhibition upto a concentration of 1.5 g/L. Naphthalene and

Phenanthrene were completely degraded by Phenanthrene- acclimatized cultures within

60 h.

Han et al (2004) investigated [124] the degradation of phenanthrene by white rot fungus

Trametes versicolor 951022 and its laccase. About 46% and 65% of 100 mg/L of

phenanthrene were removed, respectively after 36 h of incubation. Phenanthrene

degradation was maximum at pH 6 and temperature 30º C. The removal percentage of

phenanthrene was highest (76.7%) at 10 mg/L of phenanthrene concentration while the

transformation rate was maximal (0.82 mg/h) at 100 mg/L of phenanthrene concentration.

In the presence of mediator 2,2 azino-bis- (3-ethyl benzthiazoline-6-sulphonic acid)

(ABTS) or 1-hydroxy benzotriazole (HIBT) in the oxidation of phenanthrene was

increased to 40% and 30%, respectively by laccase.

24

Eibes et al (2005) reported [125] the degradation of anthracene by ligninolytic enzyme

manganese peroxidase in organic solvent mixtures (acetone, methylethylketone, methanol

and ethanol). Due to the maximal solubilisation of anthracene and minimum loss of MnP

activity, acetone (36%) was found to be important (it enhanced 143- fold the anthracene

solubility). Complete degradation of anthracene was attained after 6 h of operation under

optimal conditions.

Simarro et al (2011) determined [126] the optimum values for the biodegradation process

of PAH (naphthalene, phenanthrene and anthracene) by a bacterial consortium C2PLO5.

The most effective values of each optimised factors were: a C/N/P molar ratio of

100: 21:16, NaNO3 as nitrogen source, Fe2(SO4)3 as iron source using a concentration of

0.1 m mol L-1

, a pH of 7 and a mixture of glucose and PAHs as carbon source. It was

concluded that the high concentration of nutrients, soluble form of nitrogen and iron and

addition of glucose to PAHs as carbon source enhanced PAH biodegradation due to the

augmentation of PAH degrader microorganisms.

Thongkred et al (2011) reported [127] the oxidation of polycyclic aromatic hydrocarbons

by a tropical isolate of Pycnoporus Coccineus and its Laccase. The highest laccase yield

(5.97± 0.40 U/mL) was obtained when grown at 28ºC for 8 days in basal medium

containing 2% (w/v) glucose and 0.25% (w/v) peptone at an initial pH of 5. The

oxidation of anthracene, pyrene and fluoranthene were 59.7, 50.7 and 49.8%, when added

to liquid culture by P. Coccineus at 100 ppm within 24 h, whereas benzo[a]pyrene and

phenanthrene were oxidised to only 25.3% and 32.4 %, respectively. When P. Coccineus

crude laccase (1U/mL) was used, pyrene and anthracene were oxidised to 76.4% and

74.3% respectively within 2 h and fluoranthene to 84.2% within 24 hr. Benzo[a]pyrene

and phenanthrene oxidation by the enzyme were enhanced when 1 mM 2,2’- azino bis

(3-ethyl benzothiazoline- 6-sulphonic acid) was added, resulting in 66.3% and 50.5%

oxidation after 24 h at initial 100 ppm.

25

Jinquan et al (2006) investigated [128] the effects of chlorine dioxide (ClO2) on the

degradation of anthracene (ANTH), pyrene (PYR) and benzo[a]anthracene (BaA). The

maximum degradation ratio of ANTH, PYR and BaA was found to be 99.0%, 67.5% and

89.8% respectively, when reaction time was 30, 60 and 120 min, the concentration of

ClO2 was 0.1, 0.4 and 0.5 m mol L-1

, and pH was 7.2. The oxidation product formed in

the reaction between ANTH and ClO2 was 9,10 anthraquinone and the mechanism of

single electron transfer (SET) was proposed as the possible pathway for ANTH-ClO2

reaction.

Zaidi et al (1999) studied [129] the microbial degradation of polycyclic aromatic

hydrocarbon phenanthrene in the Caribbean coastal water. Addition of KNO3 as a source

of inorganic nitrogen (N) resulted in a 10-fold increase in the rate of phenanthrene

degradation within 125 h period, whereas, addition of K2HPO4 as a source of inorganic

nutrient phosphorus (P) had no effect. It was observed that phenanthrene in seawater

samples degraded rapidly when first pretreated with hydrogen peroxide (H2O2) and then

inoculated with a known indigenous phenanthrene degrading bacterium, Alteromonas sp.

