DEVELOPMENT OF METHODS FOR REMEDIATION OF

POLLUTANTS IN WATER BY USING ECO-FRIENDLY

MATERIALS TRANSFORMED FROM NATURAL AGRO

WASTES

GAN PEI PEI

(B.Sc. (Hons.), NUS)

A THESIS SUBMITTED

FOR THE DEGREE OF DOCTOR OF PHILOSOPHY

DEPARTMENT OF CHEMISTRY

NATIONAL UNIVERSITY OF SINGAPORE

2013

Declaration Page

I

Declaration Page

I hereby declare that this thesis is my original work and it has been written by

me in its entirety, under the supervision of Prof Sam, Li Fong Yau, (in the

laboratory S5-02/04/05/06/07), Chemistry Department, National University of

Singapore, between 12 Jan 2009 and 01 Nov 2012.

I have duly acknowledged all the sources of information which have been used

in the thesis.

This thesis has also not been submitted for any degree in any university

previously.

The content of the thesis has been partly published in:

1) P.P. Gan, S.H. Ng, Y. Huang, S.F.Y. Li, Green synthesis of gold

nanoparticles using palm oil mill effluent (POME): A low-cost and

eco-friendly viable approach, Bioresour. Technol., 113 (2012) 132-135.

2) P.P. Gan, S.F.Y. Li, Potential of plant as a biological factory to

synthesize gold and silver nanoparticles and their applications, Rev.

Environ. Sci. Biotecnol, 11 (2012) 169-206.

3) P.P. Gan, S.F.Y. Li, Biosorption of elements, Element Recovery and

Sustainability, RSC Green Chemistry, Chapter 4, publication in

progress.

4) P.P. Gan, S.F.Y. Li, Efficient removal of Rhodamine B using a rice

hull-based silica supported iron catalyst by Fenton-like process, Chem.

Eng. J., accepted.

Declaration Page

II

Gan Pei Pei

10 June 2013

Name Signature Date

Acknowledgements

III

Acknowledgements

I would like to express my sincere gratitude to my supervisor, Professor Sam

Li, who has provided me his guidance throughout my graduate studies. I

appreciate very much for the opportunity and flexibility given by him which

allowed me to carry out the research work on my interested topics.

I would like to acknowledge the financial support provided by the National

University of Singapore, National Research Foundation and Economic

Development Board (SPORE, COY-15-EWI-RCFSA/N197-1) and the

Ministry of Education (R-143-000-441-112) which provide me the opportunity

to work in a good environment with necessary resources and facilities for my

research.

I would also like to extend my appreciation to all the lab members and staffs

in Prof Sam Li’s group who encourage, support, and offer me their help during

the course of my study. The assistance and training provided by all the

CMMAC staff in both department of Chemistry and Engineering are also

appreciated.

Last but not least, I would like to thank all my family and friends for their

unconditional support and patience that given during the pursuit of my study.

Table of Contents

IV

Table of Contents

Declaration Page .............................................................................................. I

Acknowledgements ....................................................................................... III

Table of Contents .......................................................................................... IV

Summary ........................................................................................................ XI

List of Tables .............................................................................................. XIV

List of Figures ............................................................................................. XVI

List of Abbreviations ................................................................................. XXI

List of Symbols ........................................................................................ XXIII

Chapter 1 Introduction .............................................................................. 1

1.1 Crisis of Heavy Metal Pollution and Resource Scarcity ............ 1

1.1.1 Conventional Methods for Remediation of Heavy Metal

Pollution ..................................................................................... 3

1.1.2 Biosorption as a Non-destructive Remediation Tool for Heavy

Metal Pollution ........................................................................... 4

1.1.2.1 Basic Principles of Biosorption ............................................ 6

1.1.2.2 Plausible Mechanisms of Biosorption .................................. 7

1.1.3 Palm Oil Mill Effluent (POME) ............................................... 10

1.1.3.1 Potential of POME for Biosorption of heavy metals .......... 10

1.1.4 Orthogonal Array Design as a Chemometric Method for

Optimization of Biomass Pretreatment .................................... 11

1.1.5 Visual MINTEQ for the Calculations of Metal Speciation ...... 12

1.2 Crisis of Organic Dye Pollution ............................................... 12

Table of Contents

V

1.2.1 Conventional Methods for Remediation of Organic Dye

Pollution ................................................................................... 13

1.2.2 Advanced Oxidation Processes as a Destructive Remediation

Tool for Organic Dye Pollution ............................................... 14

1.2.2.1 Fenton and Fenton-like Treatment ..................................... 15

1.2.2.2 Homogeneous vs. Heterogeneous Fenton Treatment ......... 16

1.2.3 Silica as Catalyst Support for Heterogeneous Fenton Treatment

.................................................................................................. 17

1.2.3.1 Potential of Rice Hull Waste as Natural Feedstock of Silica

........................................................................................... 18

1.2.3.2 Enhancement of catalytic performance of Fenton catalyst by

incorporation of AuNps ..................................................... 19

1.2.4 Physical and Chemical Syntheses of Nanoparticles ................. 21

1.2.5 Biosynthesis of Nanoparticles .................................................. 22

1.2.5.1 Basic Principles of Biosynthesis......................................... 22

1.2.5.2 Potential of POME for Biosynthesis of AuNps .................. 25

1.3 Research Scope ........................................................................ 26

Chapter 2 Biosorption of Cd(II) and Hg(II) from Aqueous Solutions

using Palm Oil Mill Effluent (POME) as a Low-cost

Biosorbent ............................................................................... 30

2.1 Introduction .............................................................................. 30

2.2 Material and Methods ............................................................... 32

2.2.1 Solutions and Reagents ............................................................ 32

2.2.2 Preparation of Biomass Material .............................................. 33

Table of Contents

VI

2.2.3 Characterization of the POME Biosorbent ............................... 33

2.2.4 Optimization Strategy for Pretreatment of POME ................... 33

2.2.5 Batch Adsorption Experiments ................................................ 35

2.2.5.1 General Procedure for Batch Adsorption ........................... 35

2.2.5.2 Biosorption Kinetics ........................................................... 35

2.2.5.3 Biosorption Isotherms......................................................... 37

2.2.5.3.1 Langmuir Isotherm ......................................................... 37

2.2.5.3.2 Freundlich Isotherm ....................................................... 37

2.2.5.3.3 Dubinin-Radushkevich Model ....................................... 38

2.2.5.4 Thermodynamic study ........................................................ 39

2.2.5.5 Desorption and reusability studies ...................................... 39

2.3 Results and Discussions ........................................................... 40

2.3.1 Characterization of POME biosorbent ..................................... 40

2.3.2 Optimization of Pretreatment Conditions ................................ 42

2.3.3 Study of Biosorption Kinetics .................................................. 45

2.3.4 Study of Biosorption Isotherms ................................................ 48

2.3.5 Effect of Ionic Strength ............................................................ 53

2.3.6 Effect of Adsorbent dosage ...................................................... 54

2.3.7 Thermodynamic Study ............................................................. 55

2.3.8 Desorption and Reusability Studies ......................................... 57

2.4 Concluding Remarks ................................................................ 60

Chapter 3 Efficient Removal of Rhodamine B using a Rice hull-based

silica supported iron catalyst by Fenton-like process ......... 61

3.1 Introduction .............................................................................. 61

Table of Contents

VII

3.2 Materials and Methods ............................................................. 64

3.2.1 Solutions and Reagents ............................................................ 64

3.2.2 Extraction of Sodium Silicate Solution from Raw Rice Hulls . 64

3.2.3 Synthesis of RHSi-Fe Catalyst ................................................. 65

3.2.4 Characterization of RHSi-Fe Catalyst ...................................... 65

3.2.5 Heterogeneous Fenton-like degradation of RhB ...................... 66

3.2.5.1 Chemical Oxygen Demand (COD) Measurements ............ 67

3.2.5.2 Characterization of Degraded Products by FTIR Analysis 67

3.2.6 Sonochemical and Photocatalytic Experiments ....................... 67

3.2.7 Stepwise Addition Strategy Study ............................................ 68

3.3 Results and Discussions ........................................................... 68

3.3.1 Characterization of RHSi-Fe Catalyst ...................................... 68

3.3.2 Degradation Characteristic of RhB using RHSi-Fe ................. 69

3.3.3 Effect of Radical Scavengers ................................................... 74

3.3.4 Effect of RhB concentration ..................................................... 76

3.3.5 Effect of Oxidant Concentration .............................................. 78

3.3.6 Effect of Catalyst Dosage ......................................................... 80

3.3.7 Effect of temperature ................................................................ 81

3.3.8 Effect of pH .............................................................................. 82

3.3.9 Effect of Mass Transfer Resistance .......................................... 86

3.3.10 Effect of Foreign Salts and Ionic Strength ............................... 87

3.3.11 Effect of Stepwise Addition Strategy ....................................... 90

3.3.12 Comparative Study with Ultrasound and Ultraviolet Irradiations

.................................................................................................. 93

Table of Contents

VIII

3.3.13 Stability and Reusability of the Catalyst .................................. 95

3.4 Concluding Remarks ................................................................ 98

Chapter 4 Green Synthesis of Gold Nanoparticles using Palm Oil Mill

Effluent (POME): A Low-cost and Eco-friendly viable

Approach ................................................................................. 99

4.1 Introduction .............................................................................. 99

4.2 Material and Methods ............................................................. 101

4.2.1 Solutions and Reagents .......................................................... 101

4.2.2 Preparation of POME ............................................................. 101

4.2.3 Synthesis of AuNps ................................................................ 101

4.2.4 Characterization of AuNps ..................................................... 102

4.2.5 FTIR Analysis ........................................................................ 102

4.2.6 Synthesis of AuNps using POME Extracts ............................ 103

4.3 Results and Discussions ......................................................... 103

4.3.1 Effects of Initial pH on the Biosynthesis of AuNps ............... 103

4.3.2 Effect of the of HAuCl4 Concentration on the Biosynthesis of

AuNps ..................................................................................... 106

4.3.3 Effect of Temperature on the Biosynthesis of AuNps ........... 107

4.3.4 Characterization of AuNps ..................................................... 108

4.3.4.1 TEM Analysis ................................................................... 108

4.3.4.2 Particle Size Distribution .................................................. 111

4.3.4.3 X-ray diffraction Measurement ........................................ 112

4.3.4.4 Stability of Nanoparticles ................................................. 113

Table of Contents

IX

4.3.5 Possible Functional Groups involved in Biosynthesis Process

................................................................................................ 114

4.3.6 Synthesis of AuNps using Different POME Extracts ............ 115

4.3.7 Interaction of Biosynthesized AuNps with Mercury Ions ...... 118

4.4 Concluding Remarks .............................................................. 119

Chapter 5 Enhancement of Catalytic Performance of a Rice hull-based

Silica Supported Iron Catalyst by Biosynthesized AuNps for

Fenton-like Degradation of Rhodamine B in water .......... 120

5.1 Introduction ............................................................................ 120

5.2 Material and Methods ............................................................. 122

5.2.1 Solutions and Reagents .......................................................... 122

5.2.2 Synthesis of Silica-coated AuNps (Au@SiO2) ...................... 123

5.2.3 Synthesis of gold-silica-iron composite material (Au@RHSi-Fe)

................................................................................................ 124

5.2.4 Characterization of catalyst .................................................... 124

5.2.5 Degradation Experiments ....................................................... 125

5.3 Results and discussions .......................................................... 126

5.3.1 Characterization of Catalyst ................................................... 126

5.3.2 Catalytic Performance of Au@RHSi-Fe under UV-irradiation

................................................................................................ 130

5.3.3 Catalytic Performance of Au@RHSi-Fe at Room Temperature

................................................................................................ 132

5.3.4 Catalytic Performance of Au@RHSi-Fe at different pH ....... 133

5.3.5 Possible Mechanisms ............................................................. 135

Table of Contents

X

5.4 Concluding Remarks .............................................................. 138

Chapter 6 Conclusion and Future Work.............................................. 139

6.1 Summary of Results ............................................................... 139

6.2 Current Challenges and Directions for Future Work ............. 142

References ..................................................................................................... 145

Publications and Manuscripts in Preparation .......................................... 171

Appendix 1 Figure S1 Schematic diagram of oil extraction from oil palm

and POME generation ............................................................ 173

Appendix 2 Table S1.1 Structures of various dye classes ........................ 174

Summary

XI

Summary

This dissertation examines and explores the possibility of using eco-friendly

materials transformed from agro wastes for remediation of persistent

pollutants in water. The motivation of our work is to provide a more

affordable and eco-friendly approach that can serve as a pretreatment step to

enhance the treatability of those persistent pollutants in water before they are

being discharged to water body or sent to the municipal sewage treatment

plant.

Palm oil mill effluent (POME) is a residue that is discharged from palm oil

mill in large amount during the processing of oil palm for crude palm oil

production. Instead of being discharged as waste, their rich content of

polymeric functional groups could actually be utilized for other useful

purposes. In this dissertation, the potential of POME for biosorption of toxic

heavy metals have been demonstrated in Chapter 2. It was found thatPOME

modified by NaOH under optimum conditions is able to achieve rapid removal

of Cd(II) and Hg(II) from aqueous solutions at an optimum pH of 4.5.

In addition to POME, another agro waste being studied in this dissertation is

rice hull waste. Rice hull which comprises about 20% of silica appears to be

an attractive natural source of silica. Instead of being discharged as waste or

burnt openly after the rice milling process, the content of silica could be

extracted from rice hull and used for the preparation of silica-based materials.

In Chapter 3, a rice hull-based silica supported iron catalyst (RHSi-Fe) was

prepared and its catalytic activity towards heterogeneous Fenton-like

degradation of a xanthenes dye, Rhodamine B (RhB) was evaluated. The

Summary

XII

results show that this catalyst is able to work with a wide range of dye

concentrations with fast degradation rate (in 10 min) achieved at pH 3.0. In

addition, the degree of mineralization could be further enhanced by using

stepwise addition strategy, in which a more efficient utilization of oxidant

could be achieved without the requirement of additional reagent.

Although RHSi-Fe has been proven to be effective in catalyzing the

degradation of RhB in Chapter 3, the catalytic performance of this catalyst

could still be further enhanced through some modifications of the catalyst

components. One feasible approach is by incorporating gold nanoparticles

(AuNps) in the catalyst structure. Since reutilization of natural agro wastes is

one of the objectives of this research work, the potential of POME in

biosynthesis of AuNps was explored in Chapter 4 and it was found that POME

is useful for the biosynthesis of AuNps at room temperature. The

biosynthesized AuNps were predominantly spherical with an average size of

18.75 nm.

In Chapter 5, the POME-assisted biosynthesized AuNps were incorporated in

RHSi-Fe with the purpose of enhancing its catalytic activity towards

degradation of RhB. The enhancement of catalytic activity after incorporation

of AuNps was observed in both UV-assisted and non-irradiation-assisted

processes. The role of AuNps was proposed to be mainly an electron

scavenger which could consume the electrons released by the silica support

upon the exposure to UV light. The improved performance observed for non-

irradiation-assisted process could be attributed to the ability of AuNps to act as

an electron relay between the oxidation and the reduction semi-reactions.

Summary

XIII

Lastly, the research findings were concluded with a discussion on the current

challenges and suggested future work in Chapter 6.

List of Tables

XIV

List of Tables

Table 1.1 The maximum allowable concentrations of metals in trade

effluent ....................................................................................... 2

Table 1.2 The advantages and disadvantages of conventional treatment

methods as compared to biosorption .......................................... 5

Table 2.1 Assignment of factors and level settings for POME modified

with NaOH by using OA9 (34) followed by OA8 (2

7) matrix

along with the results of output responses. (initial Cd(II)

concentration: 50 mgL-1

, initial Hg(II) concentration: 15 mgL-1

,

POME dose: 3.75 gL-1

, contact time: 2 h, adsorption pH: 4.5)

.................................................................................................. 34

Table 2.2a ANOVA table for OA8 (27) matrix with Cd removal as the

output responses ...................................................................... 43

Table 2.2b ANOVA table for OA8 (27) matrix with Hg removal as the

output responses ...................................................................... 44

Table 2.3 Kinetic parameters for the adsorption of Cd(II) and Hg(II) ions

on the NaOH modified-POME biosorbent (MPOME dose: 3.75

gL-1

, temperature: 298 K, pH: 4.5, agitation speed: 300 rpm,

initial Cd(II) concentration: 50 mgL-1

, Hg(II) concentration: 50

mgL-1

) ....................................................................................... 46

Table 2.4 Parameters for Langmuir, Freundlich, and D-R isotherms at

different temperature (MPOME dose: 3.75 gL-1

, temperature:

298 K, pH: 4.5, agitation speed: 300 rpm, contact time: 1 h for

Cd(II), 2 h for Hg(II)) ............................................................... 50

Table 2.5 Thermodynamic parameters for the adsorption of Cd(II) and

Hg(II) by MPOME (MPOME dose: 3.75 gL-1

, pH: 4.5,

agitation speed: 300 rpm, contact time: 1 h for Cd(II), 2 h for

Hg(II)) ...................................................................................... 56

Table 2.6 Performance of various desorption solutions on the recovery of

metal ions and the effect of EDDS on the regeneration of

biosorbent during repeated adsorption/desorption cycles (initial

Cd(II) concentration: 50 mgL-1

, initial Hg(II) concentration: 20

mgL-1

) ....................................................................................... 59

Table 3.1 A comparison of the remaining RhB and COD in the solution at

reaction time of the 50th

and 240th

min .................................... 92

Table 4.1 Various solvent of different polarity used for the extraction of

POME .................................................................................... 116

List of Tables

XV

Table S1 Structures of various dye classes. Reproduced with permission

from reference [1] .................................................................. 175

List of Figures

XVI

List of Figures

Figure 1.1 Plausible mechanisms of biosorption. Reproduced with

permission from reference [2] .................................................... 9

Figure 1.2 Dye-Removal techniques. Reproduced with permission from

reference [3] ............................................................................. 14

Figure 1.3 Overall Fenton system classifications. Reproduced with

permission from reference [4] .................................................. 17

Figure 1.4 Catalytic generation of •OH radicals promoted by AuNps.

Reproduced with permission from reference [5] ..................... 20

Figure 1.5 Schematic illustration of the growth mechanism of Ag and Au

nanoparticles mediated by bio-reducing agents. Reproduced

with permission from a published work of the Phd candidate in

reference [6] ............................................................................. 24

Figure 2.1 Infrared spectra of (a) raw POME, (b) NaOH modified-POME,

(c) Hg(II) loaded POME, (d) Cd(II) loaded POME ................ 41

Figure 2.2 SEM micrographs of POME biomass for (a) raw powder, (b)

Cd loaded, (c) Hg loaded, (d) EDX analysis on Cd loaded

biomass, (e) EDS analysis on Hg loaded biomass .................. 42

Figure 2.3 Effects of the variables on the response for metal removal

compared with untreated biomass ( 0 M NaOH, 0 min): (a) Cd

removal, (b) Hg removal. ▲NaOH (Level 1: 0 M, Level 2: 0.2

M, Level 3: 0.6 M, Level 4: 1 M); ■ treatment time (Level 1:

60 min, Level 2: 40 min; Level 3: 20 min; Level 4: 0 min) .... 43

Figure 2.4 (a) Effect of contact time on the metal uptake (mgg-1

) of Cd(II)

and Hg(II); (b) Plot of linearized pseudo-second-order kinetic

models; (c) Plot of intra-particle diffusion-controlled kinetic

models; (d) Plot of the external mass diffusion-controlled

kinetic models. ●: Cd(II), ▲: Hg(II) ........................................ 45

Figure 2.5 (a) Effect of metal ions concentration on adsorption capacity of

NaOH modified- POME biomass; (b) Plot of linearized

Langmuir isotherms; ●: Cd(II), ▲: Hg(II); (c) Variation of Cd2+

species with increasing initial Cd concentration; (d) Variation

of HgCl2 (aq) species with increasing initial Hg concentration

(The percent of metal species was calculated by VISUAL

MINTEQ) ................................................................................. 52

List of Figures

XVII

Figure 2.6 (a) Effect of ionic strength on adsorption capacity of NaOH

modified-POME biomass for Cd removal; (b) Effect of ionic

strength on adsorption capacity of NaOH modified-POME

biomass for Hg removal (The variation of metal species was

calculated by VISUAL MINTEQ) ........................................... 53

Figure 2.7 (a) Effect of biosorbent dosage on Cd removal; (b) Effect of

biosorbent dosage on Hg removal (The variation of metal

species was calculated by VISUAL MINTEQ). ■ Cd or Hg

removal (%); ▲ Cd or Hg uptake (mg/g) ................................ 54

Figure 3.1 Chemical structure of RhB and the UV-vis absorption spectra

of RhB Solution ....................................................................... 64

Figure 3.2 TEM images and XRD-diffractogram of RHSi-Fe ................. 69

Figure 3.3 The temporal UV-vis spectra of RhB during its degradation in

RHSi-Fe/H2O2 system (a). Degradation of RhB as a function of

reaction time under heterogeneous and homogeneous

conditions (b). Reaction conditions: initial concentration of

RhB = 5 mg/L, temperature = 323 K, catalyst dosage = 1 g/dm3,

H2O2 amount = 0.98 mmol, pH 5.0 .......................................... 70

Figure 3.4 The changes of FTIR spectra during the degradation RhB in

RHSi-Fe/H2O2 system ............................................................. 72

Figure 3.5 The changes of COD and RhB concentration with respect to

reaction time are shown. Reaction conditions: initial

concentration of RhB = 50 mg/L, temperature = 323 K, catalyst

dosage = 1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0 ............ 74

Figure 3.6 Effect of radical scavengers (0.1 M) on the degradation of RhB.