2.3 Various studies on adsorption of textile dyes and polycyclic aromatic hydrocarbons

with the use of ultrasound have been reported briefly, as under;

Li et al (2009) reported [130] the removal of C.I. Direct Blue 78 dye by the combination

of ultrasound and exfoliated graphite. It was found that Direct Blue 78 was decolourised

nearly 100% by the combined ultrasound/exfoliated graphite at pH 1, dosage of

exfoliated graphite 1 g/L, contact time 20 min and 45οC. The maximum adsorption

capacity of exfoliated graphite towards the dye was found to be 152.9 mg/g. Equilibrium

data were fitted to Langmuir and Freundlich isotherm and were best described by

Freundlich isotherm model. The adsorption followed first order kinetics.

26

Entezari et al (2007) carried [131] out the removal of Reactive Black 5 (RB5) from

aqueous solutions by the sorption process in the presence and absence of ultrasound.

Sorption of the dye was investigated to determine the influence of different parameters

such as contact time, amount of sorbent and concentration of pollutant on the removal

efficiency of RB5 with and without ultrasound. The experimental data were fitted

properly to the Freundlich model and the isotherm constants were 28.2 and 7.4 for kf and

0.13 and 0.38 for 1/n in the presence and in the absence of ultrasound (20 kHz)

respectively. The data were analyzed with different sorption kinetic models and were

better fitted with a pseudo-second-order kinetic model. Two ultrasonic generators at 20

and 500 kHz were used for sonication and it was found that the rate of removal was

higher at the high frequency than at the lower one.

Sayan et al (2008) investigated [132] the decolorization of reactive dye C.I. Reactive

Blue 19 from aqueous solution using ultrasound, activated carbon and combined

ultrasound/activated carbon. The decolorization of RB 19 by using ultrasound, activated

carbon and combined ultrasound/activated carbon was found to be 36%, 91% and 99.9%,

respectively. The decolourisation was modelled statistically and optimised by means of

the Matlab computer software.

Entezari et al (2006) examined [133] the sorption of methylene blue as a basic dye onto

cellulosic materials such as waste newspaper in the presence of ultrasound

(sono-sorption) and in its absence (conventional method). The results showed that in case

of conventional method, dye removal was increased accordingly by increasing the

amount of adsorbent while it was stopped at specific amount of adsorbent in case of

sono-sorption. More than 98% removal of the dye could be achieved in a very short

period of time of sonication with respect to the conventional method.

27

2.4 Few interesting studies on adsorption of textile dyes and polycyclic aromatic

hydrocarbons without the use of ultrasound have been reported, as under;

Abdelwahab et al (2005) investigated [134] the use of rice husk, an agricultural waste for

adsorption of synthetic Direct F. Scarlet (Direct Red 23). The results indicated that the

ability of the rice husk in removing the Direct Red 23 (DR-23) were dependent on the

dye and rice husk concentrations. Maximum dye colour removal was observed after 72

hours for all the system conditions. The adsorption isotherm of DR-23 was described by

the Langmuir isotherm model. Kinetics of adsorption followed Lagergren first order

kinetic model with film diffusion being the rate- controlling step. Activated rice husk was

observed to have higher adsorption capacity than the untreated rice husk in the removal

of DR-23. The maximum adsorption capacity was ~13 mg of dye per one gram of dry

rice husk. This study showed that the rice husk could be employed as low-cost and

effective sorbent for the removal of Direct F. Scarlet from aqueous solution.

Namasivayam et al (2002) reported [135] the adsorption of Congo Red by coir pith

carbon by varying the parameters such as agitation time, dye concentration, adsorbent

dose, pH and temperature. Equilibrium adsorption data followed both Langmuir and

Freundlich isotherms. Adsorption followed second-order rate kinetics. The adsorption

capacity was found to be 6.7 mg dye per g of the adsorbent. Acidic pH was favourable

for the adsorption of Congo Red. Very low desorption of dye suggested that

chemisorption might be the major mode of dye removal by the adsorbent.