Reaction conditions: initial concentration of RhB = 5 mg/L,

temperature = 323 K, catalyst dosage = 1 g/dm3, H2O2 amount

= 0.98 mmol, pH 5.0 ................................................................ 75

Figure 3.7 Effect of RhB concentration on the degradation of RhB. Inset

shows the variation of initial rate with respect to initial

concentrations of H2O2, (represented by [H2O2]). Reaction

conditions: temperature = 323 K, catalyst dosage = 1 g/dm3,

H2O2 amount = 0.98 mmol, pH 5.0. ........................................ 77

Figure 3.8 Effect of initial H2O2 concentration on the degradation of RhB.

Inset shows the kinetic rate constants kapp determined at

different H2O2 concentrations Reaction conditions: initial

concentration of RhB = 5 mg/L, temperature = 323 K, catalyst

dosage = 1 g/dm3, pH 5.0. ....................................................... 78

List of Figures

XVIII

Figure 3.9 Effect of catalyst dosage on the degradation of RhB. Reaction

conditions: initial concentration of RhB = 5 mg/L, temperature

= 323 K, H2O2 amount = 0.98 mmol, pH 5.0. .......................... 80

Figure 3.10 Effect of temperature on the degradation of RhB (a). Inset in

Figure 3.10(a) shows the kinetic rate constants kapp determined

at different reaction temperatures. The corresponding Arrhenius

plot is shown in (b). Reaction conditions: initial concentration

of RhB = 5 mg/L, catalyst dosage = 1 g/dm3, H2O2 amount =

0.98 mmol, pH 5.0. The change of degradation rate after pH

adjustment at 323 and 303 K is shown in (c). .......................... 81

Figure 3.11 Effect of initial pH on the degradation of RhB. Inset shows the

amount of RhB adsorbed (%) at pH 3.0, 5.0 and 7.0. Reaction

conditions: initial concentration of RhB = 5 mg/L, temperature

= 323 K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98 mmol

.................................................................................................. 85

Figure 3.12 Effect of various foreign salts on RhB degradation (a) and the

corresponding changes in degradation rate when the ionic

strength of solutions was adjusted with NaCl. Reaction

conditions: initial concentration of RhB = 5 mg/L, temperature

= 323 K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98 mmol,

pH 5.0 ....................................................................................... 88

Figure 3.13 Effect of US and UV irradiations on the degradation of RhB (a).

Inset in Figure 3.13(a) shows the shift of λmax from 15th

to 45th

min when UV irradiation was applied. The corresponding rate

constants, kapp were determined and shown in (b). The

corresponding COD removal for 50 mg/L RhB after treated

with conventional heating, US and UV irradiations at pH 5.0

and 3.0 was shown in (c) .......................................................... 94

Figure 3.14 Performance of RHSi-Fe in consecutive experiments (a). The

FTIR spectrum of fresh and reused catalysts after 1st and 3

rd

cycle of repeated use was shown in (b). Reaction conditions:

initial concentration of RhB = 5 mg/L, temperature = 323 K,

catalyst dosage = 1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0

.................................................................................................. 97

Figure 4.1 Visual observations and UV-vis absorption spectra of reaction

mixtures at different pH values [(a) control HAuCl4 solution, (b)

pH 3.0, (c) pH 4.0, (d) pH 6.0, (e) pH 8.0] after 48 h of reaction

................................................................................................ 104

List of Figures

XIX

Figure 4.2 Visual observations and UV-vis absorption spectra of reaction

mixtures with varying concentration of HAuCl4 (mM) [(a) 0.1,

(b) 0.25, (c) 0.50, (d) 1.0, (e) 2.0] .......................................... 107

Figure 4.3 Visual observations and UV-vis absorption spectra of reaction

mixtures at different temperatures [(a) 25°C, (b) 40°C, (c) 60°C]

at pH 3 after 3 h of reaction ................................................... 108

Figure 4.4 TEM images of AuNps synthesized using POME at different

initial pH values: (a) pH 3, (b) pH 8 ...................................... 109

Figure 4.5 TEM images of AuNps synthesized using POME at pH 3.0 for

(C) 1 h, (D) 3 h, (E) 9 h and (F) 48 h of reaction time .......... 110

Figure 4.6 TEM image (A) and histogram of particle size distribution (B)

of AuNps synthesized using POME at pH 3 with reaction time

of 3 h. (C) XRD spectrum of AuNps synthesized using POME.

The principal Bragg’s reflections are identified ..................... 111

Figure 4.7 XRD spectrum of AuNps synthesized using POME. The

principal Bragg’s reflections are identified ........................... 112

Figure 4.8 AuNps that was freshly prepared (a) and after stored at 4°C for

more than 6 months (b) .......................................................... 113

Figure 4.9 FTIR spectra of (a) POME powder before and (b) after gold

reduction and (c) AuNps synthesized .................................... 114

Figure 4.10 UV-vis spectra of AuNps produced using various solvent

extracts after 48 h of reaction ................................................ 116

Figure 4.11 The UV-vis spectra of AuNps (A) before and (B) after mercury

treatment. Inset is the visual observation of AuNps before and

after treated with mercury ..................................................... 118

Figure 5.1 UV-vis absorbance spectra of the Au@SiO2 dispersion and

Au@RHSi-Fe gel .................................................................. 127

Figure 5.2 TEM micrographs of AuNps before (a and b) and after its

incorporation into the Au@RHSi-Fe composite (c and d) at

different magnifications ......................................................... 128

Figure 5.3 XRD-diffractogram of Au@RHSi-Fe (a) and AuNps (b) ..... 129

Figure 5.4 EDX analysis of Au@RHSi-Fe ............................................ 129

Figure 5.5 The temporal evolution of the spectral changes during the

process of RhB degradation in the Au@RHSi-Fe/H2O2 system

after its exposure to UV irradiation (a) and a comparison

List of Figures

XX

between degradation of RHSi-Fe and Au@RHSi-Fe as a

function of reaction time (b) ................................................... 130

Figure 5.6 The plot of –ln(Ct/C0) as a function of reaction time, t and the

corresponding kapp for both RHSi-Fe/H2O2 and Au@RHSi-

Fe/H2O2 .................................................................................. 131

Figure 5.7 A comparison between degradation of RhB in the presence of

RHSi-Fe/H2O2 and Au@RHSi-Fe/H2O2 systems at pH 5.0

under room temperature as a function of reaction time ........ 132

Figure 5.8 Catalytic performances of Au@RHSi-Fe/H2O2 (a) and RHSi-

Fe/H2O2 (b) systems at different pH under room temperature as

a function of reaction time. A comparison of the performance of

both catalyst at the 50th

min during the course of reaction was

depicted in (c) ......................................................................... 133

Figure 5.9 Possible mechanisms proposed for the improved catalytic

activity of Au@RHSi-Fe ........................................................ 137

Figure S1 Schematic diagram of oil extraction from oil palm and POME

generation (dashed line represents byproduct/waste stream) 174

List of Abbreviations

XXI

List of Abbreviations

A ANOVA Analysis of variance

AOPs Advanced Oxidation Processes

Au@RHSi-Fe gold-silica-iron composite material

Au@SiO2 Synthesis of silica-coated AuNps

Au/HO-npD Gold Nanoparticles grafted on nanoparticulate

diamond

AuNps Gold nanoparticles

APTMS 3-aminopropyl-trimethoxysilane

B BET Specific Brunauer-Emmett-Teller

C Cb Conduction band

COD Chemical oxygen demand

D DI Deionized

D-R Dubinin-Radushkevich

E

ecb-

Negative charge in the conduction band EDX Energy dispersive X-ray

F FTIR Fourier transform infrared spectroscopy

FE-SEM Field emission scanning electronic microscopy

H HSAB Hard and soft acids and bases

hvb+

Hole in the valence band

I ICP-OES Inductively coupled plasma optical emission

spectrometry

ICP-MS Inductively coupled plasma mass spectrometer

J JCPDS Joint committee on powder diffraction standards

M MPOME Palm oil mill effluent modified under the

optimum condition

List of Abbreviations

XXII

O OAD Orthogonal array design

P POME Palm oil mill effluent

R RHA Rice hull ash-based

RhB Rhodamine B

RHSi-Fe Rice hull-based silica supported iron catalyst

ROS Reactive oxygen species

S SPR Surface plasmon resonance

T TEM Transmission electron microscope

TEOS Tetraethoxysilane

TMOS Tetramethoxysilane

U US Ultrasonic

UV Ultraviolet

UV-vis Ultraviolet-visible

V Vb Valence band

X XRD X-ray powder diffraction

List of Symbols

XXIII

List of Symbols

b (mgg-1

) Langmuir constants related to the sorption

capacity

β (mol2kJ

-2) A constant related to sorption energy.

C0 (mgL-1

) Initial concentrations

Ct (mgL-1

) Concentrations at any time, t

Ce (mgL-1

) Concentrations at equilibrium

c Intercept

D (cm2/s) Diffusion coefficient

E (kJmol-1

) Sorption mean free energy

Ea (kJ/mol) Activation energy

Ed (%) Desorption efficiency

ε Polanyi sorption potential

F Fractional attainment of equilibrium

ΔG0

(kJ/mol) Standard free energy of change

ΔH0

(kJ)

Enthalpy

K (Lmg-1

) Langmuir constants related to the energy of

adsorption

Kf (mgg-1

) Relative adsorption capacity

k (gmg-1

min-1

) Equilibrium rate constant of the pseudo-second-

order kinetics

kapp (min-1

) Pseudo-first order apparent rate constant

kd (mgL-1

min-0.5

) Initial rate of intra-particle diffusion

kf (s-1

) Film diffusion rate constant

L (cm) Thickness of the stagnant liquid film or the pore

length

M (g) Amount of adsorbent added to solution

mb (mgg-1

) Amount of metal ions adsorbed onto the biomass

after adsorption

List of Symbols

XXIV

md (mgg-1

) Amount of metal ions released to the bulk

solution after desorption

1/n Dimensionless parameter related to the energy

heterogeneity of the system and size of the

adsorbed molecules.

Qe (mmolg-1

) Amount of heavy metals adsorbed in mmol per g

of biomass

Qm (mmolg-1

) Maximum sorption capacity in mmol per g of

biomass

qt (mgg-1

) Amount of metal ions adsorbed per unit mass of

adsorbent at any time, t

qe (mgg-1

) Amount of metal ions adsorbed per unit mass of

adsorbent at equilibrium

R (Jmol-1

K-1

) Gas constant (8.314 Jmol-1

K-1

)

R2

Regression coefficients

RL Dimensionless separation factor

λmax (nm) Maximum absorption wavelength

S/L Solid-to-liquid

ΔS0 (JK

-1)

Entropy

T (K) Temperature

Ø Reaction-diffusion modulus (Thiele modulus)

V (L) Volume of solution

Chapter 1

1

Chapter 1 Introduction

1.1 Crisis of Heavy Metal Pollution and Resource Scarcity

Heavy metals are natural components of the earth’s crust. Anthropogenic

activities have caused an increasing release of heavy metals into the

environment and made them one of the most dangerous contaminants. Primary

sources of pollutions include acid mine drainage, effluents discharged from

electroplating, tanning, pigment and battery manufacturing industries. Heavy

metals are generally more persistent in the environment than organic

pollutants because they can neither be degraded nor destroyed. Consequently,

they remain indefinitely in the environment and are subject to

biomagnifications or bioaccumulation up the food chain. These heavy metals

usually exist in their most stable oxidation states e.g. Cd2+

, Pb2+

, Hg2+

, in

which they are able to react with the body’s biomolecules to form extremely

stable biotoxic compounds, hence posing a significant threat to public health

[7].

Several cases of disasters associated with heavy metal poisoning e.g.

Minamata disease (1932-1968) and Sandoz disaster (1986) due to mercury

poisoning, Itai-itai disease (1930-1960) due to cadmium poisoning and Nigeria

disaster (2010) due to lead poisoning have led to increased public awareness

about the threat of toxic heavy metals [8-11]. In fact, efforts have been made

to reduce the use of toxic heavy metals, for example mercury is being

displaced from industry by the introduction of new technologies and the major

spread of lead in the environment has been curbed by the introduction of

unleaded gasoline. Nevertheless, the increasing use of cadmium and

Chapter 1

2

chromium which are highly toxic as well as the widespread application of

other heavy metals such as zinc, copper and nickel in various industrial

processes still remains a serious threat to humans and the environment [12].

According to the regulations set by National Environment Agency of

Singapore (NEA), the maximum allowable discharge limits of metals in trade

effluent into the public sewer are listed in Table 1.1.

Table 1.1 The maximum allowable concentrations of metals in trade

effluent.

List of Metals Limit in milligrams per litre

of trade effluent

Cadmium 1

Chromium (trivalent and hexavalent) 5

Copper 5

Lead 5

Mercury 0.5

Nickel 10

Selenium 10

Silver 5

Zinc 10

Source: http://app2.nea.gov.sg/waterpollution_te.aspx

Despite the toxicity of heavy metals, they are actually valuable natural

resources that contribute to the economic and social development. The

increasing demand for these finite stocks of metal resources will eventually

lead to the crisis of resource exhaustion, which would have serious

implications for human society [13]. Therefore, the development of

decontamination technologies that could provide the ease of metal recovery

Chapter 1

3

which allows them to be reintroduced into the industrial process or resold is of

great research interest.

1.1.1 Conventional Methods for Remediation of Heavy Metal Pollution

Conventional methods for removing heavy metals are mainly based on

physicochemical treatment, such as chemical precipitation,

coagulation/flocculation, ion exchange, membrane filtration, adsorption, and

electrochemical treatment. The most widely used method for removing heavy

metals from solution is by chemical precipitation, in which the soluble metal is

converted into an insoluble form by using precipitation agent such as

hydroxide, carbonate or sulphide. The large amount of toxic sludge generated

is the major disadvantage of this method. The second widely used method is

ion exchange, in which the ion-exchange resin acts as a concentrator of metals.

However, the efficiency of this method is limited by the presence of other

competing ions [14].

Coagulation is one of the most important methods for wastewater treatment

wherein it works through the destabilization of colloids by neutralizing the

forces that keep them apart. Coagulation is followed by flocculation of the

unstable particles in order to increase their size and form bulky floccules that

can be settled out [15]. Generally, coagulation/flocculation cannot treat heavy

metal-laden effluent completely without being used in combination with other

techniques [16]. Electrochemical methods involve the plating-out of metal

ions on a cathode surface and recovery of metals in the elemental metal state.

Nevertheless, the feasibility of this method is limited by the requirement of

relatively high capital cost and huge consumption of electricity.

Chapter 1

4

Membrane filtration is capable to remove both organic and inorganic

contaminants such as heavy metals, but the high capital cost and problem of

membrane fouling are the major drawbacks. Amongst all the treatment

methods, adsorption is regarded as the most popular method for the removal of

heavy metals from wastewater due to its flexibility in design and the ability to

produce high-quality treated effluent. Adsorption by activated carbon is the

most efficient classical way but the high cost of activated carbon is prohibitive.

The use of natural zeolite is relatively less costly but is limited by its low

efficiency [17].

1.1.2 Biosorption as a Non-destructive Remediation Tool for Heavy

Metal Pollution

Most of the conventional methods mentioned in section 1.1.1 seem to treat the

metal ions as a waste only without providing the ease of metal recovery. In

addition, these methods are either too costly or ineffective to treat large

volumes of wastewater containing low metal concentration (1-100 mg/L) [18].

In this regard, non-destructive technique such as adsorption process which

provides the possibility of recovering the sorbed metal ions through desorption

appears to be the most appropriate in treating metal laden effluents as

compared to those destructive techniques. However, due to the high cost and

limited efficiency of commercially available adsorbents, there is a need to

search for an effective technology that is economically and environmentally

viable. Since the last few decades, a great deal of research interest has been

directed toward biosorption. Biosorption is a relatively new decontamination

technology that has been proven to be a very promising process in the fight

Chapter 1

5

against metal pollution. The major advantages of biosorption over

conventional treatment methods include low operating cost, reduced

requirement for chemicals, minimization of the volume of chemical or

biological sludge and high efficiency in detoxifying very dilute effluents.

Moreover, it also offers the ease of regeneration of biosorbents and possibility

of metal recovery [2, 19]. The use of these natural biosorbents is in

compliance with the concept of “cleaner production”, in which both the

quantity and the toxicity of emissions and waste could be minimized due to

the reduced use of hazardous chemicals. A comparison of the advantages and

disadvantages of conventional treatment methods as compared to biological

treatment is made in Table 1.2.

Table 1.2 The advantages and disadvantages of conventional treatment

methods as compared to biosorption.

Methods Advantages Disadvantages

Chemical

precipitation

Simple process

Low capital cost

Applicable to most metals

Large amount of metal-laden

sludge produced

Difficult separation

Ineffective removal for low

concentration

Ion exchange

Effective removal at low

concentration

Ease of regeneration

Pure effluent metal

recovery possible

High cost

Sensitive to the presence of

competing ions

Coagulation/

flocculation

Sludge settling and

dewatering characteristic

Bacteria inactivation

capability

Large solvent consumption

High cost

High sludge disposal

requirement

Chapter 1

6

Electrochemic

al treatment

No consumption of

chemicals

Tolerant to suspended

solids

Ability to treat effluents >

2000 mg dm-3

Metal recovery possible

Ineffective for low

concentration

High running cost and power

consumption

Possible release of

inflammable H2

Additional cost for floc

filtration

Membrane

filtration

Low solid waste

generation Low chemical

consumption

High efficiency

High capital and running cost

Limited flow rates

Problem of membrane

fouling

Sensitive to the presence of

competing ions

Adsorption

Using

activated

carbon

Applicable to most metals

High efficiency

More for remove organic

pollutants

High cost of activated carbon

No regeneration

Performance varies with

adsorbent

Using natural

zeolite

Applicable to most metals

Relatively less costly

material

Low efficiency

Using

biosorbent

Low operating cost

Minimization of chemical

and biological sludge

Effective removal at low

concentration

Reduced consumption of

chemicals

Ease of regeneration of

biosorbents and metal

recovery

Characteristics of biosorbents

cannot be biologically

controlled

Irrespective of the metal

value, desorption need to be

done for further re-

employment of biosorbents

1.1.2.1 Basic Principles of Biosorption

Biosorption can be used to describe any system where interaction between a

sorbate (atom, molecule or ion) and a biosorbent (solid surface of a biological

Chapter 1

7

matrix) results in an accumulation of the sorbate at the biosorbent interface,

and therefore a reduction of the sorbate concentration in solution [20]. In fact,

biosorption is a unique property of living or dead biomass to interact with

substances such as metals whereby reduction of sorbate concentration can be

achieved. Thus, it has been widely proposed as a promising alternative for

metal remediation due to its low cost and green nature. Most of the

biosorption research has been performed with metals and related elements due

to their high affinity to biomass. In general, the emphases have been laid on

toxic heavy metals because they are not only a class of key environmental

pollutants but also possess high market value that renders their removal and

recovery a critical aspect.

1.1.2.2 Plausible Mechanisms of Biosorption

The metal-binding mechanisms responsible for metal uptake in biosorption

involve chemisorption (by ion exchange, complexation, coordination and

chelation), physical adsorption and micro-precipitation. There are also

possible oxidation/reduction reactions taking place on the biosorbent [21].

Various plausible mechanisms of biosorption for metal ions are depicted in

Figure 1.1 [2]. Due to the complexity of biomass, the biosorption process

might be a result of the interplay of several different mechanisms and the

identification of each single step is hardly achieved.

Amongst various plausible mechanisms of biosorption, ion exchange appears

to be the principal mechanism because it explains many of the observations

made during heavy metal uptake experiments [19, 22]. In biosorption, ion

exchange could be considered as the replacement of an ion previously existing

Chapter 1

8

in biosorbent with another ion in solution. Several functional groups that are

available in the structural components of many biomass material are known to

be potential ion exchange sites, such as carboxyl, sulphate, phosphate and

amine, making them potential biosorbents for sequestering metal ions [23].

The binding mechanisms involved during this ion exchange process may range

from physical (electrostatic or van der Waals forces) to chemical (ionic and

covalent), depending on the solution chemistry and available binding sites on

the biosorbent. Since most of the heavy metals precipitate at pH > 5.5, the

bound metal species can act as loci for subsequent deposition and contribute to

the metal uptake. Hence, micro-precipitation arising from the hydrolysis

product might take place simultaneously in addition to ion exchange at higher

pH.

Complexation could happen between any cations and molecules or anions

containing free electron pairs (ligands). Chelation refers to the complex

formation with multidentate ligands. It has been shown in several studies that

heavier metal ions usually have higher binding affinities to the multidentate

ligands whereby chelation of metal ions occurs. The underlying reason was

related to the stereochemical effect, wherein a larger ion might fit better to a

binding site with two distant functional groups [24]. Since most of the

biosorption processes appear to be reversible, the nature of bond formed

during the chelation process was believed to be most likely electrostatic in

nature.

Chapter 1

9

Figure 1.1 Plausible mechanisms of biosorption. Reproduced with

permission from reference [2].

In some cases, the nature of complexation in biosorption could be predicted by

the hard and soft acids and bases (HSAB) theory. According to HSAB theory,

metal cations with low polarizability such as Cr3+

, Fe3+

and Co3+

are

considered as “hard spheres” that bind preferentially to hard ligands such as

oxygen-containing functional groups. In contrast, metal cations with high

polarizability such as Cd2+

, Hg2+

, Pb2+

, and Au3+

are considered as “soft

spheres” that bind preferentially to soft ligands such as sulphur and nitrogen-

containing groups. Generally, bonds formed between hard spheres and hard

ligands will be predominantly ionic whereas bonds formed between soft

spheres and soft ligands are more covalent in nature. Although this hard and

soft scheme can be used to describe some of the biosorption process, there is

Chapter 1

10

no absolute distinction that we can follow due to varying properties of metal

species and different nature of biomass [20].

1.1.3 Palm Oil Mill Effluent (POME)

Oil palm (Elaeis guineensis) plantations are widely available in tropical

countries such as Indonesia, Malaysia, Thailand and Colombia. POME is a

residue generated during the production of palm oil. Approximately 0.5-0.75

tonnes of POME was generated from every tonne of oil palm fresh fruit

produced [25]. The disposal of large amounts of POME directly in the soil or

in natural waters may pose a risk of environmental pollution in an

uncontrolled manner [26]. Schematic diagram of POME generation during the

processing of oil palm for crude palm oil production is depicted in Figure S1,

Appendix 1 [27].

1.1.3.1 Potential of POME for Biosorption of Heavy Metals

The promising performance in metal biosorption exhibited by plant-based agro

waste particularly those containing high levels of lignocellulosic materials has

been demonstrated in many studies [28-30]. Lignocellulosic substances are

mainly composed of high-molecular weight components including cellulose,

hemicelluloses and lignin which contribute most of the mass [31]. In addition,

it also contains small quantity of low-molecular weight extractives such as

resin (terpenes, lignans, and other aromatics), fats, waxes, fatty acids, alcohols,

terpentines, tannins, and flavonoids. The abundant availability of these

lignocellulosic materials at low or no cost presents a significant advantage for

their large-scale application as biosorbent for the removal of heavy metals in

water treatment.

Chapter 1

11

Amongst those lignocellulosic materials that have been studied as potential

biosorbents, the use of oil palm-based biomass is worth to be highlighted

owing to the efficient use of land, water, nitrogen and energy resources with

relatively low pesticide consumption during the growth of this biomass. De

Vries et al. used a set of nine sustainability indicators focused on resource use

efficiency, soil quality, net energy production and greenhouse gas emissions to

evaluate the production-ecological sustainability of several major crops. Their

study reveals that oil palm, sugarcane and sweet sorghum appeared to be the

most sustainable crops in which oil palm was the most sustainable with respect

to the maintenance of soil quality [32]. Hence, this makes oil palm biomass

and its associated wastes a comparatively superior and sustainable material as

compared to many current large-scale agricultural crops.

In fact, POME contains cellulose, hemicelluloses and lignin that are

potentially useful for metal biosorption [33]. The use of POME as a low-cost

biosorbent is also more affordable to many developing countries as compared

to commercially available adsorbents. Hence, it is worthwhile exploring the

potential of POME in metal biosorption because it is not only more affordable

for poorer countries but also able to reduce the uncontrolled discharges of this

agro waste by adding a value to them.

1.1.4 Orthogonal Array Design as a Chemometric Method for

Optimization of Biomass Pretreatment

Orthogonal array design (OAD) is a chemometric approach which combines

the advantages of both simplex and factorial design. By applying this

chemometric method, much more information can be extracted from a limited

Chapter 1

12

number of experiments. To evaluate the optimum pretreatment condition of

POME, OAD was used to assign factors to a series of experiment

combinations whereby the results can then be analyzed by using a common

mathematical procedure. By arranging experiments orthogonally, the main

effects of the variable and the interactions between them can be independently

extracted. The aim of using this chemometric method is to determine the

optimum pretreatment conditions for the system, or to determine the region,

which satisfies the operating specifications with reduced number of

experiments [34, 35].

1.1.5 Visual MINTEQ for the Calculations of Metal Speciation

Visual MINTEQ provides a chemical equilibrium model for the calculation of

metal speciation, solubility equilibria, sorption etc. for natural waters. It was

the second-most used chemical equilibrium software application among

researchers publishing in Elsevier journals. It combines state-of-the-art

descriptions of sorption and complexation reactions with easy-to-use menus

and options for importing and exporting data from/to Excel.

1.2 Crisis of Organic Dye Pollution

Organic dyes are widely used in industrial applications such as textiles, leather,

food products, cosmetics, pharmaceuticals, painting and paper printing. In

particular, textile industries which are the biggest users of dyes generate large

volume of wastewater containing pigments, dyes and their derivatives during

different steps in the dyeing and finishing processes. Dyes can be

Chapter 1

13

distinguished by their compound structure, including phthalocyanins,

anthraquinones, quinine-imines and xanthenes etc. The structures of various

classes of dyes along with an example of each class were summarized in Table

S1, Appendix 2 [1]. It is estimated that approximately 10-15% of dyes

produced annually (over 7 × 105 t) is lost during the dyeing process and

released as industrial effluents [36]. These coloured effluents could affect the

photosynthetic activity in aquatic life through the hindrance of light

penetration. The hydrolysis products of many synthetic dyes are toxic and

potentially carcinogenic. This not only poses a risk to human beings but also

produces toxic effects to microorganisms which results in their resistance to

aerobic biodegradation. Due to the complex structure and synthetic origin,

many dyes do not degrade easily under mild oxidation conditions. Hence, dyes

have become one of the environmental persistent pollutants in wastewater

which render the subsequent recycling process a challenging task.

1.2.1 Conventional Methods for Remediation of Organic Dye Pollution

The most commonly used techniques for the removal of dyes from wastewater

are shown in Figure 1.2 [3]. However, most of the conventional dye-removal

techniques such as adsorption on activated carbon, sedimentation, filtration,

coagulation, reverse osmosis, etc. are non-destructive. These treatments

merely involve the transfer of pollutants from water to sludge which resulted

in the production of secondary waste that needs further disposal. Although

biodegradation is able to degrade the dye pollutants, it could be slow and

inefficient because many dye compounds are actually bio-recalcitrant [37].

Chapter 1

14

Figure 1.2 Dye-Removal techniques. Reproduced with permission from

reference [3].

1.2.2 Advanced Oxidation Processes as a Destructive Remediation Tool

for Organic Dye Pollution

Advanced Oxidation Processes (AOPs) are robust alternative destructive

techniques for dye effluent when conventional non-destructive techniques are

insufficiently effective. AOPs are based on the generation of strong oxygen-

based oxidizers (e.g. •OH, •OOH, 1

O2, O2•

) to initiate oxidative destruction of

organic pollutants. The reactive oxidizing species can be generated in situ

through either one or the combination of chemical oxidation by using ozone,

hydrogen peroxide with or without radiation sources such as ultrasound (US),

ultraviolet (UV), solar visible and thermal [4]. Among various AOPs, Fenton

reagent (e.g. Fe2+

+ H2O2) and photocatalyst (e.g. TiO2 + UV) have been

Chapter 1

15

intensively studied for environmental applications during the last few decades

due to their promising performances in the degradation of many pollutants

[38].

1.2.2.1 Fenton and Fenton-like Treatment

Classical Fenton reaction refers to the reaction between hydrogen peroxide

(H2O2) and ferrous ion (Fe2+

) in a homogeneous solution, in which hydroxyl

radical (•OH) and ferric ion (Fe3+

) are generated. The active catalyzing species

are iron ions which break down the H2O2 into •OH. Any Fenton reaction

which does not use dissolved Fe2+

homogeneous catalyst is defined as a

Fenton-like reaction. The Fenton-like reaction can be catalysed by

heterogeneous catalysts including Fe3+

, native or added iron oxides or certain

transition metals [39]. The degradation of organic dyes by Fenton process

could involve the following reactions [40-43]:

Fe2+

+ H2O2 + H+ → Fe

3+ + •OH + H2O k1 = 58 M

-1s

-1 (1.1)

Fe3+

+ H2O2 → Fe2+

+ •OOH + H+ k2 = 0.01-0.02 M

-1s

-1 (1.2)

Fe3+

+ •OOH → Fe2+

+ O2 + H+ k3 = 1.2 × 10

6 M

-1s

-1 (1.3)

Fe2+

+ •OH → Fe3+

+ OH-

k4 = 3.2 × 108 M

-1s

-1 (1.4)

H2O2 + •OH → H2O + •OOH k5 = 3.3 × 107 M

-1s

-1 (1.5)

•OH + •OH → H2O k6 = 5.3 × 109 M

-1s

-1 (1.6)

Dyes + •OH → intermediates → CO2 + H2O k7 = 3.3 × 107

M-1

s-1

(1.7)

The •OH generated is non-selective and exhibits oxidation potential of

approximately 2.80 V at pH 3.0 (•OH + H+ + e

- → H2O; E

0=2.80 V),

surpassed only by fluorine and sulfate radicals in the scale of strong oxidizing

Chapter 1

16

agents [44]. Thus, •OH radicals are useful in degrading persistent organic

pollutants.

1.2.2.2 Homogeneous vs. Heterogeneous Fenton Treatment

Fenton process can be conducted either homogeneously or heterogeneously

under various combinations as depicted in Figure 1.3 [4]. Although Fenton

and Fenton-like reagents are capable to generate •OH radicals in homogeneous

reactions, their applications are limited by narrow working pH range (< 4),

high H2O2 consumption and the difficulty of separation and recovery of the

iron species in solution [45]. To overcome these drawbacks, heterogeneous

Fenton-like catalysts have attracted extensive research interest. However, in

heterogeneous system, the reaction rate is not only affected by the chemistry

of reacting species but also the physical steps take place on the surface of the

solid catalyst where mass transfer limited adsorption of reactant molecules

occurs. Thus, the surface morphology and catalyst matrix structure are

important in determining the kinetic rate, efficiency and stability of the

reaction.

Chapter 1

17

Figure 1.3 Overall Fenton system classifications. Reproduced with

permission from reference [4] .

1.2.3 Silica as Catalyst Support for Heterogeneous Fenton Treatment

Various materials have been used as catalyst supports for immobilizing Fe3+

or

Fe2+

such as activated carbon [46], carbon nanotube [47], clay, zeolite, and

silica [46, 48]. In particular, silica supported iron-based catalysts have been

actively researched owing to their large specific surface area and excellent

stability imparted by silica backbone [49]. Silica (SiO2) can exist in a variety

of forms such as gel, crystalline and amorphous material. The structure of

SiO2 is based upon a SiO4 tetrahedron, where each silicon atom is surrounded

by four oxygen atoms and each oxygen atom is bound to two silicon atoms.

There are two types of functional groups on the surface of silica: silanol

groups (Si-O-H) and siloxane groups (Si-O-Si). The silanol groups are the

Chapter 1

18

locus of activity for any process taking place on the surface whereas the

siloxane groups are considered non-reactive [50].

In fact, silica materials such as amorphous silica and mesoporous silica have

also been used as photocatalyst for non-oxidative and oxidative reactions. In

addition, pure silica was also found to exhibit photoactivity to some extent

[51]. Therefore, the inert and possible photocatalytic properties of silica make

it a suitable catalyst support to anchor ferric ions in the synthesis of Fenton

catalyst.

1.2.3.1 Potential of Rice Hull Waste as Natural Feedstock of Silica

Although silica is the most abundant oxide in the earth’s crust, it is

predominantly prepared by synthetic means for its use in technological

applications [50]. Unlike tetraethoxysilane (TEOS) and tetramethoxysilane

(TMOS) which are well-known precursors to produce silica particles, rice hull

which comprises about 20% of silica appears to be an attractive natural source

of silica [52, 53]. Rice (Oryza sativa L.) is one of the major food sources for

over half world’s population. Approximately 600 million tonnes of rice are

produced annually. Rice hull is the outermost layer of the paddy grain, which

constitutes about 20% of the paddy weight. During the rice-milling process,

much of the rice hull produced is either burnt openly or disposed as waste. The

burning activities can lead to serious environmental and health problems [50].

In this regard, effective utilization of these rice hull wastes is important to

solve the disposal problem. Thus, the use of rice hull as a natural feedstock of

silica for the synthesis of a heterogeneous Fenton-like catalyst is advantageous.

Chapter 1

19

1.2.3.2 Enhancement of catalytic performance of Fenton catalyst by

incorporation of AuNps

Amorphous silica materials are found to exhibit photocatalytic activities in

some reactions [54]. In addition, photocatalytic degradation of organic dye by

silica nanoparticles has been reported as well [55]. Silica nanoparticles can

absorb UV light and transfer an electron from its valence band (vb) to the

conduction band (cb) upon excitation. This process generates a positively

charged hole in the valence band (hvb+) and a negative charge in the

conduction band (ecb-). The hvb

+ can then react with the chemisorbed H2O

molecules to form reactive species such as •OH radicals which could be used

to attack dye molecules successively. The ecb- could react with acceptor such

as dissolved O2 and transform it into super oxide radical anion (O2•

), leading

to the additional formation of •OOH. On the other hand, hvb+

could also

interact with donor such as OH- and •OOH and further increase the formation

of •OH radial.

The ecb- and hvb

+ could also trapped in the surface states and react with the

adsorbed species. However, the efficiency of silica catalyst is limited by the

recombination of ecb- and hvb

+. The recombination process can occur within

few nanoseconds on the surface of the particle and the resulting energy is

dissipated as heat. The presence of a gold ion can enhance the formation of

•OH radical by reducing the recombination of charges. During the reaction,

Au ions are able to trap electrons in the conduction band of SiO2 and generate

holes in which they can act as electron scavengers. Therefore, the

incorporation of AuNps in the silica matrix can reduce the electron-hole

Chapter 1

20

recombination and increase the photocatalytic efficiency by accelerating the

formation of •OH radical [56].

In addition to the enhancement of photocatalytic activity of silica material,

AuNps grafted on nanoparticulate diamond (Au/HO-npD) have also been

reported to be a highly efficient catalyst in a Fenton process that use

exclusively H2O2 without using any of sources of irradiation (performed in the

dark). The authors proposed that the catalytic generation of •OH radical

involves a swing between positive and neutral Au states in which Au act as an

electron relay from the oxidation to the reduction semi-reaction as shown in

the Figure 1.4 [5].

Figure 1.4 Catalytic generation of •OH radicals promoted by AuNps.

Reproduced with permission from reference [5].

Chapter 1

21

1.2.4 Physical and Chemical Syntheses of Nanoparticles

Currently, the most common method for the synthesis of metallic

nanoparticles is through chemical reduction of metal salts in solution phase

[57]. Depending on the condition of the reaction mixture, metal ions may

favor either the process of nucleation or aggregation to form small metal

clusters. This synthetic method usually employs chemicals such as hydrazine,

sodium borohydride and hydrogen as reducing agents [58, 59]. Synthetic or

natural polymers such as natural rubber [60], chitosan [61], cellulose [62] and

copolymer micelles [63] have been used as stabilizers against oxidation and

coalescence in nanocomposites. Due to the hydrophobicity of these chemicals,

organic solvents such as ethanol, dimethyl formamide, ethylene glycol,

toluene and chloroform are usually used [64]. Although some of the organic

solvents such as dimethyl formamide and ethylene glycol are biodegradable to

some extent, toluene and chloroform are toxic and hazardous. The use of these

organic solvents could be detrimental to the environment and limits the scale

of production. Besides, some of the toxic chemicals might contaminate the

surfaces of nanoparticles and make them unsuitable for certain biomedical

applications [65].

Physical approaches to synthesize metallic nanoparticles include UV

irradiation [66], sonochemistry [67], radiolysis [68], laser ablation [69] and so

forth. During physical fabrication, metallic atoms are vaporized followed by

condensation on various supports, in which the metallic atoms are rearranged

and assembled as small cluster of metallic nanoparticles [63]. The main

advantage of the physical approach is that nanoparticles with high purity and

Chapter 1

22

desired size can be selectively synthesized [70]. However, these processes

usually require complicated instruments, electrical and radioactive heating as

well as high power consumption, which results in high operating cost.

1.2.5 Biosynthesis of Nanoparticles

Biosynthesis of nanoparticles is an alternative way that is compatible with

green chemistry principles, in which biomolecules secreted by the biomass can

act as both reducing and capping agents during the reaction. Therefore, this

reaction can be considered as a green chemical process that can minimize the

usage of hazardous chemicals [71].

1.2.5.1 Basic Principles of Biosynthesis

During the formation of metallic nanoparticles, there are three mandatory

components, which are the reducing agent, stabilizing agent and a solvent

medium that can solubilize the metal of interest [64]. Biosynthesis of

nanoparticles is regarded as a green process because the biomass itself can act

as both reducing and stabilizing agents. In addition, most of the plant-

mediated syntheses of metallic nanoparticles could be performed in aqueous

medium instead of organic solvents, which is apparently more

environmentally benign and cost-effective. In the absence of other strong

ligands, metallic ions could interact with the biomass through ionic binding

with the bioorganic reducing agents such as flavonoids or terpenoids. It is

believed that the adsorption of bioreducing agents on the surface of metallic

nanoparticles is attributable to the presence of π-electrons and carbonyl groups

in their molecular structures.

Chapter 1

23

Similar to biosorption, HSAB theory is also applicable here. A soft metal like

Au(III) existing in the form of [AuCl4]- prefers to bind to the biomass mainly

through soft ligands such as amino and sulfhydryl groups, especially when the

soft ligands are positively charged at low pH [65, 72]. Several studies

suggested that the mechanism of Ostwald ripening may be involved during the

biosynthesis process, wherein smaller or quasi-solid nascent particles could

migrate and feed into their closest neighboring particles of larger size, causing

the growth of larger particles and depletion of smaller particles. The lower

surface energy of the large crystals is thermodynamically favored, whereas the

nucleation of small crystals is kinetically favored [57, 73, 74]. The small

crystals tend to dissolve in the solution and grow into larger ones, which can

achieve a lower energy state with greater stability. The excess Gibb’s free

energy associated with the nanoparticles could then be minimized by

transformation into more energetically favorable shapes, which is directed by

the bioorganic capping molecules [75].

To stabilize the nanoparticles in a dispersing medium, there must be sufficient

repulsive forces to overcome the van der Waals force that causes coagulation

[76]. The repulsive forces can be exerted through either electrostatic or steric

stabilization. Physisorbed surfactants or polymers from the bioorganic capping

agents may exert steric or electrostatic barriers around the particle surface [64].

At the beginning of the reaction, spherical nanoparticles were produced due to

the protection by sufficient stabilizing molecules. However, the particles that

formed at later stage were less stable due to availability of less stabilizing

molecules. The instability of the nanoparticles rendered the formation of

anisotropic nanostructures like triangular nanoprisms. These nanoprisms

Chapter 1

24

possess high surface energy and they undergo a shrinking process in order to

reduce the surface energy, resulting in the formation of truncated

nanotriangles [77]. Depending on the reducing power and availability of

bioreducing and stabilizing agents, the biosynthesized nanoparticles obtained

may have higher proportion of a particular shape. A possible growth

mechanism for the formation of different shapes of nanoparticles such as

nanorods, nanowires, nanoprisms, hexagonal nanoparticles is illustrated in

Figure 1.5.

Figure 1.5 Schematic illustration of the growth mechanism of Ag and Au

nanoparticles mediated by bio-reducing agents. Reproduced with permission

from a published work of the Phd candidate in reference [6].

Although chemical and physical methods have successfully produced well-

defined nanoparticles, these processes are usually expensive and involve the

use of toxic chemicals. In addition, the contact of nanoparticles to human body

is inevitable for certain biomedical applications such as drug delivery. Hence,

Chapter 1

25

it is necessary to develop an eco-friendly production method that can provide

nanoparticles with low toxicity and biocompability. Thus, among the primary

goals of nanotechnology is to produce nanoparticles in a clean and cost-

effective manner. To achieve this goal, many researchers have diverted their

interest to biological synthesis of nanoparticles [78, 79]. However, due to the

limited capacity of plants for reducing metal ions, biosynthesis process usually

works well for metal ions with large positive electrochemical potential such as

Au and Ag ions [80]. Hence, biosynthesis was used as a low-cost and eco-

friendly approach in Chapter 4 for the synthesis of AuNps.

1.2.5.2 Potential of POME for Biosynthesis of AuNps

In addition to biosorption, the high content of phenolic acids and flavonoids in

POME are also principal components for the biosynthesis of nanoparticles

[81]. The hydroxyl groups available in these compounds could participate in

the bioreduction of metal ions and result in the formation of nanoparticles.

Thus, biosynthesis is another attractive area for the application of POME. In

fact, for metal ions with large positive electrochemical potential such as Au

and Ag ions, both the processes of biosorption and biosynthesis could compete

with each other. The dominating effect of each could be adjusted by varying

the overall surface charge of the biomass, which is possible to be achieved by

tuning the solution pH.

Chapter 1

26

1.3 Research Scope

The main objective of this dissertation is to explore the potential use of eco-

friendly materials transformed from low-cost agro wastes for applications in

remediation of persistent pollutants in water. The motivation of our work is to

provide a more affordable solution for developing countries that are not

designed and equipped for handling toxic wastes. The incorporation of a low-

cost decontamination strategy as a “pretreatment step” to treat these persistent

pollutants before they are being discharged into the water body or sent to

municipal sewage treatment plants is important to minimize possible

environmental pollution and facilitate the subsequent recovery process. In

addition, the effective utilization of these low-cost agro wastes might also

bring in new opportunities for the development of biomass pre-processing

plants, especially in developing countries that are rich with biomass resources.

Heavy metals are persistent inorganic pollutants that can neither be degraded

nor destroyed. Some of the heavy metals are highly toxic. In particular, toxic

heavy metals such as cadmium, lead and mercury are able to form extremely

harmful biotoxic compounds which had led to severe disasters in human

history. Nevertheless, the demand for these finite stocks of metal resource is

always unavoidable in some anthropogenic activities which are important for

the economic and social development. Thus, non-destructive techniques which

provide the possibility of recovering the sorbed metals through desorption

appears to be a feasible approach. In view of this, biosorption which offers the

advantages of low operating cost, reduced chemical consumptions, high

Chapter 1

27

adsorption efficiency and minimization of the production of secondary sludge

was investigated as a remediation tool for heavy metal pollution.

A common rationale for selecting the potential biosorbent material is based on

their ease of availability and cost of production. POME which is an agro waste

discharged in large amount from the oil palm industry appears to be a good

candidate of biosorbent by taking into account their ease of availability and

massive growth in tropical countries. In Chapter 2, the potential use of POME

as a low-cost biosorbent for the removal of Cd(II) and Hg(II) was investigated.