Bozlur et al (2010) investigated [136] the use of acid activated saw dust in removing

Lurazol Brown pH (LBP) dye as a function of agitation time, adsorbent dosage and initial

dye concentration. The adsorption obeyed both Langmuir and Freundlich isotherm and

the values of their constants indicated favourable and beneficial adsorption. It was found

that only 2.5 gm of acid activated sawdust was capable of removing 94% of dye from an

initial concentration of 300 mg/L of 40 mL of solution. A two-stage filtration system has

been designed, constructed and demonstrated.

28

Shabudeen et al (2006) reported [137] the use of activated carbon which was prepared

from waste kapok hull in removing Rhodamine dye-B. The factors affecting the rate of

adsorption such as initial dye concentration, agitation time, carbon dose, particle size and

pH variation have been studied at various temperature ranges at 300, 318 and 330K. The

experimental data was analyzed for possible agreement with the Lagergren, Langmuir

and Freundlich adsorption isotherm equations. The intraparticle diffusion rate constant,

adsorption rate constants, diffusion rate constants and diffusion coefficients were

determined. The adsorption rate was mainly controlled by intraparticle diffusion. The

structural and morphological of activated carbon were characterized by XRD and SEM

studies respectively.

Singh et al (2011) investigated [138] application of saw dust (raw and modified) for the

removal of Orange G from its aqueous solutions. The raw saw dust was modified by

sulphuric acid and sodium bicarbonate. Equilibrium data were fitted to Langmuir and

Freundlich isotherm and were best described by Freundlich isotherm model. The

maximum adsorption capacity was found to be 0.40 mg g-1

for an initial concentration of

2.5 mg L-1

of Orange G at a dose of 1.0 g L-1

of the adsorbent. A maximum removal

(44%) of Orange G was found at 298 K. The negative values of ΔG were indicative of the

spontaneous nature of the process of dye removal, whereas the negative values of ΔH and

ΔS indicated the exothermic nature of adsorption and decrease in degree of freedom of

the adsorbed species, respectively.

Zawani et al (2009) investigated [139] the removal of Remazol Black 5 from the

synthetic wastewater using palm kernel shell activated carbon in terms of initial pH,

initial concentration, contact time and temperature. The optimum pH was found at acidic

range, pH 2. Equilibrium data were fitted to Langmuir and Freundlich isotherm and were

best described by Freundlich isotherm model. The maximum adsorption capacity at 30ºC,

40ºC and 50ºC were found to be 58.8. 96.7 and 98.6 mg/g, respectively. The pseudo

first-order kinetic model fitted very well with the adsorption behaviour of RB5 dye. The

negative value of ΔGº and ΔHº indicated that the adsorption was a spontaneous and

exothermic process.

29

Sharma et al (2009) investigated [140] the application of activated carbon developed

from rice husk, an agricultural waste product for the removal of Malachite Green from

aqueous solutions and wastewaters. The removal increased from 93.75 to 94.91% by

decreasing the initial concentration from 100 to 60 mg/L. Time of equilibrium was found

to be 40 min. Equilibrium data were fitted to Freundlich and Langmuir isotherm

equations and the isotherm constants were determined. Kinetics of adsorption followed

first order kinetic model. The monolayer adsorption capacity of rice husk activated

carbon for adsorption of the dye was found to be 63.85 mg/g at room temperature.

Nourouzi et al (2009) studied [141] the adsorption of two reactive dyes, Reactive Black 5

and Reactive Red E onto palm kernel shell-based activated carbon. The experiment was

carried out to investigate three models: film diffusion model, film surface and film-pore

diffusion models. The results showed that the external coefficients of mass transfer

decreased with increasing of initial adsorbate concentration. Film surface diffusion model

was able to fit experimental better than film pore diffusion model.

Jayarajan et al (2011) discovered [142] the effect of agricultural waste product

nano-porous adsorbent of Jack fruit peel waste for removing Rhodamine dye (RD) from

aqueous solution. The effects of adsorption isotherm were studied at different adsorbent

dosages, temperature and pH. Adsorption behaviour of the dye could be described well

by Langmuir and Freundlich models. The monolayer adsorption capacity determined was

determined to be 4.361 to 1.98 mg/g.