In this study, OAD has been used as a chemometric method to determine the

optimum pretreatment condition for the modification of biosorbent. The

influences of different process parameters such as adsorption time, initial

metal concentration and ionic strength of metal solution were studied. The

adsorption behaviour of metal ions was also investigated by fitting the

experimental data with different kinetic models and adsorption isotherms. In

addition, the recovery of metal ions and reusability of biosorbent were tested

as well.

Organic dyes are another major source of pollutants generated from the

industries. Since most of the dyes are bio-recalcitrant and have been proven to

be carcinogenic, destructive techniques such as AOPs are useful in enhancing

the treatability of these dye-containing effluents. Among various AOPs,

Fenton process which utilizes iron-based technologies has been extensively

studied due to its low cost and the fact that it is is relatively more practical as

compared to other AOPs. In Chapter 3, a rice hull-based silica supported iron

catalyst was used for the heterogeneous Fenton-like degradation of

Chapter 1

28

Rhodamine B (RhB) which has been selected as a representative of dyes from

the xanthenes class. Due to the low cost and abundant availability of rice hull

waste in agricultural industry, the use of rice hull as an alternative source of

silica for the synthesis of Fenton-like catalyst for remediation of organic dye

pollution is of great research interest. Since the use of rice hull-based silica

supported iron catalyst (RHSi-Fe) for heterogeneous Fenton-like degradation

of textile dyes is rarely studied, the aim of our work is to assess the catalytic

performance of this catalyst on the dye degradation under different operating

conditions. The influences of various reaction parameters as well as the effects

of ultrasonic (US) and ultraviolet (UV) irradiations were evaluated. Moreover,

to improve the degree of mineralization without additional dose of oxidant, a

stepwise addition strategy was also used in this study. Since the improved

performance of Fenton catalyst by incorporating AuNps incorporated into the

catalyst matrix has been reported in the literature, the effect of AuNps for the

enhancement of catalytic performance of the RHSi-Fe for Fenton-like

degradation of RhB was examined as well.

To make the preparation of catalyst a greener procedure, biosynthesized

AuNps were used. Biosynthesis of nanoparticles can be considered as a green

chemical process in which the biomass itself can act as both reducing and

capping agents during the reaction. In addition, most of the biosynthesis

processes can be carried out in aqueous solution instead of organic solvents.

The minimal use of hazardous chemicals in the biosynthesis process makes it a

low-cost and eco-friendly viable approach. In view of the high phenolic and

flavonoids content of POME, its potential in biosynthesis of AuNps was

explored in Chapter 4. The influence of various reaction parameters such as

Chapter 1

29

reaction pH, concentration of gold precursor and the interaction time to the

morphology and size of biosynthesized AuNps was investigated. The

biosynthesized AuNps were then embedded in the structure of RHSi-Fe and its

catalytic performance toward degradation of RhB was evaluated in Chapter 5.

Chapter 2

30

Chapter 2 Biosorption of Cd(II) and Hg(II) from Aqueous

Solutions using Palm Oil Mill Effluent (POME) as a Low-cost

Biosorbent

2.1 Introduction

Heavy metal pollution is a growing threat to the environment and humanity.

Among all the toxic heavy metals, cadmium and mercury are well-known for

their high toxicity and negative health impacts, and they have been reported to

cause major poisoning or health hazards. Cadmium and its compounds are

relatively water soluble as compared to other heavy metals so they are mobile

in soil and have higher tendency of bioaccumulating in living system [12].

Adverse health effects of cadmium to human beings include carcinogenicity,

kidney dysfunction, skeletal deformation (Itai-itai), cardiovascular diseases

and hypertension [82]. Mercury and its compounds from a variety of sources

could be converted into the more toxic form i.e. methylmercury chloride by

aquatic living organisms and enter the food chain [83]. Mercury poisoning can

provoke neurological damages, including irritability, paralysis, blindness,

insanity, chromosome breakage and birth defects [84]. Since these inorganic

micro-pollutants are non-biodegradable, there is a pressing need to remove

them at the site of emission before they spread to the environment.

Conventional methods for the removal of heavy metals include chemical

precipitation, reverse osmosis, electro dialysis, ultra filtration, ion exchange,

evaporation and carbon adsorption [85]. However, these techniques usually

involve high operating costs, limiting their applications in large scale

Chapter 2

31

processes for wastewater treatment. In the past two decades, the application of

several agricultural materials as biosorbents for the removal of toxic pollutants

has been reported [86-88]. During the biosorption process, the pollutants are

transferred from the water effluent to a solid phase (biosorbent) by different

mechanisms. Strategies for managing spent biosorbent include metal

desorption from the biomass, biomass dissolution, and biomass incineration

[18]. The most feasible post-treatment strategy is the regeneration of

biosorbent through metal desorption from the biomass which allows the

regenerated biosorbent for repeated uses and also offers the possibility to

reclaim the desorbed metals. The exhausted biosorbent could then be

converted into energy fuels such as methane, hydrogen, and bioethanol

through fermentation process or transformed to thermal energy by combustion.

Oil palm (Elaeis guineensis) plantations are widely available in tropical

countries such as Indonesia, Malaysia, Thailand and Colombia.

Approximately 0.5-0.75 tonne of palm oil mill effluent (POME) was generated

from every tonne of oil palm fresh fruit produced [25]. The disposal of large

amounts of POME directly in the soil or in natural waters may pose a risk of

environmental pollution in an uncontrolled manner. During the decomposition

of this waste material, the dissolved oxygen of the river or stream is rapidly

depleted, which could endanger the aquatic species [26]. In this context, the

application of POME as biosorbent not only reduces the cost with commercial

adsorbents but also provides economical and environmental advantages to

some developing countries.

Chapter 2

32

The metal uptake capacity of biomass can generally be improved by chemical

or physical treatments. NaOH modification of lignocellulosic materials has

been reported to be one of the most effective and economical methods [85].

However, the best removal efficiency is only achieved by using biomass that

was modified under the optimum condition [89]. In this study, orthogonal

array design (OAD) was used as a chemometric method to determine the

optimum pretreatment condition of POME which is to be used as a biosorbent

for the removal of Cd(II) and Hg(II) in aqueous solutions [34, 35]. In addition,

the influences of various process parameters, such as initial metal

concentration, adsorbent dose, pH and ionic strength of solutions have been

investigated by batch adsorption experiments. Recovery of metal ions and

regeneration of biosorbent were evaluated as well.

2.2 Material and Methods

2.2.1 Solutions and Reagents

All chemicals used throughout the experiments were of analytical grade and

solutions were prepared by using deionized (DI) water (resistivity ≈ 18.2 MΩ

cm) obtained from a Millipore (Billerica, MA) Direct Q purification unit.

Stock solutions of Cd(II) and Hg(II) were prepared by dissolving solid

cadmium chloride (CdCl2·2.5H2O) and mercury chloride (HgCl2) in acidified

de-ionized water to 1000 mgL-1

. Working solutions were obtained by diluting

the metal stock solutions to the desired concentrations. To adjust pH of the

metal solutions, 0.10 to 1.0 M sodium hydroxide (NaOH) or hydrochloric acid

Chapter 2

33

(HCl) solutions were used. The pH of the solutions was measured using

Metrohm 827 pH Lab meter.

2.2.2 Preparation of Biomass Material

POME used was collected from a palm oil industry (Kluang Palm Oil

Processing Mill Sdn. Bhd.) located at Kluang, Johor, Malaysia, through a one-

time collection. It was dried at 60°C in an air-supplied oven for approximately

24 h. After that, the POME was ground by an electrical blender and

subsequently sieved. The part of ground POME with diameter of particles ≤

53 µm was used. The POME powder was first washed with distilled water

under stirring at 65°C for 1 h and then washed again with n-hexane/ethanol

(1:1, v/v) under reflux conditions for 4 h. After that, the powder was dried at

105 °C and stored in a desiccator.

2.2.3 Characterization of the POME Biosorbent

The POME biosorbent was characterized by fourier transform infrared

spectroscopy (FTIR, Varian Excalibur 3100). 1 mg of POME biosorbent was

blended with 100 mg of IR grade KBr and pressed into a pellet. The spectra

were recorded at a resolution of 4 cm-1

with a scan range from 400 to 4000 cm-

1. The morphology and elemental analysis were carried out by using field

emission scanning electronic microscopy (FE-SEM) equipped with an energy

dispersive X-ray (EDX) spectrometer (JEOL JSM-6701F SEM).

2.2.4 Optimization Strategy for Pretreatment of POME

To optimize the NaOH pretreatment conditions for POME, OAD approach

was applied. The optimum pretreatment condition was obtained by using an

Chapter 2

34

OA9 (34) followed by an OA8 (2

7) matrix [90]. The removal efficiencies (%) of

Cd(II) and Hg(II) were set as the output responses. The assignment of

different factors and level settings was depicted in Table 2.1. A preliminary

study was performed and the optimum adsorption pH was determined to be pH

4.5 for both the adsorptions of Cd(II) and Hg(II). All the subsequent

adsorption tests were performed by using POME modified under the optimum

condition (MPOME). For each pretreatment, POME was treated with NaOH in

an Erlenmeyer flask at 20 °C. After that, the biomass was rinsed thoroughly

with distilled water until the pH of filtrate became neutral. NaOH modified-

POME was then dried at 105 °C and stored in a desiccator for future use.

Table 2.1 Assignment of factors and level settings for POME modified with NaOH by

using OA9 (34) followed by OA8 (2

7) matrix along with the results of output responses.

(initial Cd(II) concentration: 50 mgL-1

, initial Hg(II) concentration: 15 mgL-1

, POME dose:

3.75 gL-1

, contact time: 2 h, adsorption pH: 4.5)

OA9 (34) matrix OA8 (2

7) matrix

Trial

No.

Column No. a Responses

Trial

No.

Column No. a Responses

1

A

2

B

3

#

4

#

Cd

removal

Hg

removal

1

A

2

B

3

A x

B

4

#

5

#

6

#

7

#

Cd

removal

Hg

removal

1 0.2 60

96.01 96.15 1 0.2 40 97.57 100.00

2 0.2 40

97.36 95.94 2 0.2 40 97.36 100.00

3 0.2 20

66.39 85.84 3 0.2 20 68.22 89.05

4 0.6 60

84.85 95.72 4 0.2 20 66.39 85.84

5 0.6 40

96.68 95.95 5 0.6 40 98.05 99.38

6 0.6 20

92.75 95.76 6 0.6 40 96.68 100.00

7 1 60

74.41 86.96 7 0.6 20 92.75 94.29

8 1 40

80.37 95.10 8 0.6 20 92.33 94.20

9 1 20

94.01 96.11

a A: NaOH concentration (M), B: treatment time (min), A x B: interaction between factor A and B

Chapter 2

35

#: unassigned column (dummy factors)

2.2.5 Batch Adsorption Experiments

2.2.5.1 General Procedure for Batch Adsorption

The adsorption tests were conducted by agitating POME powder with heavy

metal solutions in Erlenmeyer flasks. The mixture was stirred at 300 rpm on a

magnetic stirrer equipped with a thermostat-controlled water bath. For kinetic

studies, stirring was briefly interrupted while aliquots of sample solution were

taken at certain time intervals. The suspensions were centrifuged at 5000 rpm

for 20 min and the supernatant solutions were filtered by 0.45-µm membrane

filters. Blanks without biosorbents were carried out in order to verify the

possibility of metal precipitation at the adsorption pH.

The residual metal concentrations in the filtrates were analyzed by a Perkin-

Elmer Dual-view Optima 5300 DV inductively coupled plasma optical

emission spectrometry (ICP-OES) system. The removal efficiency (%) of

heavy metals was computed as below:

emoval ( - )

x 100% (2.1)

where C0 and Ce (mgL-1

) are the metal concentrations before and after

adsorption. All adsorption tests were performed in triplicate and the mean

values were used. The standard errors were controlled within 1% for Cd(II)

and 5% for Hg(II) uptake.

2.2.5.2 Biosorption Kinetics

Chapter 2

36

The biosorption kinetics was studied by using 50 mgL-1

Cd(II) and Hg(II)

solutions at 20°C with contact time varied between 1 and 120 min. Pseudo-

second-order kinetic model was selected due to its good applicability in most

biosorption processes [91]. The pseudo-second-order equation is generally

represented as below:

2

1 1

t ee

tt

q kq q

(2.2)

where qt (mgg-1

), qe (mgg-1

) and k (gmg-1

min-1

) are the amount of metal ions

adsorbed at time t, the amount of metal ions adsorbed at equilibrium, and the

equilibrium rate constant of the pseudo-second-order kinetics, respectively.

The contribution of intra-particle diffusion was investigated by using Webber

and Morris equation, which is represented as below [92]:

1/ 2t dq k t c (2.3)

where kd (mgL-1

min-0.5

) is the initial rate of intra-particle diffusion and c is the

value of intercept.

Liquid film diffusion model was used to evaluate the external mass transfer

phenomenon [93, 94]. This could be applied to determine the transport of

solute molecules from the liquid phase up to the solid phase boundary:

ln(1 ) fF k t (2.4)

where F = qt/qe is the fractional attainment of equilibrium and kf (s-1

) is the

film diffusion rate constant.

Chapter 2

37

0( )ee

C C Vq

M

2.2.5.3 Biosorption Isotherms

2.2.5.3.1 Langmuir Isotherm

Langmuir isotherm in its linearized expression is shown as below [95]:

1e e

e

C C

q bK b

(2.5)

where (mgg-1

) is the solute uptake per unit weight of POME

powder at equilibrium; C0 and Ce (mgL-1

) are the initial and equilibrium metal

concentrations; V (L) is the volume of solution; M (g) is the amount of

adsorbent added to solution; b (mgg-1

) and K (Lmg-1

) are the Langmuir

constants related to the sorption capacity and energy of adsorption,

respectively.

An essential feature of Langmuir isotherm is the dimensionless separation

factor, RL, which is defined by the following equation:

1

(1 )LR

KCe

(2.6)

If 0 < RL < 1, it indicates favourable adsorption; if RL > 1, it indicates

unfavourable adsorption; if RL = 1, it indicates linear adsorption and if RL = 0,

it indicates an irreversible adsorption process [96].

2.2.5.3.2 Freundlich Isotherm

Freundlich isotherm in its linearized expression is shown as below [95]:

1log log loge f eq K C

n

(2.7)

Chapter 2

38

where Kf (mgg-1

) represents the relative adsorption capacity, and 1/n

(dimensionless) is related to the energy heterogeneity of the system and size of

the adsorbed molecules.

2.2.5.3.3 Dubinin-Radushkevich Model

Dubinin-Radushkevich (D-R) model is represented as below [97]:

2

0ln lne mQ Q (2.8)

where Qe (mmolg-1

) is the amount of heavy metals adsorbed per g of biomass,

Qm (mmolg-1

) represents the maximum sorption capacity, and β (mol2kJ

-2) is a

constant related to sorption energy.

ε is the Polanyi sorption potential calculated as below:

1ln(1 )

e

RTC

(2.9)

where R is the gas constant (8.314 Jmol-1

K-1

), T is the temperature in K and Ce

is the metal concentration at equilibrium. β and Qm can be calculated from the

slope and intercept of lnQe versus .

The mean free energy (E, kJmol-1

) required to transfer one mole of metal ions

from the infinity of a solution to the surface of biomass can be determined as

below:

12( 2 )E

(2.10)

If 8 kJmol-1

< E < 16 kJmol

-1, the adsorption takes place by chemical ion

exchange; if E < 8 kJmol-1

, the adsorption takes place by physical process.

2

0

Chapter 2

39

2.2.5.4 Thermodynamic study

The Gibbs free energy, the enthalpy and the entropy for the biosorption

process were computed using the following equations [94]:

0 lnG RT b (2.11)

0 0

lnS H

bR RT

(2.12)

where ΔG0

(kJ/mol) is the standard free energy of change, R is the gas constant

(8.314 Jmol-1

K-1

), T is the temperature in K, and b is the Langmuir constant.

The entropy (ΔS0) and enthalpy (ΔH

0) changes can be computed from a plot of

lnb versus 1/T.

2.2.5.5 Desorption and reusability studies

For desorption study, biosorption experiment was first carried out by

contacting the dried biomass with metal solution at pH 4.5 with a solid-to-

liquid (S/L) ratio of 3.75 gL-1

. Cd(II) and Hg(II) loaded biomass samples were

then contacted with the desorption solutions for 1 h and 2 h, respectively.

After the biosorption experiment, the biomass was then dried in an oven at

105 °C and stored in a desiccator until desorption tests were carried out. Each

desorption solution was adjusted to a molarity of 0.2 M and S/L is maintained

at 10 gL-1

in each desorption test. All the tests were done at 20°C. Desorption

efficiency of each solution was computed as follow [98]:

x 100% (2.13)

Chapter 2

40

where Ed is the desorption efficiency (%), md (mgg-1

) is the amount of metal

ions released to the bulk solution after desorption and mb (mgg-1

) is the

amount of metal ions adsorbed onto the biomass after biosorption. Reusability

of the biomass was evaluated by repeating the biosorption-desorption cycle for

five times using the same biomass.

2.3 Results and Discussions

2.3.1 Characterization of POME biosorbent

FTIR technique was used to examine the surface groups of POME biosorbent

and to evaluate the functional groups responsible for the metal adsorption.

Infrared spectra of the biosorbent and metal-loaded biosorbent samples are

shown in Figure 2.1. The spectra display a number of absorption peaks which

indicate the complex nature of the POME biomass. The band at 3413 cm-1

is

the stretching vibrations of surface hydroxyl groups, the bands at 2922 cm-1

and 2918 cm-1

are assigned to the symmetric vibrations of CH2 especially

alkenes, and the band at 2361 cm-1

can be associated with the stretching

vibration of O-H from strong hydrogen-bonded –COOH group[94, 99]. The

band at 1743 cm-1

corresponds to the C=O in carbonyls that may be attributed

to the lignin aromatic groups [100], the bands at 1647 and 1508 cm-1

are more

likely due to the presences of COO-, C=O groups in cellulose, the band at

1375 cm-1

reflects the stretching vibrations of the ionic carboxylic groups (-

COOH) present in pectin, the band at 1246 cm-1

is assigned to the stretching

vibration of –SO3 of hemicelluloses, the band at 1050 cm-1

is probably due to

Chapter 2

41

the –OCH3 group present in lignin structure, and the band at 559 cm-1

shows

the characteristic band of –SH2 and –PO4 functional group [94, 101].

Figure 2.1 Infrared spectra of (a) raw POME, (b) NaOH modified-POME,

(c) Hg(II) loaded POME, (d) Cd(II) loaded POME.

As shown in the FTIR spectrum, POME consists of mainly polymeric OH,

CH2 and COO- groups. All these functional groups were known to have

certain affinity to heavy metals [102]. As shown in Figure 2.1, the amount of

polymeric OH functional groups was increased after NaOH modification. This

suggests that the OH functional groups may play an important role for the

removal of metal ions. The disappearance of peaks at 2852, 1743, 1246 cm-1

may indicate the dissolution of some components of lignin and hemicellulose

by NaOH. It was observed that the peak at 2361 cm-1

disappeared after Cd(II)

and Hg(II) adsorptions, which indicates that O-H functional group in the -

COOH may be responsible for the biosorption process.

Chapter 2

42

SEM images of the raw POME biosorbent without contact with metal solution

and after contact with Cd(II) and Hg(II) solutions at pH 4.5 are shown in

Figure 2.2. The image in Figure 2.2a shows that POME biomass is made up of

several thin flakes that are loosely bound together. After the metal uptake, the

biomass surface appears littered with nodules of an electron dense material as

shown in Figure 2.2b and 2.2c. The EDX of these nodules show the presences

of Cd and Hg that were most probably adsorbed on the biomass surface

(Figure 2.2d and e). This suggests that surface precipitation may be involved

in the metal sorption process.

Figure 2.2 SEM micrographs of POME biomass for (a) raw powder, (b)

Cd loaded, (c) Hg loaded, (d) EDX analysis on Cd loaded biomass, (e) EDX

analysis on Hg loaded biomass.

2.3.2 Optimization of Pretreatment Conditions

In order to investigate the effects of NaOH concentration and treatment time,

OA9 (34) matrix combined with OA8 (2

7) matrix were applied. In OA9 (3

4)

design, the interactions between variables were not considered. Due to the

Chapter 2

43

clearer trend of responses in three-level design, the aim of using this three-

level design is to locate the region of optimum responses (Figure 2.3a and b).

Figure 2.3 Effects of the variables on the response for metal removal

compared with untreated biomass ( 0 M NaOH, 0 min): (a) Cd removal, (b)

Hg removal. ▲NaOH (Level 1: 0 M, Level 2: 0.2 M, Level 3: 0.6 M, Level 4:

1 M); ■ treatment time (Level 1: 60 min, Level 2: 40 min; Level 3: 20 min;

Level 4: 0 min)

The optimum regions were selected from 0.2 M to 0.6 M for NaOH

concentration and from 20 min to 40 min for treatment time. After that, an

OA8 (27) matrix was used for further optimization, in which more exact levels

of the variables were chosen and the interaction effects between the variables

were considered separately. The results were then analyzed by using analysis

of variance (ANOVA) and F-test as shown in Table 2.2a and b.

Table 2.2a ANOVA table for OA8 (27) matrix with Cd removal as the output

responses.

Source of

variances

Sum of

squares

Degrees of

freedom

Mean

square F value Significance

a

Percent

contribution

[NaOH] 315.85 1.00 315.85 462.91 P < 0.001 25.24

time 612.21 1.00 612.21 897.25 P < 0.001 48.92

A X B 320.70 1.00 320.70 470.02 P < 0.001 25.63

Error b 2.73 4.00 0.68

0.22

Total 1251.50 7.00

100.00

a The critical F value is 61 at p < 0.001, 18 at p < 0.01 and 4.32 at p < 0.1.

b Pooled error results from pooling unassigned column effects.

Chapter 2

44

Table 2.2b ANOVA table for OA8 (27) matrix with Hg removal as the output

responses.