Ofomaja et al (2008) investigated [143] the biosorption of Methyl Violet dye from

aqueous solution using Mansonia wood sawdust. The equilibrium biosorption data were

analyzed using three isotherm models; Langmuir, Freundlich and Redlich-Peterson

isotherm. Best fits were yielded with Langmuir and Redlich-Peterson isotherms. The

removal percentage of dye became significant above pH 7, which was slightly higher

than the pHPZC of the sawdust material. The thermodynamic parameters, such as

∆G°, ∆H°, and ∆S° suggested that the biosorption was a spontaneous and endothermic

process.

30

Annadurai et al (2002) investigated [144] the adsorption of Methyl Orange (MO),

Methylene Blue (MB), Rhodamine B (RB), Congo Red (CR), Methyl Violet (MV) and

Amido Black 10B (AB) by low-cost banana and orange peels as adsorbents. The

adsorption capacities for both peels decreased in the order: MO > MB > RB > CR > MV

> AB. The Freundlich equation showed a better fit than does the Langmuir equation for

adsorption of dyes using banana peel, but exactly reversed using orange peels. An

alkaline pH was favorable for the adsorption of dyes. Based on the adsorption capacity, it

was shown that banana peel was more effective than orange peel. Kinetic parameters of

adsorption such as the Lagergren rate constant and the intraparticle diffusion rate constant

were determined. For the present adsorption process intraparticle diffusion of dyes within

the particle was identified to be rate limiting.

Bhattacharyya et al (2003) reported [145] the adsorption of Brilliant Green dye on

powdered dry leaves of the Neem tree. The adsorption process was carried out in a batch

process with different concentrations of the dye solution (2.07 x 10-2

to 10.36 x 10-2

m mol dm-3

), different adsorbent doses (0.13-0.63 g dm-3

), and temperature (300, 303,

313, and 323K). The removal of dye was nearly 25% with 0.13 g dm-3

of adsorbent after

1 hr of contact time which increases to 71.8% after 5 hr and further increment upto 98%

was achieved by an increase in adsorbent dose to 0.63 g dm-3

. The suitability of the

adsorbent was tested by fitting the adsorption data with Langmuir and Freundlich

isotherms and by computing equilibrium thermodynamic and kinetic parameters.

Suteu et al (2005) reported [146] the sorption of reactive dye Brilliant Red HE-3B from

aqueous solutions on commercially powdered activated charcoal as a function of solution

pH, initial dye concentration, temperature and contact time. The equilibrium sorption data

fitted well to Langmuir model. The apparent thermodynamic parameters of sorption

suggested an entropy-driven, endothermic sorption process. The kinetic of the sorption

fitted well to the pseudo-second order kinetic model, indicated an intraparticle diffusion

mechanism.

31

Sarioglu et al (2006) reported [147] the mechanism of Methylene Blue adsorption on

biosolid (waste sludge) through batch experiments. The effects of various experimental

parameters, such as pH (3 - 11), bio solid dosage (1-10 g L-1

), contact time (5-1440 min)

and initial dye concentration were investigated. The results showed that the dye removal

increased with increase in the initial concentration of the dye and also increased with

amount of bio solid used and initial pH. Adsorption data was modelled using the

Freundlich adsorption isotherm. The results show that bio solid could be employed as an

effective and low cost material for removal of dyes and colour from aqueous solution.

Filipkowska et al (2002) investigated [148] the adsorption of reactive dyes onto chitin

from aqueous solutions at pH 3. Ten reactive dyes were examined, including 5 dyes from

the helactine group, 3 dyes from the polactine group and 2 dyes from the remazol group.

The K and b constants were calculated from the Langmuir equation that assumes the

presence of two sites of different nature. It was found that the dye adsorption on chitin in

type I and II sites differed in both the adsorption affinity and maximum adsorption

capacity. Based on the dimensionless separation factor RL it was found that the dye

adsorption mechanism in type I sites was an ion exchange in the following decreasing

order: helactine dyes > polactine dyes > remazol dyes, whereas in the case of type II sites

it was a physical adsorption in the following increasing order: helactine dyes < polactine

dyes < remazol dyes. The experiments indicated a correlation between dye adsorption on

chitin and a number of sulphonic groups in a dye molecule while correlation with the

number of aromatic rings, benzene or naphthalene was not found.

Maximova et al (2008) investigated [149] the adsorption of two basic dyes, Basic Red 18

and Basic Blue 5, onto two fractions, Perfil M-100 and Perfil M-150. The equilibrium

sorption isotherms were correlated well by the Langmuir and Redlich-Peterson isotherms

for Basic Blue 5 and by the Freundlich model for Basic Red 18. The adsorption capacity

of two sorbents for Basic Red 18 followed the order: perfil M-100 > perfil M-150 and for

Basic Blue 5 it was perfil M-150 > Perfil M-100. The kinetics models of sorption were

analysed using pseudo- first order and pseudo- second order kinetic models and were

successfully fitted by pseudo-second order model.

32

Chiou et al (2006) reported [150] the adsorption behaviour of acid dye

(MY, metanil yellow) and reactive dye (RB15, reactive blue 15) in aqueous solutions by

the cross-linked chitosan beads. The adsorption capacities was 3.56 m mol g-1

(1334 mg g-1

) for dye MY and 0.56 m mol g-1

(722 mg g-1

) for dye RB15 at pH 4, 30 °C.

The Langmuir model agreed very well with the experimental data (R2 > 0.996). The first-

order kinetic model fitted well for lower initial dye concentrations, while the second-

order kinetic model fits well for higher initial dye concentrations. It was found that high

adsorption capacity of anionic dyes onto chemically cross-linked chitosan beads was due

to strong electrostatic interaction between the -NH3+ of chitosan and the dye anions. The

competitive adsorption favoured the dye RB15 in the mixture solution

(initial conc. (mM): MY = 1.34; RB15 = 1.36); while it favoured the dye MY in the

mixture solution (initial conc. (mM): MY = 3.00; RB15 = 1.34) and the adsorption

kinetics for dye RB15 has the tendency to shift to a slower first order model.

Nasuha et al (2011) reported [151] the adsorption of cationic, Methylene blue (MB) and

anionic, Congo Red (CR) dye from aqueous solution by using chemically modified

papaya seed (PS). The results showed that the chemical modification esterification of

carboxyl group could increase the adsorption capacity of papaya seed in removal of

cationic dye compared to anionic dye. The effect of different experimental parameters

such as initial dye concentration (50-500 mg/L) and various adsorbents (NPS and EPS)

were investigated. The Langmuir isotherm showed a better fit than does the Freundlich

for adsorption of MB, but for CR the best fit isotherm model was Freundlich. Maximum

monolayer adsorption capacity was found to be 250.00 mg/g and 200.00 mg/g at 30°C for

EPS and for NPS respectively. The adsorption kinetics obeyed pseudo-second order with

higher coefficient of correlation R2 0.99.

Vacha et al (2006) investigated [152] the adsorption of benzene, naphthalene, anthracene,

and phenanthrene by means of molecular dynamics simulations. Potentials of mean

force, i.e., free energy profiles were evaluated when moving the studied molecules across

an aqueous slab. In all cases, deep surface free energy minima, corresponding to orders of

magnitude of surface enhancement of the aromatic molecule, were located. It was found

33

that the hydration free energies computed from MD calculations and the experimentally

determined values were in good agreement. Data pertaining to the importance of the

air–water interface in the adsorption and transport of PAHs on micron sized water

droplets were described. The relevant data on adsorption and reaction (ozonation and

photochemical) at the air–water interface of planar surfaces and droplets were also

summarized.

Saba et al (2011) examined [153] the adsorption kinetics of two polycyclic aromatic

hydrocarbons anthracene and phenanthrene on soil from different areas of Attock

Refinery Limited (ARL), Rawalpindi, Pakistan. It was found that the high concentration,

acidic pH and high organic matter contributed to higher adsorption capacities. The

maximum adsorption exhibited for anthracene and phenanthrene were 95.2% and 95.4%,

respectively under optimum condition of pH and initial concentration. The experimental

results were analysed by Freundlich and Langmuir isotherms.

Gok et al (2008) investigated [154] the adsorption kinetics of naphthalene onto

organo-sepiolite from aqueous solutions. Natural sepiolite was modified by a surfactant,

which was dodecyltrimethyl ammonium (DTMA) bromide. The optimum pH values and

the equilibrium contact time for the adsorption of naphthalene onto DTMA-sepiolite were

found as 6 and 75 min, respectively. The adsorption capacity of DTMA sepiolite

increases with increasing temperature indicated that the adsorption process was

endothermic in nature. The maximum adsorption capacity of DTMA-sepiolite for

naphthalene was found to be 2.867 mg/g.