Source of

variances

Sum of

squares

Degrees of

freedom

Mean

square F value Significance

a

Percent

contribution

[NaOH] 21.08 1.00 21.08 15.75 P < 0.01 9.86

time 161.91 1.00 161.91 121.01 P < 0.001 75.78

A X B 25.33 1.00 25.33 18.93 P < 0.01 11.85

Error b 5.35 4.00 1.34

2.50

Total 213.67 7.00

100.00

a The critical F value is 61 at p < 0.001, 18 at p < 0.01 and 4.32 at p < 0.1.

b Pooled error results from pooling unassigned column effects.

The results from ANOVA analysis show that NaOH concentration and

pretreatment time as well as their interactions have exerted significant effects

on the metal removal at p < 0.001 and p < 0.01. By simply using 17

experimental trials, the optimum condition was determined to be 0.2 M NaOH

with 40 min treatment time (run 1 and 2 in OA8 (27) matrix) for both the

removal of Cd(II) and Hg(II) The concentration of NaOH used during

pretreatment is critical because it can solubilise hemicelluloses and pectin that

embedded in the cell wall of POME, which may expose the active binding

sites to the metal ions. Besides, the solubilisation of chemical oxygen demand

(COD) was also improved after NaOH treatment, which may attributed to the

solubilisation the surface lipids by removing base-soluble tannin that may

stain the water and hence increase COD. It was observed that COD increases

to 315 mg/L after unmodified POME biomass was contacted with metal

solution, whereas only 25 mg/L of COD was found after metal uptake by

MPOME. In addition, the efficiency of MPOME was tested in real wastewater

matrix (made up of surface water collected at 4 meters deep in the Kranji

water reservoir, Singapore) spiked with the target metal ions as well.

Chapter 2

45

Approximately 20% drop in removal efficiencies were observed for Cd(II)

( from 97 to 78%) and Hg(II) ( from 98 to 80%) spiked wastewaters with

concentration of 50 mg/L and 15 mg/L respectively.

2.3.3 Study of Biosorption Kinetics

It is important to study the biosorption kinetics in the treatment of aqueous

effluents because they provide useful information on the mechanism of the

adsorption process. In order to investigate the biosorption kinetics of Cd(II)

and Hg(II) ions by POME biosorbent, three diffusion kinetic models were

tested, as shown in Figure 2.4a to 2.4d. The kinetic parameters for the

adsorption of Cd(II) and Hg(II) on the NaOH modified-POME biosorbent are

listed in Table 2.3.

Figure 2.4 (a) Effect of contact time on the metal uptake (mgg-1

) of Cd(II)

and Hg(II); (b) Plot of linearized pseudo-second-order kinetic models; (c) Plot

of intra-particle diffusion-controlled kinetic models; (d) Plot of the external

mass diffusion-controlled kinetic models. ●: Cd(II), ▲: Hg(II)

Chapter 2

46

When the biomass and metal ions are in contact, there is a finite distribution of

metal ions between the solid and liquid phases. The minimum contact time

between adsorbent and adsorbate to reach this equilibrium state were

determined to be 1 h for Cd(II) and 2 h for Hg(II) as shown in Figure 2.4a.

The curve was characterized by a strong increase during the first few minutes

of contact, followed by a slow increase until the equilibrium state was reached.

The experimental data were then regressed against the pseudo-second-order

kinetic model based on Equation 2.2. The equilibrium capacity (qe ) and the

rate constant (k) can be calculated from the slope and intercept of the curve.

According to Figure 2.4b, the kinetic results fit very well to the pseudo-

second-order kinetic model for the Cd(II) and Hg(II) ions using POME as

biosorbent with regression coefficient, R2, between 0.998 and 1.000 (Table

2.3). In addition, the calculated and experimental qe values were very close to

each other. As a result, the biosorption systems for both Cd(II) and Hg(II)

appear to follow the pseudo-second-order kinetic model, which describes

chemisorptions that involve the sharing or exchange of electrons between

Table 2.3 Kinetic parameters for the adsorption of Cd(II) and Hg(II) ions on the NaOH modified-

POME biosorbent (MPOME dose: 3.75 gL-1

, temperature: 298 K, pH: 4.5, agitation speed: 300 rpm,

initial Cd(II) concentration: 50 mgL-1

, Hg(II) concentration: 50 mgL-1

).

Pseudo-second-order kinetic model Intra-particle diffusion-controlled

kinetic model

External mass transfer-

controlled kinetic model

qe exp

(mgg-1

)

qe cal

(mgg-1

)

k

(gmg-1

min-1

)

R2

kd

(mgg-1

min-0.5

)

y

R2

kf

(cms-1

)

R2

Cd(II) 11.181 11.211 0.261 1.000 0.0647 10.747 0.9599 0.0431 0.9713

Hg(II) 11.330 11.416 0.027 0.998 0.3435 7.790 0.9873 0.0295 0.9665

Chapter 2

47

adsorbent and adsorbate through the formation of covalent bonds or ion

exchange [103]. k values were calculated to be 0.261 and 0.027 gmg-1

min-1

for

Cd(II) and Hg(II) sorptions which show that Cd(II) ions have a faster rate of

adsorption as compared to Hg(II) ions (Table 2.3). This could also explain the

shorter equilibrium time required by Cd(II) as compared to Hg(II).

When metal ions are in contact with the biosorbent, transport of the metal ions

from the solution through the interface between the solution and the adsorbent

and into the particle pores may become effective [104]. One of these steps

may control the rate at which solute is adsorbed. Intra-particle diffusion and

film diffusion models were examined to verify the influence of mass transfer

resistance on the adsorptions of Cd(II) and Hg(II) to the POME biosorbent

based on Equation 2.3 and 2.4. The intra-particle diffusion constant, kd (mgg-

1min

-0.5), can be obtained from the slope of the plot of qt (mgg

-1) versus t

0.5

(min0.5

). The intercept of the plot (y) indicates the boundary layer effect. This

might be related to the external mass transfer (film diffusion) effect on the

adsorption process. The film diffusion rate constant, kf (s-1

), can be obtained

from the slope of the plot of ln(1-F) versus t (min). As shown in Figure 2.4c

and 2.4d, the best-fit straight lines of these two models do not pass through the

origin for both the adsorptions of Cd(II) and Hg(II). This suggests that both

intra-particle diffusion and film diffusion may be involved in the biosorption

process but they are not the rate-determining steps. In fact, the kinetics of

interaction between Cd(II) and Hg(II) with the POME biosorbent should not

be overwhelmingly controlled by only one mechanism. Although pseudo-

second-order mechanism is more likely the dominating mechanism,

Chapter 2

48

contributions from intra-particle and film diffusion could not be ruled out

completely as well.

2.3.4 Study of Biosorption Isotherms

The study of biosorption isotherm provides us the relationship between the

amount of adsorbate taken up by the adsorbent (qe) and the adsorbate

concentration remaining in the solution after the system attained equilibrium

(Ce). According to the regulations of National Environment Agency of

Singapore (NEA), the maximum allowable concentrations for trade effluent

discharged into public sewer are 1 mgL-1

for cadmium and 0.5 mgL-1

for

mercury. In order to evaluate the performance of POME for the removal of

Cd(II) and Hg(II) in a more practical way, test solutions with concentrations

ranges from 1 to 100 mgL-1

were used for the study of different adsorption

isotherms.

For the adsorption of Cd(II) and Hg(II), the Langmuir, Freundlich and

Dubinin-Radushkevich isotherm models were investigated. The parameters of

these isotherms give useful information for the adsorption mechanism, the

surface properties and affinity of the adsorbent [105]. For Cd(II) adsorption,

Langmuir and Freundlich isotherms were tested based on Equation 2.5 and 2.7

and the results are shown in Table 2.4.

According to the R2 values, Cd(II) sorption shows a better fit to Langmuir

model as compared to Freundlich model, especially at higher temperatures.

The effect of metal ions concentration on adsorption capacity of NaOH

modified-POME biomass and the plot of linearized Langmuir isotherms for

both Cd(II) and Hg(II) are shown in Figure 2.5(a) and (b). Langmuir model

Chapter 2

49

suggests that when an adsorbate occupies a site on the adsorbent, there is no

further sorption allowed at that site. All sites are energetically equivalent and

there is no interaction between adsorbates [106]. In this case, biosorption stops

at one monolayer, which is consistent with the specific adsorption onto

functional binding sites on the biomass. In addition, the exchange reaction

between surface active binding sites and previously adsorbed metal ions is

only a monolayer or less [107]. All the RL values calculated based on Equation

2.6 for Cd(II) and Hg(II) biosorption were between 0.001 to 0.4, which

indicate favourable adsorption. However, Freundlich isotherms are important

also because they do not assume a homogeneous surface.

Chapter 2

50

Table 2.4 Parameters for Langmuir, Freundlich, and D-R isotherms at different temperatures (MPOME dose: 3.75 gL-1

, temperature: 298 K,

pH: 4.5, agitation speed: 300 rpm, contact time: 1 h for Cd(II), 2 h for Hg(II)).

Cd(II) Hg (II)

T

(K) Langmuir isotherm

Freundlich isotherm

Langmuir isotherm

Dubinin-Radushkevich isotherm

b

(mgg-1

)

K

(Lmg-1

)

R2

1/n

Kf

(mgg-1

)

R2

b

(mgg-1

)

K

(Lmg-

1)

R2

Qm

(mmolg-1

)

β

( x 10-

6)

E

(kJ

mol-1

)

R2

283 20.877 1.7046 0.993 0.157 11.167 0.8754 19.230 3.347 0.9628 0.127 -0.005 10 0.9054

293 21.978 2.2525 0.9923 0.187 12.039 0.9954 9.921 4.175 0.9242 0.140 -0.006 9.13 0.9102

313 21.413 1.4918 0.9739 0.172 11.048 0.7863 11.481 2.165 0.9624 0.010 -0.6 0.91 0.8871

333 21.786 1.8142 0.9881 0.199 11.321 0.9462 2.354 2.515 0.9258 0.003 -1 0.70 0.6137

Chapter 2

51

For Hg(II) sorption, besides Langmuir isotherm, D-R model was examined as

well and the parameters are listed in Table 2.4. According to the sorption

mean free energies (E) calculated from D-R model, the adsorption processes at

283 and 293 K most probably took place through chemical ion exchange. The

fitting of this model decreased significantly when the temperatures were

increased to 313 and 333 K, in which physical adsorption may become

dominant. According to Langmuir model, the maximum adsorption capacity

for Cd(II) and Hg(II) are 22 mgg-1

at 293 K and 19.2 mgg-1

at 283 K,

respectively. This suggests that POME biosorbent is a fairly good biosorbent

for the removal of Cd(II) and Hg(II) from aqueous solutions at low metal

concentration. Due to the complex nature of POME biomass, the fitting of

experimental data with a certain isotherm is of limited utility and it cannot be

used to predict the extent of adsorption if the solution variables change.

However, such models are useful to assess the performance of different

biomass under various operating conditions, especially for the design and

optimization of adsorption system.

As suggested in previous studies, Cd2+

and HgCl2(aq) appear to be the

principal species that are involved during the biosorption processes [97, 108].

It was proposed that Hg(II) may be adsorbed on the biomass surface via a

mechanism of HgCl2 molecular adsorption. The effects of the amount of Cd2+

and HgCl2(aq) species on the removal efficiency of metal ions were evaluated

using a software (VISUAL MINTEQ). As shown in Figure 2.5c and 2.5d, the

removal efficiency of Cd(II) decreases slightly when the initial metal

concentration increases. This trend could be explained by the availability of

less binding sites when the amount of metal ions competes for the binding

Chapter 2

52

sites increased. It was observed that this trend also corresponds to the gradual

decrease of the amount of Cd2+

species in more concentrated solutions. For the

case of Hg(II), the removal efficiency changes with the trend of HgCl2(aq)

species when the initial metal concentration increases up to 25 mgL-1

. Above

this concentration, the sorption capacity was most probably limited by other

factors that may become dominant. Hence, it was deduced that metal

biosorption is a function of free-metal activity rather than total metal

concentration.

Figure 2.5 (a) Effect of metal ions concentration on adsorption

capacity of NaOH modified- POME biomass; (b) Plot of linearized Langmuir

isotherms; ●: Cd(II), ▲: Hg(II); (c) Variation of Cd2+

species with increasing

initial Cd concentration; (d) Variation of HgCl2 (aq) species with increasing

initial Hg concentration (The percent of metal species was calculated by

Visual MINTEQ).

Chapter 2

53

2.3.5 Effect of Ionic Strength

Wastewater originating from different industries normally contains a certain

amount of electrolytes with a variety of mineral salts. The presence of these

salts or co-ions in solutions leads to high ionic strength that may interfere with

the sorption capacity of biosorbents. Thus, the effect of ionic strength on the

uptake of Cd(II) and Hg(II) by MPOME biosorbent was examined with 0.001,

0.01, and 0.1 M NaCl solutions at pH 4.5. As shown in Figure 2.6, the

variation of ionic strength had a significant effect on the extent of metals

uptake by biosorbent. This phenomenon can be attributed to the increasing

Na+ ions that may compete with the metal ions for the adsorption sites on

biomass surface. In addition, it was observed that the trend of metal sorption

was influenced by the percentage of Cd2+

and HgCl2 (aq) species as well,

which were affected by the ionic strength of the solutions.

Figure 2.6 (a) Effect of ionic strength on adsorption capacity of NaOH

modified-POME biomass for Cd removal; (b) Effect of ionic strength on

adsorption capacity of NaOH modified-POME biomass for Hg removal (The

variation of metal species was calculated by VISUAL MINTEQ).

Chapter 2

54

2.3.6 Effect of Adsorbent Dosage

The study of biosorbent dosages for the removal of Cd(II) and Hg(II) from

aqueous solution was carried out using biosorbent dosages ranging from 1.0 to

5.0 gL-1

and fixing the initial Cd(II) and Hg(II) concentration at 100 mgL-1

and

17 mgL-1

, respectively. For both metals, it was observed that the amount of

metal removal reached a plateau at biosorbent dosages of 3.75 gL-1

. For

biosorbent dosages higher than this value, the removal of the metals remained

almost constant. As shown in Figure 2.7a and 2.7b, an increase of the

biosorbent dosage was followed by an increase of metal removal efficiency

but a decrease of metal uptake per unit weight of biosorbent.

Figure 2.7 (a) Effect of biosorbent dosage on Cd removal; (b) Effect of

biosorbent dosage on Hg removal (The variation of metal species was

calculated by VISUAL MINTEQ). ■ Cd or Hg removal (%); ▲ Cd or Hg

uptake (mg/g)

The increase of metal removal efficiency with biosorbent dosages could be

attributed to increases in the biosorbent surface areas, augmenting the number

of adsorption sites available for adsorption. On the other hand, the increase in

the biosorbent dosages promotes a remarkable decrease in the amount of metal

uptake per unit weight of biosorbent (q). This could be attributed to the

Chapter 2

55

availability of less metal ions per unit mass of adsorbent with some of the

active sites remaining unsaturated during the biosorption process. Besides, the

formation of aggregates between the biomass particles at high dosage of

biosorbent may also reduce the availability of those active binding sites and

exert a practical limit on the extent of metal uptake.

2.3.7 Thermodynamic Study

Gibbs free energy change (ΔG0) is the driving force that indicates the

spontaneity of biosorption process. Negative ΔG0 indicates the adsorption

process occurs spontaneously. The results of thermodynamic parameters are

shown in Table 2.5. The negative values for ΔG0

in both the adsorption of

Cd(II) and Hg(II) indicate that the biosorption processes are spontaneous in

nature. For Cd(II) adsorption, the values for ΔH0 and ΔS

0 were found to be

positive, which indicates that the biosorption of Cd(II) was endothermic in

nature. The increase of randomness at the solid-liquid interface during

biosorption shows the affinity of biomass to Cd2+

ions. The increase of

biosorption capacity at increasing temperatures may also be due to the pores

enlargement or activation of binding sites on the biomass surface. On the other

hand, Hg(II) adsorption is quite different with the case of Cd(II). The negative

values of ΔH0 and ΔS

0 suggest that the biosorption mechanism is exothermic

and the decrease of randomness may be related to the molecular adsorption

mechanism.

Chapter 2

56

Table 2.5 Thermodynamic parameters for the adsorption of Cd(II) and Hg(II) by MPOME. (MPOME dose: 3.75 gL-1

, pH: 4.5, agitation

speed: 300 rpm, contact time: 1 h for Cd(II), 2 h for Hg(II)).

Cd(II) Hg(II)

T

(K)

lnb

ΔG

(kJmol-1

)

ΔS

(Jmol-1

K-1

)

ΔH

(kJmol-1

)

R2

T

(K)

lnb

ΔG

(kJmol-1

)

ΔS

(Jmol-1

K-1

)

ΔH

(kJmol-1

)

R2

283 3.04 -7.15

27.60 0.66 0.9976

283 2.96 -6.96

-95.00 -33.91 0.9991 313 3.06 -7.97 293 2.53 -6.16

333 3.08 -8.53 333 0.81 -2.25

Chapter 2

57

2.3.8 Desorption and Reusability Studies

Desorption tests were carried out by using different desorption solutions

following biosorption process. According to the results in Table 2.6, it was

observed that all the desorption solutions show good performance for the

recovery of Cd(II), with the desorption efficiency between 82 and 99%. As

compared to the recovery of Cd(II), good recovery for Hg(II) (> 95%) was

only achieved by using S,S-ethylenediamine-N,N’-disuccinic acid (EDDS) as

eluent. It was found in another study that HNO3 and HCl usually show good

performances (< 90%) during the first cycle of desorption but the metal uptake

capacity decreases in the subsequent biosorption/desorption cycles. As

suggested by the authors, it is most probably be due to the gradual destruction

of the functional groups and nitrification of the biomass surface [98, 109].

Hence, HNO3 and HCl are not suitable for regeneration purposes.

Ethylenediaminetetraacetic acid (EDTA) and EDDS are both powerful

chelating agents that are useful for extracting metals. However, the poor

biodegradability of EDTA made it a persistent organic pollutant in the

environment. As compared to EDTA, (S,S)-EDDS is an environmentally

benign alternative whose metal complexes are highly biodegradable. Thus, the

subsequent reusability study was focused on the performance of EDDS. The

results in Table 2.6 shows that EDDS is useful for the recovery of Cd(II) and

Hg(II) ions and the regeneration and recovery efficiencies remain high even

after five repeated cycles. The high regeneration efficiency may be attributed

to the formations of new binding sites on the biomass surface after elution by

EDDS. The powerful chelating ability of EDDS may help to increase the

Chapter 3

58

metal uptake capacity during the next biosorption. This is especially useful if

expensive toxic metals are to be removed and recovered quantitatively.

Chapter 2

59

Table 2.6 Performance of various desorption solutions on the recovery of metal ions and the effect of EDDS on the regeneration of

biosorbent during repeated adsorption/desorption cycles (initial Cd(II) concentration: 50 mgL-1

, initial Hg(II) concentration: 20 mgL-1

).

Desorption solution

(0.20M)

Desorption

efficiency (%)

A/D cycle

using

EDDS

Cd Hg

Cd Hg

MPOME

regeneration

efficiency (%)

a Recovery

efficiency (%)

b MPOME regeneration

efficiency (%)

Recovery

efficiency (%)

EDTA 82.61 16.29 1 > 99 81.82 > 99 95.31

EDDS 81.82 95.31 2 > 99 86.80 > 99 109.23

HCl 97.66 16.93 3 > 99 86.93 > 99 88.46

HNO3 99.43 1.03 4 > 99 85.30 92.05 89.44

5

85.98

95.15

a Recovery efficiency = amount of metal recovered/ amount of metal adsorbed x 100%

b MPOME regeneration efficiency = regenerated adsorption capacity/original adsorption capacity x 100%

Chapter 2

60

2.4 Concluding Remarks

Palm oil mill effluent which is an agricultural waste discharged from palm oil

industry was used as biosorbent for the removal of Cd(II) and Hg(II) from

aqueous solutions. The pretreatment of biosorbent, adsorption time, initial

metal concentration and ionic strength of metal solution were investigated.

According to the results of batch kinetic and equilibrium experiments, both the

adsorptions of Cd(II) and Hg(II) were best fitted to pseudo-second-order

kinetic model and Langmuir isotherm. The contact time required to reach

equilibrium was 1 and 2 h at 298K, for Cd(II) and Hg(II), respectively. The

optimal biosorbent dosage was found to be 3.75gL-1

for both the adsorptions of

Cd(II) and Hg(II). The role of metal speciation was investigated by using

VISUAL MINTEQ. Maximum removal efficiencies were found to be 97% for

50 mgL-1

Cd(II) and 98% for 15 mgL-1

Hg(II). Good recoveries of metal ions

were obtained by using EDDS as desorption agent.

Chapter 3

61

Chapter 3 Efficient Removal of Rhodamine B using a Rice

hull-based silica supported iron catalyst by Fenton-like process

3.1 Introduction

Many dyes are highly water-soluble. It is estimated that approximately 10-15%

of dyes produced annually (over 7 × 105 t) is lost during the dyeing process

and released as industrial effluents [36]. These colored effluents could affect

the photosynthetic activity in aquatic life through the hindrance of light

penetration. Hence, dyes have become one of the major environmental

pollutants in wastewater which render the subsequent recycling process a

challenging task. Conventional methods to treat the dye waste effluents

include flocculation, activated carbon adsorption, and biological treatment

[110]. Since many dyes are bio-recalcitrant, biological treatment is insufficient

to degrade many dye compounds [37]. Moreover, these methods usually do

not work efficiently as they are non-destructive, costly and merely involve the

transfer of pollutants from water to sludge which resulted in the production of

secondary waste that needs further disposal. Advanced oxidation processes

(AOPs) which are based on the generation of hydroxyl radicals (•OH) to

initiate non-selective oxidative destruction of organic pollutants have emerged

as a promising alternative. Among various AOPs, Fenton reagent (e.g. Fe2+

+

H2O2) and photocatalyst (e.g. TiO2 + UV) have been intensively studied for

environmental applications during the last decades due to their promising

performances in the degradation of many pollutants [38].

While there are many reports describing the use of advanced synthetic

materials such as carbon nanotube [47], graphene [111], zeolite [43, 48] as

Chapter 3

62

well as other nanocomposites [112-114] in the synthesis of solid catalyst for

Fenton and photocatalysis processes, serious doubts about their practical

feasibility are still harbored. In fact, the synthesis of these advanced materials

could be quite tedious in which the use of hazardous chemicals and harsh

reaction conditions are required. Considering the cost of production and

possible impact to the environment, the use of an environmentally friendly and

low cost material is important to make the Fenton treatment more affordable

as a general decontamination technology. Silica supported iron-based catalysts

have been actively researched owing to their large specific surface area and

excellent stability imparted by silica backbone [49].Unlike tetraethoxysilane

(TEOS) and tetramethoxysilane (TMOS) which are well-known precursors to

produce silica particles, rice hull which comprises about 20% of silica appears

to be an attractive natural source of silica [52, 53]. Due to its low economic

value and abundant availability, the use of rice hull as a feedstock of silica in

the synthesis of a heterogeneous Fenton-like catalyst for dye degradation

shows a great potential.

Rhodamine B (RhB) exhibits high stability under various pH with

considerably high resistance to photo and oxidative degradation [115, 116].

Toxic and carcinogenic effects upon exposure to RhB have been

experimentally proven [117, 118]. A number of approaches have been

reported in the literature for the degradation of RhB. These include Fenton-

based oxidation [119-121], photocatalytic degradation [117, 122-124],

sonochemical degradation [37, 125-127], ozonation [128-130], and biological

degradation [131]. Adsorptive removal of RhB from aqueous solution by

using rice hull-based adsorbents has been reported in several studies [117,

Chapter 3

63

132-134]. However, the use of rice hull-based catalyst for heterogeneous

Fenton-like degradation of textile dyes is rarely reported. In particular, Daud

and Hameed have reported on the use of a rice hull ash-based (RHA) catalyst

for decolorization of Acid Red 1 by Fenton-like process [135]. RHA is the

residue obtained from pyrolysis of rice hull at high temperature in which high-

energy consumption is involved. In this study, the catalytic performance of a

rice hull-based silica supported iron catalyst (RHSi-Fe) synthesized through a

low-energy method using the sodium silicate solution extracted from rice hull

was investigated.

Our preliminary study shows that efficient degradation of RhB could be

achieved by using silica loaded with iron content as low as 3 wt.%. As

suggested by Liou and Lin, prolonged aging time might result in the decrease

of surface area due to the dominant growth of silica particles instead of

nucleation [136]. Hence, an aging time of 2 h was used for the resulted silica

gel in this study. The use of low iron content and short aging time is

practically favorable due to the reduction in material consumption and

production time. The major objectives of this study are (i) to evaluate the

catalytic performance of RHSi-Fe for heterogeneous Fenton-like degradation

of RhB; (ii) to investigate the influences of various reaction parameters on the

rate of degradation; (iii) to improve the degree of mineralization without

additional dose of oxidant through the use of stepwise addition strategy; and

(iv) to evaluate the photocatalytic and sonocatalytic properties of RHSi-Fe by

applying ultrasonic (US) and ultraviolet (UV) irradiations in a comparative

study.

Chapter 3

64

3.2 Materials and Methods

3.2.1 Solutions and Reagents

RhB (90%) was procured from Sigma-Aldrich and used as received. H2O2 (30%

w/w) was obtained from Scharlau (Barcelona, Spain). Ferric nitrate,

Fe(NO3)3∙9H2O (98%) was purchased from Fluka (Buchs, Switzerland). Rice

hull waste was obtained through a one-time collection from a rice processing

mill located at Bangkok, Thailand. All the other chemicals were of analytical

grade and used without further purification. The RhB solutions were prepared

by using deionized (DI) water (resistivity ≈ 18.2 MΩ cm) obtained from a

Millipore (Billerica, MA) Direct Q purification unit. Figure 3.1 shows the

chemical structure of RhB and the UV-vis absorption spectra of RhB solution.

Figure 3.1 Chemical structure of RhB and the UV-vis absorption spectra of

RhB solution.

3.2.2 Extraction of Sodium Silicate Solution from Raw Rice Hulls

Sodium silicate solution was prepared from raw rice hull through a modified

alkali extraction method [137]. In a typical procedure, 35 g of dried rice hull

Chapter 3

65

was acid washed with 500 mL 1.0 mol/L HNO3 under stirring for 48 h to

remove the metal impurities. The rice hull was filtered and washed thoroughly

with DI water until the pH became almost neutral. The acid treated rice hull

residue was then dried at 383 K in an oven. The dried residue was stirred in

500 mL of 1.0 M NaOH for 24 h and the resulting dark brown suspension was

centrifuged at 4000 rpm for 20 min to obtain the sodium silicate solution.

3.2.3 Synthesis of RHSi-Fe Catalyst

For the synthesis of RHSi-Fe catalyst, 3.0 M HNO3 containing appropriate

mass of Fe(NO3)3∙9H2O to obtain 3 wt.% of Fe3+

was added dropwise into the

sodium silicate solution under stirring. The dropwise addition was continued

until the suspension become pH 3.0. The resulting gel solution was aged in the

mother liquor for 2 h at room temperature without disturbing. The gel was

recovered by centrifugation at 4000 rpm for 10 min followed by washing with

DI water under sonication for another 10 min. To wash the gel thoroughly, the

steps of centrifugation and sonication were repeated 3 times. The gel after

washing was dried in an oven at 383 K for 24 h to obtain RHSi-Fe.

3.2.4 Characterization of RHSi-Fe Catalyst

In order to analyze the crystallinity of RHSi-Fe, X-ray powder diffraction

(XRD) was carried out using Bruker-AXS Smart Apex CCD single-crystal

diffractometer (Karlsruhe, Germany). The X-ray source was CuKα radiation (λ

= 0.1542 nm). The diffractogram was recorded in 10-80° 2θ range, with a

0.025° step size and a collecting time of 1s per point. The iron content of

catalyst was determined using an Agilent 7500cx inductively coupled plasma

mass spectrometer (ICP-MS) (Santa Clara, CA).

Chapter 3

66

The Fourier transform infrared (FTIR) spectroscopy was conducted to study

the surface functional groups of the RHSi-Fe. FTIR spectra were obtained on a

Shimadzu (Kyoto, Japan) IR Prestige-21 operated in transmission mode (400-

4000 cm-1

) at a resolution of 4.0 cm-1

with the sample as KBr pellets.

Specific Brunauer-Emmett-Teller (BET) surface area was calculated from the

N2 adsorption isotherms measured at -196 °C using a Micromeritics ASAP

2020 (Norcross, GA). The morphology of RHSi-Fe was analyzed using high

resolution images obtained with a JEOL (Tokyo, Japan) 3010 transmission

electron microscope (TEM).

3.2.5 Heterogeneous Fenton-like degradation of RhB

Typical reaction mixture contained 50 mg of RHFe-Si in 50 mL of 5 mg/L

RhB solution unless otherwise stated. Before addition of H2O2, the dye and

catalyst suspension was magnetically stirred at 500 rpm for 1 h at room

temperature to allow the establishment of adsorption/desorption equilibrium.

After this pretreatment time, 2 mL aliquot was withdrawn to determine the

concentration C0. The oxidant (0.98 mmol) of H2O2 was added to the

suspension in a thermostat-controlled water bath after the temperature

stabilized at 323 K [138]. This time was recorded as the starting time of the

reaction. At certain reaction intervals, 2 mL aliquot was withdrawn and the

concentration of RhB (Ct) was analyzed immediately to avoid errors due to

further reactions. For the determination of RhB concentration, all the aliquots

withdrawn were centrifuged at 14000 rpm for 5 min followed by subsequent

UV-vis analysis of the supernatant solutions. The UV-vis spectrum of RhB

solutions were recorded from 190 to 700 nm using Hach (Loveland, CO) DR

Chapter 3

67

5000 UV-vis spectrophotometer. The RhB concentration was determined

based on the constructed calibration curves at maximum absorption

wavelength (λmax) of 554 nm. Each experiment was performed in triplicates.

All results were expressed as a mean value of 3 experiments with a standard

deviation of ≤ 5%.

3.2.5.1 Chemical Oxygen Demand (COD) Measurements

Chemical oxygen demand (COD) measurements were carried out by using a

USEPA-approved dichromate method (Hach Method 8000). As H2O2 disturbs

the COD measurement, catalase (from bovine liver cell culture, Sigma-Aldrich)

was used to destroy residual H2O2 before measurement. The catalase

contribution to the COD results was deducted from all presented values [139].

3.2.5.2 Characterization of Degraded Products by FTIR Analysis

The degraded products were characterized by FTIR analysis. The suspensions

after reaction were filtered to remove the RHSi-Fe particles and the filtrates

were evaporated by freeze-drying method. For FTIR measurement, the dry

residues were supported on KBr pellets at a fixed weight ratio (1%).

3.2.6 Sonochemical and Photocatalytic Experiments

The catalytic activity of RHSi-Fe under US and UV irradiations was evaluated.

Sonochemical experiments were conducted in a Cole-Parmer 8890 ultrasonic

cleaning bath operated at 42 kHz. Temperature of the ultrasonic bath was

controlled at 323 K during the reaction. Photocatalytic experiments were

conducted using a Blak-Ray B-100AP/R UV lamp (365 nm, 100 W) in a self-

made black wooden box at room temperature. Distance between the source of

Chapter 3

68

UV light and test solution was about 30 cm. No temperature control was

applied during the course of reaction.

3.2.7 Stepwise Addition Strategy Study

A series of batch experiments were performed to determine the degradation of

RhB under different stepwise dosing modes. The same experimental condition

as described in the previous section was used except the initial concentration

of RhB was selected at 100 mg/L. In each run, the total amount of H2O2 was

maintained at 0.98 mmol for 50 mL of test solution. For purpose of

comparison, punctual addition that dosed all of the H2O2 at the beginning of

reaction was conducted as well. Efficiency of COD removal and rate of

decolorization were measured to evaluate the performance of each dosing

strategy.

3.3 Results and Discussions

3.3.1 Characterization of RHSi-Fe Catalyst

The result from ICP-MS analysis reveals that iron content of RHSi-Fe catalyst

was about 3.18 wt.%. TEM micrograph in Figure 3.2a and b shows that the

RHSi-Fe consists of aggregated spherical particles with diameters ranging

between 20 and 30 nm. No sharp peak was observed in the XRD-

diffractogram (Figure 3.2c). This indicates a result of good dispersion of

microcrystalline iron particles on the amorphous silica support in which the

microcrystalline particles are too small to diffract X-rays. BET surface area of

RHSi-Fe was measured to be 422 m²/g, which is much higher than the values

Chapter 3

69

reported in reference [140]. The higher surface area obtained might be a result

of shorter aging time and lower iron content that were used in this study.

Figure 3.2 TEM images and XRD-diffractogram of RHSi-Fe.

3.3.2 Degradation Characteristic of RhB using RHSi-Fe

The residual concentration of RhB left in the solution during its degradation in

the RHSi-Fe/H2O2 system was monitored by measuring its absorption at 554

nm. The temporal evolution of the spectral changes is presented in Figure 3.3a.

Chapter 3

70

Figure 3.3 The temporal UV-vis spectra of RhB during its degradation in

RHSi-Fe/H2O2 system (a). Degradation of RhB as a function of reaction time

under heterogeneous and homogeneous conditions (b). Reaction conditions:

initial concentration of RhB = 5 mg/L, temperature = 323 K, catalyst dosage =

1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0

According to the results, the main absorbance at 554 nm decreased with

increasing reaction time and disappeared after 40 min, which indicates that the

structure of RhB molecules was destroyed during the reaction. To verify the

catalytic efficiency of RHSi-Fe, control experiments were conducted by using

reaction mixtures containing RhB alone, RhB in the presence of RHSi-Fe and

RhB in the presence of H2O2 under the same experimental conditions (Figure

3.3b). The results show that degradation of RhB under natural light at our

experimental conditions is negligible. No noticeable decolorization was

observed in the presence of either H2O2 or RHSi-Fe alone. The degradation of

RhB in the presence of both RHSi-Fe and H2O2 reached almost 100% within

40 min. This observation suggests that RHSi-Fe is efficient to catalyze the

oxidative degradation of RhB, which resulted in the decolorization of RhB

solution. On the other hand, it was noticed that the degradation curve of RhB

in RHSi-Fe/H2O2 system shows a sigmoidal profile, which is typical for

autocatalytic or radical reactions [141]. Basically, two regions can be

identified, the initial one representing an induction period and the second one

Chapter 3

71

after the inflection point representing the steady state. This indicates that a

certain minimum number of free radicals are required for the onset of the

degradation process [126]. The presence of induction period during

heterogeneous dye degradation was also observed in other studies [126, 135,

138, 142].

To further investigate the catalytic activity of RHSi-Fe catalyst, leaching tests

were performed by measuring the amount of iron leached into the solution

after a reaction time of 120 min. The results show that about 5.0 mg/L of Fe3+

were leached into the solution. Since iron leaching was detected during the

reaction, homogeneous contribution of leached Fe3+

to total activity was

examined. A homogeneous Fenton process was carried out using the same

concentration of Fe3+

(prepared from Fe(NO3)3∙9H2O) that were leached from

RHSi-Fe under the same operating conditions. As shown in Figure 3.3b, the

degradation of RhB is very slow in which only about 17% of decolorization

was observed after 40 min of reaction. This suggests that although there is a

homogeneous catalytic contribution resulted from iron leaching, the fast

degradation of RhB observed within 40 min in the presence of RHSi-Fe is

mainly accelerated by a heterogeneous process.

Figure 3.4 shows the FTIR spectra changes of RhB during its degradation in

RHSi-Fe/H2O2 system with increasing reaction time. The curve of untreated

RhB consists of peaks at 1591 and 1475 cm-1

which were attributed to the

stretching vibrations of C C and CH2 bond in the aromatic ring, while the

peaks at 1342 cm-1 corresponds to C N stretching in the structure of hB

molecule [143]. During the process of Fenton-like degradation, these

Chapter 3

72

characteristic peaks disappeared after a reaction time of only 20 min. The new

peaks observed at 1633, 1382 and 1159 cm-1 were assigned to the stretching

vibrations of N O bond, N O and N H bond, and C O bond, respectively

[144]. In addition, a wide absorption peak around 3460 cm-1, which is due to

the stretching vibrations of O H became apparently stronger during the

degradation process. This results indicate that the structure of RhB had been

destroyed and some hydroxylated adducts were generated.

Figure 3.4 The changes of FTIR spectra during the degradation RhB in

RHSi-Fe/H2O2 system.

Similar FTIR spectra for another basic dye with similar structure, methylene

blue, after 40 min reaction with a total of 37 mmol H2O2 in layered birnessite-

type manganese oxides suspension was also observed in reference [144]. As

compared to 0.98 mmol of H2O2 used in this study, it further implies the

efficient use of H2O2 achieved by RHSi-Fe in the degradation of dye with

similar structure. While the reaction proceeded up to 120 min, the peak

intensities decreased and shifted to 1679, 1396, and 808 cm-1

, respectively.

The more intense peak observed at 1679 cm-1

could be due to the stretching

Chapter 3

73

vibrations of C=O bond whereas the two small peaks observed at 1396, and

808 cm-1

might be attributed to C-N stretching and N-H bending. The FTIR

spectra changes suggest that large conjugated chromophore structure of RhB

was destroyed with the formation of some smaller carboxylated compounds.

Since complete decolorization of RhB does not mean that the dye is

completely oxidized, the degradation of dye in terms of COD removal was

investigated. The changes of RhB and COD concentration with respect to

reaction time in the RHSi-Fe/H2O2 system are depicted in Figure 3.5. As

shown in the figure, about 75% of COD was removed after a reaction time of

240 min. The reduction in COD values indicates the degree of mineralization

achieved. The results obtained show that RHSi-Fe/H2O2 is effective in the

destruction of RhB in which considerable degree of mineralization could be

achieved. As compared to the decolorization process (40 min), the rate of

COD removal was much slower. The longer reaction time required for COD

abatement as compared to the decolorization process has been reported in

several studies [110, 145, 146]. Thus, it can be assumed that during

heterogeneous Fenton-like degradation of RhB, the more easily degradable

color substrates undergo oxidation at the beginning and then more resistant

compounds are oxidized in a later stage. The latter substances most probably

cause a slower rate of reduction in COD in the treated RhB solution.

Chapter 3

74

Figure 3.5 The changes of COD and RhB concentration with respect to

reaction time are shown. Reaction conditions: initial concentration of RhB =

50 mg/L, temperature = 323 K, catalyst dosage = 1 g/dm3, H2O2 amount =

0.98 mmol, pH 5.0.

3.3.3 Effect of Radical Scavengers

Effects of radical scavengers on the rate of RhB decolorization were examined

to evaluate possible reactive oxygen species (ROS) that are responsible for the

degradation process. Sodium azide, p-benzoquinone and dimethyl sulfoxide

were selected as 1O2, O2

• and •OH radical scavengers, respectively [144, 147].

As shown in Figure 3.6, the decolorization of RhB was almost completely

depressed in the presence of sodium azide while the addition of dimethyl

sulfoxide has significantly slowed down the decolorization of RhB with less

than 40% of RhB decolorized within 40 min. Nevertheless, a slight increase of

the initial rate of RhB decolorization was observed in the presence of p-

benzoquinone.

Chapter 3

75

Figure 3.6 Effect of radical scavengers (0.1 M) on the degradation of RhB.

Reaction conditions: initial concentration of RhB = 5 mg/L, temperature = 323

K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0.

These results suggest that 1O2 and •OH radical are the main OS responsible

for the decolorization of RhB in the RHSi-Fe/H2O2 system. In fact, 1O2 is a

powerful electrophile that can react with highly conjugated organic dyes.

Although the oxidation capability of 1O2 is not as powerful as •OH radical, it

has a longer lifetime than •OH [148] and has been reported as major ROS in

the decolorization of several organic dyes in literature [144, 149]. In contrast,

O2•

is not important for the degradation of RhB in the RHSi-Fe/H2O2 system.

The slight increase of initial decolorization rate when p-benzoquinone was

added can be explained by the formation of p-benzohydroquinone in the

presence of H2O2 [150], which can quickly reduce Fe3+

to Fe2+

and accelerate

the Fenton’s reaction [151].

Based on the experimental data and a survey of literature [152-155], possible

reaction pathways for the degradation of RhB in RHSi-Fe/H2O2 system were

proposed as below:

Chapter 3

76

Fe3+

+ H2O2 → Fe2+

+ •OOH + H+ k1 = 0.001-0.01 M

-1s

-1 (3.1)

Fe2+

+ H2O2 → Fe3+

+ •OH + OH-

k2 ≈ 70 M-1

s-1

(3.2)

Organic dyes + •OH → oxidized products → CO2 + H2O

(3.3)

•OOH + •OOH → 1O2 + H2O2 k3 = 8.3 x 10

5 M

-1s

-1 (3.4)

Organic dyes + 1O2 → Oxidized products → CO2 + H2O (3.5)

Fe3+

+ •OOH → Fe2+

+ O2 + H+ k4 = 1.2 × 10

6 M

-1s

-1 (at pH 3)

(3.6)

3.3.4 Effect of RhB concentration

As suggested by several studies, the decolorization of RhB in water is best

described by either first [156, 157] or pseudo-first-order kinetics [125, 158].

To examine the reaction order for the degradation of RhB by RHSi-Fe, several

kinetic models were tested. It was found that the experimental data were best

fitted with the pseudo-first-order equation as given below:

Ct = Cexp(-kappt) (3.7)

where C0 and Ct are the initial concentration and the concentration at any time

t of RhB whereas kapp is the pseudo-first order apparent rate constant. The kapp

constants were obtained from the slopes of the straight lines by plotting –

ln(Ct/C0) as a function of time, t, through regression. The kinetic rate constants

kapp were determined at different RhB concentrations with corresponding

regression coefficients, R2, ranges from 0.94 to 0.97. The effect of initial RhB

concentration on the decolorization rate of RhB is shown in Figure 3.7.

Chapter 3

77

Figure 3.7 Effect of RhB concentration on the degradation of RhB. Inset

shows the variation of initial rate with respect to initial concentrations of H2O2,

(represented by [H2O2]). Reaction conditions: temperature = 323 K, catalyst

dosage = 1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0.

The results show that the rate of decolorization increases with increasing

concentration of RhB from 2.5 to 50 mg/L whereas further increase in the RhB

concentration to 100 mg/L results in a decrease of reaction rate. The ROS

formed near the surface of catalyst have very short lifetime (in the range of a

few nanoseconds to microseconds). By increasing the quantity of RhB

molecules per volume at higher concentration, there is a higher chance of

collision between dye molecules and the near surface active catalyzing species,

leading to an increase in the degradation efficiency. However, when the RhB

concentration was further increased to 100 mg/L, the number of RhB

molecules could be more than the number of available reactive sites, whereby

the reaction was slowed down. In addition, the results also imply that RHSi-Fe

is efficient for the heterogeneous Fenton-like degradation of RhB in a wide

range of concentrations. This is particularly useful for urgent applications to

specific problems including emergency situations that require decontamination

of wastewater with high concentrations of dye.

Chapter 3

78

3.3.5 Effect of Oxidant Concentration

H2O2 as the simplest form of peroxides is an integral component of several

chemical oxidation technologies such as Fenton, photo-Fenton, UV-based

chemical oxidation etc. From the standpoint of green chemistry, the advantage

of using H2O2 arises from its high atom-efficiency and the generation of water

as the only by-product. Hence, it is often used as an oxidant to enhance the

rate of catalytic reactions in several AOPs [159].