Huang et al (2010) investigated [155] the potential of Gomphidius viscidus, a kind of

ectomycorrhizal fungi, for phytoremediation of anthracene in soil. Adsorption capacity of

1886.79 mg/g and 1515.15 mg/g at 25º C, pH 6.0 for active and inactive mycelia

respectively, was obtained. It was found that the process of equilibrium could be reached

within 40-60 min and described well by the Langmuir isotherm model and pseudo-second

order sorption kinetics. The adsorption capacity increased with an increase in initial

anthracene concentration while decreased with rise in temperature at 25ºC and 35ºC.

34

Eiceman et al (1983) examined [156] the adsorption of six polycyclic aromatic

hydrocarbons on fly ash. Direct measurement of adsorption using elution analysis with

gas-solid chromatography has been made on fly ash from a municipal incinerator and a

coal fired power plant. Results showed that the irreversible adsorption occurred at

concentration below 30 µg PAH per gram of flyash, while at concentration above 30 µg/g

adsorption of PAH were governed by vapour pressure.

Duc et al (2010) examined [157] the adsorption of polycyclic aromatic hydrocarbon on

graphite surfaces. Coronene (C24H12) was chosen as a model compound to study the

interaction between a typical PAH and a graphite surface. They investigated the problem

by adopting an applied mathematical modelling approach and therefore exploit the

continuous atomistic approximation together with the Lennard Jones potential instead of

using conventional computational methods, such as full ab initio, which was not feasible

owing to the large molecules involved. They described the mechanism of adsorption of a

Coronene molecule on a graphite surface using an analytical expression for the

interaction energy.

Yuan et al (2010) studied [158] the adsorption of polycyclic aromatic hydrocarbons:

naphthalene, fluorene, phenanthrene, pyrene and fluoranthene from aqueous solution

using petroleum coke-derived porous carbon. The specific surface area (SSA) of the

carbons ranges from 562 to 1904 m2/g, while their point of zero charges pHpzc varies

from 2.6 to 8.8. The adsorption capacity parameter Kf for any given PAH, increase with

the SSA and pHpzc of the carbons, confirming the roles of dispersive interactions. The

value of Kf followed the order: naphthalene > fluorene > phenanthrene > pyrene. Results

indicated that the uptake process was likely to be controlled by diffusive transport

processes.

35

Walters et al (1984) examined [159] the adsorption of polycyclic aromatic hydrocarbons

onto activated carbon. An evaluation of Henry’s law, Langmuir, BET, Freundlich and

Redlich-Peterson equations indicated that the Langmuir equation was most useful for

representing the data. The Henry’s Law adsorption constants ranged from

2390 (mg/g)/(mg/L) for naphthalene to 326000 (mg/g)/(mg/L) for chrysene. The

adsorption capacity obtained from Langmuir equation ranged from 580 mg/g for

naphthalene to 14.7 mg/g for benzo[a]anthracene. Results indicated that the adsorption of

PAH onto activated carbon was much stronger than was adsorption of PAH onto soils,

sediments and suspended organic matter.

Hall et al (2009) investigated [160] the adsorption of polycyclic aromatic hydrocarbons

onto silica gel. Acenaphthene was selected as target compound. Kinetic results were

analysed using pseudo-first and pseudo-second order equations, and equilibrium data

were described by the Langmuir and Freundlich isotherms. It was concluded that the

lower molecular mass PAHs, such as acenaphthene, were well represented by the pseudo-

first order equation. Adsorption mechanisms were analysed using boundary layer mass

transfer and intra particle diffusion models, indicating intraparticle diffusion as the main

adsorption controlling step.

Low et al (1988) examined [161] the adsorption of anthracene, pyrene and

dibenzo[a,c]anthracene on nine fly ash samples from the combustion of Australian

bituminous and brown coals by a high-performance liquid chromatographic procedure.

Results indicated that the adsorption data was well represented by the Freundlich

equation. From the correlation analysis between the adsorption capacity of PAHs and

physical and chemical characteristics of fly ashes, it was indicated that the residual

carbon content was the main regulating parameter.