The effect of initial H2O2 concentration on the decolorization rate of RhB is

shown in Figure 3.8. The results show that when H2O2 concentration increases

from 0.49 to 0.98 mmol, there is a great increase in the decolorization rate of

RhB. The dependence of reaction rate (kapp) on H2O2 concentration ([H2O2])

was evaluated by plotting lnkapp against ln[H2O2] (Inset of Figure 3.8). A cubic

relationship was obtained and the equation for this relationship is expressed as:

y = 0.4142x3 + 4.0861x

2 + 13.341x + 11.111 (R

2 = 1.00) (3.8)

Figure 3.8 Effect of initial H2O2 concentration on the degradation of RhB.

Inset shows the kinetic rate constants kapp determined at different H2O2

concentrations. Reaction conditions: initial concentration of RhB = 5 mg/L,

temperature = 323 K, catalyst dosage = 1 g/dm3, pH 5.0.

Chapter 3

79

Equation 3.8 suggests that the degradation of RhB was very susceptible to the

concentration of H2O2 at low dose (between 0.49 and 0.98 mmol). A marked

increase in decolorization rate may occur by increasing the dose of H2O2 in

small increment within this range. For example, if one increases the amount of

H2O2 from 0.50 to 0.55 mmol, according to Equation 3.8, the increment of kapp

was calculated to be 3.2 ×10-3

min-1

. On the other hand, if the amount of H2O2

was increased from 0.95 to 1.00 mmol, the increment of kapp was calculated to

be only 8×10-4

min-1

. One possible explanation could be the amount of ROS

generated at low dose of H2O2 is not enough and hence the oxidation rate

increases significantly with increasing dose of H2O2. By further increasing the

H2O2 concentration, it is expected that the decolorization rate would increase

accordingly due to the formation of more ROS. Nevertheless, the increase in

decolorization rate becomes less significant when the amount of H2O2 was

further increased from 0.98 to 1.96 and 2.94 mmol. This could be attributed to

the scavenging of •OH which occurs through the following reaction:

H2O2 + •OH → H2O + •OOH k5 = 3.3 × 107 M

-1s

-1 (3.9)

Therefore, the selection of optimum H2O2 concentration is important for

practical applications. This optimum H2O2 concentration appears to be around

0.98 mmol under the conditions used in this study. The slightly higher

oxidation rates obtained above this concentration is not preferred due to the

increasing scavenging effect of H2O2 and also the cost of H2O2.

Chapter 3

80

3.3.6 Effect of Catalyst Dosage

The effect of catalyst dosage on decolorization rate of RhB is illustrated in

Figure 3.9. The result indicates that higher decolorization efficiency was

achieved when the catalyst dosage increased from 0.5 to 1.0 g/dm3. This is

mainly attributed to the availability of more iron active sites that could

accelerate the production of ROS on the catalyst surface. The higher amount

of near surface ROS generated could then attack the sorbed pollutants and

resulted in an increased rate of reaction. A decline in decolorization rate was

observed when catalyst dosage was further increased from 1.0 to 2.0 g/dm3.

This could be attributed to the increasing formation of aggregates between

particles which resulted in a decrease of surface area [49, 160]. Thus, it is

more likely that the optimum catalyst dosage is 1.0 g/dm3

under our

experimental conditions.

Figure 3.9 Effect of catalyst dosage on the degradation of RhB. Reaction

conditions: initial concentration of RhB = 5 mg/L, temperature = 323 K, H2O2

amount = 0.98 mmol, pH 5.0.

Chapter 3

81

3.3.7 Effect of Temperature

Temperature is a critical parameter to the reaction rate, the product yield and

distribution. The kinetic rate constants kapp were determined at different

temperatures with corresponding R2 ranges from 0.94 to 0.97 (Equation 3.7).

The effect of temperature on the decolorization rate of RhB is shown in Figure

3.10a. It can be seen that temperature exerts a strong effect on the

decolorization rate of RhB. The decolorization rate increased with increasing

temperature with a significant rise in reaction rate at 323 K. This phenomenon

is expected due to the exponential dependence of the rate constant with the

reaction temperature as shown in the inset of Figure 3.10a. This is because

higher temperature could increase the reaction rate between H2O2 and the

catalyst, thus accelerating the generation of ROS. The Arrhenius plot is

presented in Figure 3.10b by plotting lnkapp against 1/T. From the slope of

Arrhenius plot, –Ea/R, where R is universal gas constant (8.314 kJ/mol K),

activation energy (Ea) was calculated to be 82.53 kJ/mol under our

experimental conditions.

Chapter 3

82

Figure 3.10 Effect of temperature on the degradation of RhB (a). Inset in

Figure 3.10(a) shows the kinetic rate constants kapp determined at different

reaction temperatures. The corresponding Arrhenius plot is shown in (b).

Reaction conditions: initial concentration of RhB = 5 mg/L, catalyst dosage =

1 g/dm3, H2O2 amount = 0.98 mmol, pH 5.0. The change of degradation rate

after pH adjustment at 323 and 303 K is shown in (c).

According to Figure 3.10a, it appears that the degradation of RhB at near room

temperature (303 K) is very slow. In order to evaluate the feasibility of using

RHSi-Fe catalyst for the degradation of RhB at room temperature, the solution

pH was adjusted with other experimental conditions remaining the same.

Figure 3.10c shows that when initial solution pH was adjusted from pH 5.0 to

3.0, the reaction rate could be greatly enhanced. Almost complete

decolorization of RhB could be achieved within 80 min at 303 K. The results

suggest that at more acidic pH, the degradation of RhB is possible to proceed

at room temperature despite a longer reaction time is required.

3.3.8 Effect of pH

Dye waste effluent is discharged at different pH. The ability of a Fenton-like

catalyst to work under different pH conditions is highly desirable. The effect

of initial pH on the degradation of RhB was studied at pH 3.0, 5.0 and 7.0 and

Chapter 3

83

the results are presented in Figure 3.11. As shown in the figure, RHSi-Fe is

able to achieve almost complete decolorization of RhB at the three tested pH

values. A maximum rate of degradation was observed at pH 3.0 within 10 min

with the other reaction parameters kept at optimum conditions obtained from

sections 3.4 to 3.7 (initial concentration of RhB = 5 mg/L, temperature = 323

K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98 mmol). The rate of

decolorization of RhB decreases with increasing pH. Similar observations

have also been reported for the degradation of other dyes [49, 135]. One

possible reason might be attributed to the tendency to form less reactive

hydroperoxyl radical (•OOH) rather than •OH radical at less acidic condition

(Equation 3.1). As the pH increases, the excess OH- ions might lead to the

formation of ferric hydroxide complexes which deactivate the catalyst and

slow down the reaction.

Since the degradation of RhB by RHSi-Fe is more likely a surface-catalyzed

reaction, the initial adsorption of RhB on the catalyst surface was investigated.

The removal of RhB at different initial pH was recorded during the 1 h

pretreatment time in the absence of H2O2 (Inset in Figure 3.11). It can be seen

that the amount of RhB adsorbed decreased from 40 to 30% when pH

increased from 3.0 to 7.0. As suggested by Lin and Gurol, •OH radicals are

quite reactive and could most probably react with the sorbed species before

being able to diffuse to the solution. Thus, the organic compound might be

oxidized in the sorbed state by near-surface •OH [161]. By correlating the

higher sorption rate of RhB to its faster degradation rate at lower pH, this

further implies the dominant role of surface-catalyzed mechanism. The initial

Chapter 3

84

attractive force that could bring the dye molecules toward the catalyst surface

is of relevance in controlling the rate of degradation.

In fact, the optimum pH for the removal of RhB has been reported to be about

3 by several authors. Guo et al. studied the effect of pH on RhB adsorption

using rice-husk-based carbon and reported a pH of 3.45 as the optimum pH

[132]. Mohammadi et al. utilized a palm shell derived activated carbon as

adsorbent for the removal of RhB and reported a pH of 3 as the optimum pH

[162]. The amount of dye adsorbed on the catalyst surface at different pH is

affected by the surface charge of the catalyst and the distribution of dye

species in solution. A change in solution pH results in the formation of

different ionic species and different surface charge of the catalyst. RhB exists

in both cationic and zwitterionic forms in polar solvents (Figure 3.11b). At

pH below 4, the RhB ions are in cationic and monomeric molecular form. The

fast degradation rate observed at pH 3.0 is due to the favorable adsorption of

RhB in cationic form to silica matrix where the net surface charge is negative.

The adsorbed dye species were then oxidized at the sorbed state.

At pH greater than 4, the transformation of the RhB species from cationic to

zwitterionic forms occurs by deprotonation of the carboxyl group in the

cationic form (RhB+) [132, 162]. The zwitterionic form of RhB in water may

increase the aggregation of RhB by forming dimers through electrostatic

interactions between the xanthenes and the carboxyl group of RhB monomers

[132, 163]. The aggregation of RhB hinders the adsorption of dye molecules

into the pore structures of catalyst and results in decreased amount of dye

species being adsorbed on the active sites of catalyst. In addition, the

Chapter 3

85

adsorbed RhB species in aggregated forms might not be effectively oxidized

during the catalytic process. Thus, these could be possible reasons for the

lower degradation rates observed at pH 5.0 and 7.0.

Figure 3.11 Effect of initial pH on the degradation of RhB. Inset shows the

amount of RhB adsorbed (%) at pH 3.0, 5.0 and 7.0 (a). Reaction conditions:

initial concentration of RhB = 5 mg/L, temperature = 323 K, catalyst dosage =

1 g/dm3, H2O2 amount = 0.98 mmol. Cationic and zwitterionic forms of RhB

(b).

Chapter 3

86

3.3.9 Effect of Mass Transfer Resistance

The apparent rate of a heterogeneous reaction is usually dominated by either

the rate of intrinsic reaction on the surface or the rate of diffusion of the

solutes to the surface. In order to estimate the effects of external and internal

diffusion resistances on the reaction rate, we have evaluated the reaction-

diffusion modulus (Thiele modulus, Ø) which is expressed as the ratio of the

reaction rate to the diffusion rate based on the following equation:

Ø = [k/(D/L2)]

0.5 (3.10)

k is the reaction rate constants (s-1

), D is the diffusion coefficient (cm2/s), and

L is the thickness of the stagnant liquid film or the pore length (cm). If Ø is

estimated to be < 0.5, the rate of diffusion is concluded to be faster than the

reaction rate whereas if Ø is estimated to be > 5, this suggests the existence of

a strong liquid diffusion resistance.

In this study, we considered the external mass transfer resistance as diffusion

through liquid film and the internal mass transfer resistance as pore diffusion.

The diffusion coefficient of the solutes in liquids is typically ~10-5

cm2/s and

the stagnant layer thickness of the liquid film can be estimated by the Film

theory as about 10-3

cm on the basis of a typical mass transfer coefficient (0.01

cm/s) in agitated vessels. The maximum k values measured for the degradation

rate of RhB at different initial concentrations of RhB and H2O2 is about 2.2 x

10-3

s-1

.Thus, Ø was estimated to be 0.015 for external mass transfer.

According to FTIR analysis, a band at 1633 cm-1

which corresponds to the

bending vibration of the trapped water molecules in the narrow pores of silica

Chapter 3

87

matrix was detected even the catalyst was dried overnight at 110°C before the

spectra were recorded. This suggests that the catalyst is most likely a high

moisture product made up of a network of interconnected pores in the silica

matrix. Thus, the effect of pore resistance was estimated. To estimate the rate

of internal diffusion, the pore length for spherical porous particles can be

taken as equal to one-third of the radius of the particles size. Hence, the pore

length for the catalyst particles can be estimated as 5 x 10-7

cm. The effective

diffusivity of the solutes in the pores generally is about 10-6

cm2/s. Therefore,

the Ø was estimated to be 2.34 x 10-5

for the internal mass transfer.

These values of Ø imply that the average rate of RhB degradation on the

RHSi-Fe catalyst surface is far slower than its diffusion rate to the surface

through either external film or internal pores. Therefore, the effects of external

and internal diffusion resistances on the reaction rate are negligible in this

study.

3.3.10 Effect of Foreign Salts and Ionic Strength

The presence of ionic species in water may affect the degradation process via

adsorption of the contaminants or reaction with ROS. This is an important

point that needs to be considered as industrial effluents usually contain a

certain quantity of inorganic salts with varying concentrations. As an example,

NaCl is used to promote the transfer of dyestuff to the fabric and it commonly

exists with Na2SO4 in textile mill effluents [164, 165].

The effect of foreign salts on RhB removal was investigated in this RHSi-

Fe/H2O2 system and the results obtained are depicted in Figure 3.12a. As

shown in the results, the degradation rate of hB (≈ 10-5

M) in 0.01 M of

Chapter 3

88

Na2SO4, NaCl, KCl, MgCl2 and CaCl2 declines in sequence, and the initial

degradation rate in Na2SO4 and NaCl solutions is greater than without adding

any salts deliberately (control sample). To further investigate the effect of

ionic strength on removal of RhB, solutions of varying ionic strengths (0.01,

0.05 and 0.1 M) adjusted with NaCl were tested (Figure 3.12b). The results

obtained show an increased rate of RhB removal as the ionic strength was

increased from 0.01 M to 0.05 M. However, no noticeable improvement was

observed when the ionic strength was further increased to 0.1 M.

Figure 3.12 Effect of various foreign salts on RhB degradation (a) and the

corresponding changes in degradation rate when the ionic strength of solutions

was adjusted with NaCl. Reaction conditions: initial concentration of RhB = 5

mg/L, temperature = 323 K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98

mmol, pH 5.0.

This observation seems contradictory to the common realization that chloride

and sulfate ions could inhibit the degradation process through radical

scavenging and complexation [166-168]. The ionic strength was reported as an

important parameter for the adsorption of dye on oxide surface because they

can influence electrostatic interactions between the oxide surface and the dye

Chapter 3

89

species [169]. As the degradation of RhB in this RHSi-Fe/H2O2 system mainly

occurred through a heterogeneous surface-catalyzed reaction, the reason for

this observed phenomena might be related to the presence of excess anions

that affected the equilibria between positively charged RhB species and the

catalyst surface [170]. At our experimental condition (pH 5.0), the net surface

charge of SiO2 is negative (pHpzc = 2.5), the anions are expected to have no

influence on the dye adsorption due to repulsive interactions. However, this

only applies if the net surface is considered homogeneously negatively

charged. In practice, if the heterogeneity of the surface is taken into

consideration, there might be a distribution of positive, negative and neutral

sites on the surface, in which the interactions between positive surface sites

and anions could take place. One possible explanation is the addition of anions

might allow the neutralization of the positive sites of Fe3+

and hence the

electrostatic repulsion barrier is hindered and non-electrostatic interactions

between cationic RhB and neutral site can occur through low energetic H-

bonds, or van der Waals short-ranged interactions. In particular, the better

degradation efficiency obtained in the presence of divalent sulfate anion could

be attributed to its higher valency in comparison with the monovalent chloride

ion, in which a stronger electrostatic field might be more efficient in reducing

the electrostatic repulsion barrier.

On the other hand, a deleterious effect on RhB removal was observed with the

increase of cations valences. Divalent cations such as Mg2+

and Ca2+

have

slowed down the degradation process to a larger extent than monovalent

cations such as K+ and Na

+. This implies that multivalent cations with stronger

electrostatic attraction to the negative silica surface are more efficient in

Chapter 3

90

competing with RhB molecule for the adsorption on catalyst surface. In

addition, the rate of degradation decreases as the hydrated ionic radii of the

added foreign cations having the same valence decreases, being K+ (3.31 Å) <

Na+ (3.58 Å); Ca

2+ (4.12 Å) < Mg

2+ (4.28 Å) [171]. Since the greater the ion’s

hydration, the farther it is from the adsorbing surface and the weaker its

adsorption, cations with greater hydrated ionic radii might be less efficient in

competing with the cationic RhB for the adsorption on catalyst surface and

hence result in a higher rate of RhB degradation.

In view of the above results, it shows that RHSi-Fe is effective in the presence

of foreign salts. In particular, common salts found in textile wastewater such

as Na2SO4 and NaCl could even enhance the degradation rate in this catalyst

system.

3.3.11 Effect of Stepwise Addition Strategy

As shown in Equation 3.9, high localized concentrations of H2O2 might lead to

undesirable reactions that consume •OH radicals and reduce the oxidative

capacity of the reactants over time. In order to maximize the performance of

the added H2O2, two series of experiments using even-dose (50 µL + 50 µL)

and asymmetrical-dose (75 µL + 25 µL) stepwise addition of H2O2 were

performed at different time intervals. The dosing strategy and comparison of

results for both dosing modes are illustrated in Table 3.1. When one-step

addition dosing mode was used, it was observed that dosing 100 µL (0.98

mmol) of H2O2 gives higher RhB and COD removal than dosing 50 µL (0.49

mmol) of H2O2. This indicates that fast depletion of ROS at lower dosage of

H2O2 could limit the extent of degradation process and hence gave a lower

Chapter 3

91

efficiency of COD removal. As two-step addition dosing mode was used, both

even- and asymmetrical-dosing of H2O2 showed negative effect on RhB

removal as compared to the one-step addition. In particular, when 50 µL

instead of 75 µL of H2O2 was used for the first dosing, the rate of

decolorization was slowed down more significantly, which implies that the

rate of decolorization depends largely on the initial amount of oxidant in the

solution.

However, it is noteworthy that at the dosing modes of 75 + 25 µL (T = 0 +

110 min), 50 + 50 µL (T = 0 + 110 min) and 75 + 25 µL (T = 0 + 60 min),

positive improvement of COD removal was observed with increasing order.

This suggests that both the dosing time and dosage of oxidants are important

in determining the efficiency of stepwise addition. A moderate distribution of

oxidant over time is favourable to improve the degree of mineralization in

which undesirable adverse reactions could be minimized. In addition, it was

noted that that the time interval between first and second dosing could not be

too long (as shown at T = 0 + 180 min). This might be due to the depletion of

H2O2 at the later stage of degradation process in which the oxidative reaction

might stop at less heavily oxidized products before the second dose of H2O2

was injected.

Chapter 3

92

Table 3.1 A comparison of the remaining RhB and COD in the solution at reaction time of the 50th and 240th min.

Dosing mode

(H2O2 addition)

Dosing time

(min)

(%)

(%)

Improvement (%) versus

One step addition

(%)

(%)

1st dosing 2

nd dosing

50 µL 100 µL 50 µL 100 µL

Two-step

addition

Even-dose

stepwise

50 µL H2O2 50 µL H2O2

0

15 12 51 40.0 0.0 19.1 -21.4

25 21 47 -5.0 -75.0 25.4 -11.9

60 18 46 10.0 -50.0 27.0 -9.5

110 18 38 20.0 -50.0 39.7 9.5

180 20 51 0.0 -66.7 19.1 -21.4

Asymmetrical-

dose stepwise

75 µL H2O2 25 µL H2O2

0

15 16 47 22.0 -30.0 25.4 -11.9

25 14 48 32.5 -12.5 23.8 -14.3

60 16 39 20.0 -33.3 38.1 7.1

110 14 36 30.0 -16.6 42.9 14.3

180 15 43 25.0 -25.0 31.8 -2.4

One-step

addition

50 µL H2O2 0 -- 20 63 -- -- -- --

100 µL H2O2 0 -- 12 42 -- -- -- --

Chapter 3

93

In light of the experimental results, stepwise addition strategy has been

demonstrated to be potentially useful in this RHSi-Fe/H2O2 system for

improving the efficiency of COD abatement. This is highly environmentally

and economically favourable as the degree of mineralization could be

improved by just using an optimum dosing mode without the requirement of

additional reagent.

3.3.12 Comparative Study with Ultrasound and Ultraviolet Irradiations

Investigations of various AOPs combination methods have been reported in

literature to improve the destruction of recalcitrant organic compounds. The

most popular combinations are UV and US irradiations combined with the use

of heterogeneous catalyst and oxidizing agents (e.g. Fenton catalyst, H2O2 and

ozone) [172]. In this work, the performance of RHSi-Fe in the presence of

H2O2 under US and UV irradiations was tested. Control experiments were

conducted and the results show that the degradation of RhB alone under US

(at 323 K) and UV (at room temperature) irradiations is very slow with 16%

and 6% of decolorization after 1 h irradiation time. The kinetic rate constants

kapp were determined at different type of treatment with corresponding

regression coefficients R2 ranges from 0.97 to 1.00 (Equation 3.7). The effect

of treatment type on the decolorization rate of RhB is shown in Figure 3.13a

and b.

According to the results, US irradiation provides the highest rate of

degradation among all the three tested conditions. UV irradiation shows

considerably high rate of degradation during the initial 25 min but the reaction

rate slows down between 25 and 40 min. Unlike the cases in US and

Chapter 3

94

conventional heating at 323 K, a shift of λmax from 554 to 519 nm was

observed during this period as shown in the inset of Figure 3.13a. This

suggests that de-ethylation (characterized by shift of λmax) and cleavage of

conjugated chromophore structure (characterized by change in λmax) might

happen in a stepwise manner. This might account for the slower rate as

observed between 25th

to 45th

min when UV irradiation was applied.

A comparison of COD removal efficiency after treated with conventional

heating, US and UV irradiations at pH 5.0 and 3.0 was presented in Figure

3.13c.

Figure 3.13 Effect of US and UV irradiations on the degradation of RhB (a).

Inset in Figure 3.13(a) shows the shift of λmax from 15th

to 45th

min when UV

irradiation was applied. The corresponding rate constants, kapp were

determined and shown in (b). The corresponding COD removal for 50 mg/L

RhB after treated with conventional heating, US and UV irradiations at pH 5.0

and 3.0 was shown in (c).