36

Crisafully et al (2008) investigated [162] the removal of polycyclic aromatic

hydrocarbons (PAHs) from petrochemical wastewater using various low-cost adsorbents

of natural origin including sugar cane bagasse, green coconut shells, chitin, and chitosan.

Adsorption experiments of mixtures of PAHs (5.0-15.0 mg/L) have been carried out at

ambient temperature (28+/-2 degrees C) and pH 7.5. The adsorption isotherms of PAHs

were in agreement with a Freundlich model, while the uptake capacity of PAHs followed

the order: green coconut shells > sugar cane bagasse > chitin > chitosan. The adsorption

properties of green coconut shells were comparable to those of some conventional

adsorbents such as Amberlite T.

Chen et al (2011) investigated [163] the removal of polycyclic aromatic hydrocarbons

onto wood chips (WC), ryegrass (RR), orange peels (OP), bamboo leaves (BL) and pine

needles (PN). Biosorption isotherms fitted well with Freundlich equation and the

mechanism was dominated by partition process. The magnitude of phenanthrene partition

coefficients (Kd) followed the order of PN > BL > OP > RR > WC, ranged from

2484 ± 24.24 to 5306 ± 92.49 L/kg. Except the WC sample, the Kd values were

negatively correlated with sugar content, polar index [(N + O)/C] of the biosorbents,

while the aromatic component exhibited positive effects.

Ma et al (2011) determined [164] the cell wall-cosolvent partition coefficients (Km) of

polycyclic aromatic hydrocarbons for Rhizopus oryzae cell walls by controlling the

volume fraction of methanol (f) ranging from 0.1 to 0.5. Five cosolvent models were

employed for extrapolating the cell wall-water partition coefficients (Kw) in pure water.

The extrapolated Kw values of PAHs on R. oryzae cell walls were ranged from 2.9 to 5.1.

Comparison of various Kw values of pyrene generated from extrapolation and the QSPR

model, together with predicted different (PD), mean percentage deviations (MPD), and

root mean square errors (RSE), revealed that the performance of the LL and Bayesian

models were the best among all five tested cosolvent models. This study suggested that R.

oryzae cell walls played an important role in the partitioning of PAHs during

bioremediation because of the high Kw of fungal cell walls.

37

Long et al (2008) investigated [165] the adsorption of naphthalene on a spherical

microporous carbon adsorbent (CR-1), which had been obtained by carbonization and

activation of the waste polysulfonated cation exchange resin. . The adsorption data fitted

best with Polanyi-Dubinin-Manes model, although tested simultaneously with

Freundlich, Langmuir, Brunauer-Emmett-Teller models as well. Through both isotherm

modeling and constructing "characteristic curve", Polanyi theory was useful to describe

the adsorption process of naphthalene by CR-1, providing evidence that a micropore

filling phenomenon was involved. Adsorption process was found to follow pseudo-first-

order kinetic equation.

Chang et al (2004) examined [166] the adsorption of naphthalene on zeolite from

water-butanol solution, which was a surfactant-enriched scrubbing liquid. The adsorption

data was successfully evaluated by Langmuir, Freundlich, and linear isotherms. The

adsorption kinetics fitted best with the pseudo-second order equation, although tested

simultaneously with pseudo-first order and Elovich rate equation.

Sener et al (2010) investigated [167] the adsorption behaviour of naphthalene onto

sonicate talc from aqueous solution. The naphthalene uptake of talc was found as

276 mg/g and increased to 359 mg/g after the sonication. Adsorption studies also showed

that the adsorption of naphthalene onto the sonicated talc was not affected by changes in

pH suggesting that the main driving forces for naphthalene adsorption onto talc was

hydrophobic bonding rather than electrostatic force. The Freundlich isotherm best fitted

for the adsorption of naphthalene onto talc and the adsorption process followed the

pseudo-second-order rate expression for different initial naphthalene concentrations.

38

Long et al (2009) investigated [168] the adsorption of naphthalene onto two adsorbents, a

hypercrosslinked polymeric adsorbent (NDA-150) and a macroporous polymeric

adsorbent (XAD-4), which have the same matrix of poly (styrene-divinylbenzene) and

similar surface area. The equilibrium data were correlated with Polanyi–Manes–Dubinin

equation, and the adsorption enthalpy changes (ΔHmeas) were determined. The good fits

of Polanyi–Manes–Dubinin equation and high-exothermic adsorption enthalpies

indicated that micropore filling was involved during the adsorption of naphthalene onto

two adsorbents. The higher adsorption capacity of hypercrosslinked adsorbent compared

to macroporous adsorbent was due to its larger micropore volume. The adsorption of

naphthalene was found to follow a pseudo-second-order process which was controlled by

intra-particle diffusion. The adsorption rate of macroporous adsorbent (XAD-4) was

faster than that of hypercrosslinked adsorbent (NDA-150) as it has more suitable porous

structure for transporting the adsorbate into the adsorption sites.

Demirkan et al (2011) investigated [169] the adsorption of two nonpolar petroleum

contaminants, naphthalene, and o-xylene, onto seven fly ashes with varying carbon

contents, with powdered activated carbon (PAC). Six equilibrium isotherm models were

used to evaluate the batch data. The naphthalene and o-xylene adsorption capacity of the

fly ashes was correlated with the unburnt carbon content, specific surface area of the

sorbent, and the percentage of the anisotropic and isotropic carbon content of the ash. A

pore-filling mechanism was the dominant mechanism for the adsorption of nonpolar

organic chemicals onto PAC as evaluated on the basis of the Polanyi-Dubinin-Manes

model, whereas the adsorption onto fly ash was likely to be governed by the unburned

carbon content and the specific surface area of the ash.

Manop et al (2009) investigated [170] the adsorption of anthracene (C14H10) and pyrene

(C16H10) on two types of soot, rubber-wood furnaces soot (RB) and diesel engine soot

(DE). Characteristics of the soot, such as carbon contents, specific surface area and pore

volume were investigated. Adsorption isotherms models, i.e., Freundlich (Fre) and

Polanyi-Dubinin-Manes (PDM) were fitted well. KF in Fre and VO in PDM illustrated

adsorption degree of the soot in which KF of anthracene on RB and on DE were 0.97 and

39

0.74 and VO of anthracene was 200.30 and187.78 cm3 mol

-1, respectively. On the other

hand, KF of pyrene on RB and on DE were 0.97 and 0.74 and VO of pyrene was 50.07

and 44.66 cm3 mol

-1, respectively. Results indicated the amount of adsorption of

anthracene was higher than that of pyrene and the adsorption capacities of RB was larger

than DE. The energy of adsorption E as calculated from PDM model for anthracene on

RB and DE was found to be -8.84 and -8.62 kJ mol-1

and those of pyrene on RB and DE

were -7.39 and -7.45 kJ mol-1

, respectively.

Owabor et al (2010) investigated [171] the adsorption and desorption kinetics of

naphthalene, anthracene and pyrene in a soil slurry reactor to ascertain the mechanism

controlling the retention and release rates of these compounds. A stirred flow method was

employed to perform the experiments. The extent of partitioning for the polycyclic

aromatic hydrocarbons (PAHs) tested was found to be dependent on their solubility and

diffusivity in the aqueous phase. Apparent adsorption and desorption rate coefficients

were determined using the Langmuir and Freundlich isotherm models. Equilibrium

adsorption and desorption at the external surface and in the internal pore of the soil

particle obeyed the Freundlich isotherm equation. The pseudo-equilibrium condition

established at the minimum contact time suggests that equilibrium adsorption attained for

the contaminant PAHs was not instantaneous but rather time dependent.

Dyachuk et al (2009) examined [172] the adsorption of PAHs by solid phase

luminescence from micro heterogeneous media containing an anionic surfactant

(sodium dodecyl sulphate, SDS) and a nonionogenic polymer (polyethylene glycol, PEG

1000), on cellulose and polyurethane foams (PUF). Thus, the adsorption of PAHs by the

modified cellulose considerably decreased the detection limit of PAHs in aqueous media.

The highest sensitivity of the determination was attained in the phosphorescence method.

The best results were obtained in the determination of PAHs using the preconcentration

in the microvolume of SDS micelles modified with polyethylene glycols followed by the

solid-phase extraction of PAHs and the analysis in the adsorbent phase.