Chapter 3

95

The reduction in COD values in the treated RhB solution indicates the

mineralization of dye molecules along with the color removal. The maximum

COD removal at pH 5.0 and 3.0 were found to be 75.2% and 84.4% after

treated with conventional heating and UV irradiation, respectively. Despite US

irradiation gives the highest decolorization rate, the effect of COD removal is

unexpectedly lower. As suggested by Merouani et al., degradation products of

RhB are recalcitrant towards sonochemical treatment. This is due to the fact

that the intermediate products have very low possibilities of making contact

with •OH radicals, which react mainly at the interface of the bubble. Thus,

sonochemical action that gives rise to products bearing more hydroxyl (or

carboxylic) groups is of low efficiency towards COD abatement [37]. The

results obtained show that RHSi-Fe/H2O2 is effective in the destruction of

RhB in which considerably high degree of mineralization could be achieved.

3.3.13 Stability and Reusability of the Catalyst

It is important to evaluate the stability and reusability of catalyst for practical

implementation of a heterogeneous catalytic system. Thus, reusability and

leaching tests were performed during three consecutive cycles under identical

experimental conditions. To recover the catalyst, the reaction effluent was

centrifuged, and washed with DI water for three times at the end of the

oxidation process. It is worth noting that the filtered catalyst was reused

directly without further drying between consecutive cycles. This is to test the

ease of reusing this catalyst if it is to be used as a fixed bed in continuous

operation. The regeneration of catalyst by simply washing with water without

submitting the catalyst for further drying can reduce the energy consumption

Chapter 3

96

and lower the overall process cost. Figure 3.14a shows the performance of

reused catalyst in terms of RhB removal in 3 consecutive cycles. Amount of

iron leached into the solution during three consecutive cycles was measured to

be 5.0 mg/L (1st cycle), 0.8 mg/L (2

nd cycle), and 0.1 mg/L (3

rd cycle).

Although the initial rate of decolorization decreased gradually during the

successive cycles, no significant decay in decolorization degree of the dye

solution was observed after a reaction time of 2 h when the catalyst was reused.

The decrease in initial rate of decolorization could be a result of the loss of

catalyst and active phase leaching. Nevertheless, deactivation of catalyst could

be attributed to a diversity of factors, such as poisoning of the active catalytic

sites due to adsorbed organic species. This problem could be avoided by

submitting the catalyst to an intermediate calcinations step to restore its

catalytic activity [173]. However, the high energy consumption involved

during the calcinations step might increase the overall process cost. Other

factors including reduction of catalyst specific surface area and dissolution of

some metal oxides from catalyst into the reaction medium are also possible

reasons for the reduction of catalytic activity.

FTIR spectrum of fresh and reused catalysts is shown in Figure 3.14b. The

band around 3400 cm-1

is due to OH stretching vibration of the silanol or

adsorbed water molecules on the silica surface. The band at 1633cm-1

is due to

the bending vibration of the trapped water molecules in the silica matrix. The

strong band at 1070 cm-1 corresponds to the asymmetric vibration of the

siloxane bond, Si O Si, which forms the backbone of the silica matrix. The

band at 792 cm-1

is attributed to the stretching vibration of the Si O Si bond.

Chapter 3

97

The band at 455cm-1 is assigned to the bending vibration of the Si O Si bond.

The band at 950 cm-1 indicates the Si O stretching vibration of the silanol

group [174]. As shown in the figure, there are no significant changes of FTIR

spectrum for the fresh and recovered catalysts. This implies that the surface

functional groups of recovered catalyst remain unchanged after 3 cycles of

repeated use. Hence, the excellent stability of the catalytic activity of RHSi-Fe

could be attributed to the low loss of iron content during oxidation cycles and

to the structural stability of the solid.

Figure 3.14 Performance of RHSi-Fe in consecutive experiments (a). The

FTIR spectrum of fresh and reused catalysts after 1st and 3

rd cycle of repeated

use was shown in (b). Reaction conditions: initial concentration of RhB = 5

mg/L, temperature = 323 K, catalyst dosage = 1 g/dm3, H2O2 amount = 0.98

mmol, pH 5.0.

Chapter 3

98

3.4 Concluding Remarks

The objective of this study is to evaluate the catalytic performance of a rice

hull-based silica catalyst on heterogeneous Fenton-like degradation of an azo-

dye, RhB. The catalyst was found to consist of aggregated silica nanospheres

with about 3 wt.% of iron loaded. The effects of various reaction parameters

such as initial dye concentration, catalyst dosage, H2O2 concentration, pH and

temperature on the catalytic degradation of RhB were studied. Almost 100%

decolorization was achieved within 10 min at an initial pH of 3.0. A

comparative study by applying US and UV irradiations was performed. COD

measurement shows that considerably high degree of mineralization could be

achieved along the decolorization of RhB. The degree of mineralization was

further enhanced by using stepwise addition strategy in which the impact of

adverse reactions could be minimized. In addition, the effects of foreign ions

and ionic strength were investigated and both promoting and inhibiting effects

on the degradation rate were observed. FTIR analysis and ICP-MS results

shows that the catalyst exhibits low iron leaching, good structural stability and

no loss of performance in at least three times of repeated use.

Chapter 4

99

Chapter 4 Green Synthesis of Gold Nanoparticles using Palm

Oil Mill Effluent (POME): A Low-cost and Eco-friendly viable

Approach

4.1 Introduction

Gold Nanoparticles (AuNps) are found to be useful in many applications such

as biomedicine, catalysis, biosensing, electronic and magnetic devices.

Although existing chemical and physical methods have successfully produced

well-defined nanoparticles, these processes are usually costly and involve the

use of toxic chemicals. In addition, synthesis of AuNps using chemical

methods could still lead to the presence of some toxic chemical species being

adsorbed on the surface of nanoparticles which may cause adverse effects in

medical applications. In this case, synthesis of nanoparticles using

microorganisms or plants could be advantageous, in which biomolecules

secreted by the biomass can act as both reducing and capping agents during

the reaction and resulted in nanoparticles that are more biocompatible [74].

Over the past several years, biological synthesis of nanoparticles by using

microorganisms such as bacteria [175] , fungi [176], and yeast [177] have

been reported by several research groups. However, a major drawback of

microbial synthesis is the difficulty to provide good control over size

distribution, shape and crystallinity of nanoparticles [79]. The manipulation of

reaction parameters such as pH and temperature might inactivate the microbes

and hinder the bioreduction process. In addition, the requirements of

specialized facilities and long incubation time could limit the scale of

Chapter 4

100

production as well. In contrast, plant-mediated synthesis of nanoparticles is

comparatively simpler and more cost-effective, in which nanoparticles with

morphology comparable to those synthesized by chemical and physical

methods were obtained [178]. Several bioorganic compounds in plant systems

such as flavonoids, terpenoids, proteins, reducing sugars and alkaloids were

suggested to be involved as either reducing or capping agents during the

formation of nanoparticles [179]. Extracellular syntheses of gold and AgNps

by using leaf extract of Mangifera indica [180], Syzygium cumini [181],

Aloysia citrodora [182] have been shown to produce nanoparticles with

different morphology.

To the best of our knowledge, the use of agro waste from lignocellulosic plant

biomass has not been investigated so far for their ability in biosynthesis of

nanoparticles. Palm oil mill effluent (POME) is a residue generated during the

production of palm oil. Approximately 0.5-0.75 tonne of POME was

generated from every tonne of oil palm fresh fruit produced and discharged in

soil and natural waters as waste [25, 26]. In the present study, we have

demonstrated the ability of POME to synthesize AuNps without addition of

any external surfactant, capping agent or template. The biosynthesized AuNps

were found to be predominantly spherical, with some triangular and hexagonal

shapes observed. The influence of various reaction parameters to the

morphology and size of biosynthesized AuNps was also investigated.

Chapter 4

101

4.2 Material and Methods

4.2.1 Solutions and Reagents

All chemicals used throughout the experiments were of analytical grade and

solutions were prepared by using de-ionized (DI) water (resistivity ≈ 18.2 MΩ

cm) obtained from a Millipore (Billerica, MA) Direct Q purification unit.

Chloroauric acid (HAuCl4 ∙ 3H2O) was procured from Sigma-Aldrich and used

as received. Stock solution of Au(III) was prepared by dissolving solid

chloroauric acid in acidified de-ionized water to 100 M. Working solutions

were obtained by diluting the metal stock solutions to the desired

concentrations. To adjust pH of the metal solutions, 0.10 to 1.0 M NaOH or

HCl solutions were used. The pH of the solutions was measured using

Metrohm 827 pH Lab meter.

4.2.2 Preparation of POME

POME used was collected from a palm oil industry (Kluang Palm Oil

Processing Mill Sdn. Bhd.) located at Kluang, Johor, Malaysia, through a

one-time collection. It was dried at 60°C in an air-supplied oven for

approximately 24 h. After that, the POME was ground by an electrical blender

and subsequently sieved. The portion of ground POME with diameter of

particles ≤ 53 µm was used.

4.2.3 Synthesis of AuNps

The source of gold precursor used in all the experiments was HAuCl4 ∙ 3H2O

in deionized water. Typical reaction mixtures contained 0.1 g of POME

powder in 10 ml of 1.0 mM HAuCl4 solution unless otherwise stated. The

Chapter 4

102

reaction mixtures were stirred at 300 rpm for 48 h at room temperature. At

the end of reaction, the mixture was drawn and filtered for analysis. The effect

of pH on the formation of AuNps was investigated by adjusting the pH of

reaction mixtures (0.1 g POME, 1.0 mM HAuCl4 solution) from 2.0 to 11.0 by

using 0.1 M to 1.0 M HCl and NaOH. Subsequent tests were performed at the

optimum pH as determined. The effect of initial gold salt concentration was

determined by changing the concentration of HAuCl4 to 0.10, 0.25, 0.50, 1.0

and 2.0 mM. To study the effect of temperature on the formation of AuNps,

reaction mixtures containing 0.1 g POME powder, and 10 ml of 1.0 mM

HAuCl4 were incubated at 25 °C, 40 °C and 60 °C, respectively.

4.2.4 Characterization of AuNps

AuNps were characterized by using a Shimadzu UV2450 UV-vis

Spectrophotometer operated at a resolution of 1 nm by using DI water as blank

for each of the sample sets. X-ray diffraction (XRD) diffractogram of dry

nanoparticles powder was obtained using Bruker-AXS Smart Apex CCD

single-crystal diffractometer with CuKα radiation (λ 0.1542 nm). The

measurement was carried out using thoroughly dried thin films of

nanoparticles on Si(111) wafers. The morphology of the AuNps was analyzed

using high resolution images obtained with a JEOL 3010 transmission electron

microscope (TEM).

4.2.5 FTIR Analysis

Samples of POME powder before and after reaction with HAuCl4 as well as

the AuNps produced were analyzed. The FTIR spectra were obtained on a

Chapter 4

103

Shimadzu IR Prestige-21 operated in transmission mode (400-4000 cm-1

) at a

resolution of 4.0 cm-1

with the sample as KBr pellets.

4.2.6 Synthesis of AuNps using POME Extracts

POME powder (0.1 g) was suspended in 2 ml of extraction solvent, sonicated

for 2 h and centrifuged at 14000 rpm for 10 min at room temperature using the

corresponding solvent. The supernatant solutions were collected and kept for

future use. During the biosynthesis, the extracted fractions were added to

appropriate amount of HAuCl4 and made up to a final volume of 10 ml to

obtain a concentration of 1 mM solution. The reaction mixture was stirred for

48 h at room temperature. For the extraction using boiling water, 1 g of the

POME powder was suspended in 50 ml of water and the mixture was boiled

for 10 min.

4.3 Results and Discussions

4.3.1 Effects of Initial pH on the Biosynthesis of AuNps

(Please note that the experimental work included in this section has been

performed by Ng Shi Han, an honors student in 2011, under the supervision of

the PhD candidate)

The initial pH value of the aqueous HAuCl4 solutions was an important

parameter in the synthesis of AuNps using POME. It was observed that the

color of gold solutions changed from pale yellow to purplish-pink (pH 2.0, 3.0)

and wine red (pH 4.0, 5.0) followed by violet and greyish blue (pH 6.0, 7.0,

8.0, 9.0) when the initial pH of solution was increased. However, no color

change was observed for gold solutions at highly alkaline pH (pH 10.0 and

11.0). The color change would be an indication of gold bioreduction mediated

Chapter 4

104

by POME and the subsequent formation of AuNps after 48 h of reaction

(Figure 4.1).

Turkevich has proposed that red is associated with particle sizes smaller than

40 nm and violet is associated with larger particles formed by the aggregation

of colloidal gold [183, 184]. These colors arise due to the excitation of surface

plasmon resonance (SPR) in the AuNps. The UV-vis spectra obtained from

solutions at different initial pH after 48 h of reaction are shown in Figure 4.1.

It can be seen that a SPR absorption peak occurs at around 530 to 560 nm

which are the characteristic absorption bands of AuNps except for pH 10.0

and 11.0. The shift of wavelength could be an indication of the increasing

aggregation of AuNps occurred at higher pH whereas the disappearance of

absorption peak indicates the absence of AuNps. Therefore, the process of

bioreduction of gold ions to AuNps could be followed by UV-vis spectroscopy

and color change. This observation provides a possibility to control the size

and morphology of AuNps produced through pH adjustment.

Figure 4.1 Visual observations and UV-vis absorption spectra of reaction

mixtures at different pH values [(a) control HAuCl4 solution, (b) pH 3.0, (c)

pH 4.0, (d) pH 6.0, (e) pH 8.0] after 48 h of reaction.

Chapter 4

105

In addition to biosynthesis of AuNps, the capability of POME to bind and

concentrate gold from aqueous solutions was also observed at acidic pH. It

was found by ICP-MS analysis, the maximum efficiency of gold biosorption

reached ~95% at pH 2.0 after 2 h of reaction. However, the solution turned to

pink color after 24 h. The change of color suggests the formation of AuNps in

solution after long contact time. Similar result was also observed by Antunes

et al., where the optimal removal of gold ions was found at pH 2.0 within 3 h

of reaction time [185]. In general, biosorption mechanism of gold is ionic

rather than covalent due to the dependence of gold binding with pH. At pH 2.0,

gold is present in solution in anionic form [AuCl4]- and the functional groups

of active biocompounds on the biomass surface such as hydroxyl groups tend

to undergo protonation and become positively charged. The overall positively

charged surface could promote the interaction between protonated functional

groups and the negatively charged [AuCl4]- through electrostatic attraction or

electrovalent bond [179]. As a result, biosorption was preferred over

bioreduction of gold ions. As suggested by Kuyucak and Volesky, the

bioreduction of gold occurred through the oxidation of hydroxyl to carbonyl

groups as shown in Equation 4.1 [186]:

AuCl4- + 3R-OH Au

0 + 3R=O + 3H

+ + 4Cl

- (4.1)

POME has a high content of phenolic acids and flavonoids. The abundant

hydroxyl groups available in these compounds could participate in the gold

bioreduction. The formation of AuNps at pH 2.0 after 24 h suggested that the

bioreduction of gold ions is a slow process at acidic pH due to the protonation

of hydroxyl groups and the competing biosorption process. However, the

Chapter 4

106

biosorption process might be reversible and hence the AuNps could still be

formed as the reaction proceeds.

At higher initial pH values (10.0 and 11.0), no formation of AuNps was

observed as characterized by the disappearance of SPR peak in Figure 4.1.

This might be due to the increasing amount of OH- , which could act as strong

complexing agents of gold ions that interfere with the capping ability of

biomolecules in POME and also compete with the [AuCl4]- for binding to the

biomolecules.

4.3.2 Effect of the HAuCl4 Concentration on the Biosynthesis of AuNps

(Please note that the experimental work included in this section has been

performed by Ng Shi Han, an honors student in 2011, under the supervision of

the PhD candidate)

According to LaMer model, it is predicted that the formation of nanoparticles

could only happen when the precursor concentration is within a suitable range

for nucleation. However, this range might vary amongst different biomass-

assisted synthesis approaches. Therefore, the effect of precursor concentration

on this POME-mediated synthesis was studied (Figure 4.2). It was found that

there was no formation of AuNps by using 0.10 and 0.25 mM of gold chloride.

The formation of AuNps increased as the concentration of gold chloride

increased from 0.50 to 2.0 mM. This is in agreement with the range of initial

precursor concentration (≤10-3

M) that was generally used by most of the

biological methods reported in the literature [179, 187, 188].

Chapter 4

107

Figure 4.2 Visual observations and UV-vis absorption spectra of reaction

mixtures with varying concentration of HAuCl4 (mM) [(a) 0.1, (b) 0.25, (c)

0.50, (d) 1.0, (e) 2.0].

4.3.3 Effect of Temperature on the Biosynthesis of AuNps

(Please note that the experimental work included in this section has been

performed by Ng Shi Han, an honors student in 2011, under the supervision of

the PhD candidate)

The effect of temperature on the formation of AuNps was also investigated

(Figure 4.3). The solution changed from pink to greyish blue followed by dark

brown as the temperature increased. The darkening of solution indicates that

the process of aggregation of smaller AuNps to form larger particles was

favoured over nucleation to form new nanoparticles at higher temperature. The

slight shift of absorption wavelength from 530 to 545 nm when the

temperature was increased from 25 °C to 60 °C suggested the increasing size

of AuNps. This phenomenon may due to the increased rate of reduction at

higher temperature which may hinder the capping process and resulted in the

aggregation of AuNps. In fact, the effect of temperature in directing the shape

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and size of nanoparticles has been reported by several research groups [74,

187, 189]. The ability of POME to biosynthesize well-dispersed AuNps at

room temperature could not only increase the ease of handling but also

provide the advantage of energy-saving.

Figure 4.3 Visual observations and UV-vis absorption spectra of reaction

mixtures at different temperatures [(a) 25°C, (b) 40°C, (c) 60°C] at pH 3 after

3 h of reaction.

4.3.4 Characterization of AuNps

4.3.4.1 TEM Analysis

The morphology of the AuNps was observed by TEM. Figure 4.4a and b

shows the representative TEM images of AuNps synthesized using POME at

pH 3 and pH 8. The TEM images revealed that AuNps formed at pH 3 were

predominantly spherical with some others having occasionally triangular,

truncated triangular and hexagonal shapes. When the pH was increased to 8,

the edges of the nanostructures became not well-defined which could be a

result of aggregation (Figure 4.4 a and b). This observation is consistent with

the shift of wavelength in UV-vis spectra and the change of solution colour

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from pink to greyish blue as shown in Figure 4.1. According to the figures, pH

3 appears to be an optimum pH for this biosynthesis process due to the well-

dispersed and uniform nanostructures observed.

Figure 4.4 TEM images of AuNps synthesized using POME at different

initial pH values: (a) pH 3, (b) pH 8.

Figure 4.5a-d shows the change of morphology of AuNps when the reaction

time was increased from 1 to 48 h. It was found that AuNps with well-defined

structures were observed after 1 h of reaction which suggests that the

biosynthesis process mediated by POME is relatively rapid. As the reaction

proceeded for about 9 and 48 h, aggregation of AuNps became dominant with

some of the nanoparticles overlapped with each other and clustered together.

This observation revealed that reaction time could be an important factor in

controlling the size and morphology of AuNps during the biosynthesis process.

In order to produce AuNps with desired morphology and monodispersity, it is

important to select an appropriate time for the biosynthesis process.

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Figure 4.5 TEM images of AuNps synthesized using POME at pH 3.0 for

(C) 1 h, (D) 3 h, (E) 9 h and (F) 48 h of reaction time.

In fact, the formation of AuNps could be a result of heterogeneous nucleation

and growth followed by Ostwald ripening. The heterogeneous nucleation

could be catalyzed by extraneous material released by POME, in which the

foreign particles act as a scaffold for the crystal to grow on. Since there is no

requirement for the incipient surface energy, this nucleation process is

kinetically favoured. Along with the crystal growth, Ostwald ripening took

place and responsible for the coarsening of size distribution. During Ostwald

ripening, small crystals tend to dissolve in the solution and transfer their mass

to grow larger particles which lead to a greater stability with lower energy

state. The excess Gibb’s free energy associated with the nanoparticles could

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then be minimized by transformation into more energetically favourable

shapes, which was directed by the bioorganic capping molecules [75].

4.3.4.2 Particle Size Distribution

TEM image and histogram of particle size distribution of the AuNps formed

after 3 h of reaction time at pH 3 are shown in Figure 4.6.

Figure 4.6 TEM image (A) and histogram of particle size distribution (B)

of AuNps synthesized using POME at pH 3 with reaction time of 3 h. (C)

X D spectrum of AuNps synthesized using POME. The principal Bragg’s

reflections are identified. (Please note that the data in Figure 4.6(B) was

calculated by Ng Shi Han, an honors student in 2011, under the guidance of

the PhD candidate)

The average particle size was determined to be 18.75 ± 5.96 nm by counting

258 particles on a representative TEM micrograph. The co-existence of

AuNps in smaller and larger size was due to the AuNps formed in early and

later stages of the reaction, which shows that both nucleation to form new

nanoparticles and aggregation to form larger particles happened consecutively.

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4.3.4.3 X-ray diffraction Measurement

The corresponding XRD-diffractogram was analyzed (Figure 4.7). Diffraction

peaks observed at 38.1°, 44.1°, 64.7° and 77.8° can be indexed to the (111),

(200), (220) and (311) Bragg’s reflections of cubic structure of metallic gold

respectively (JCPDS No. 04-0784). Another series of diffraction peaks at

45.9°, 57.0°, 71.6° and 76.1° which can be indexed to the (100), (103), (006)

and (105) Bragg’s reflections of hexagonal phases (JCPD No. 41-1402) were

also observed .

Figure 4.7 XRD spectrum of AuNps synthesized using POME. The

principal Bragg’s reflections are identified.

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XRD analysis suggested that the biosynthesized AuNps are composed of

crystalline gold and they are biphasic in nature. In addition to the assigned

Bragg peaks, additional yet unassigned peaks found at 28.4° and 32.1° were

also noticed, which might be a result of crystallization of some bioorganic

compounds or proteins that are present in the POME on the surface of AuNps.

Similar results were also reported by Shankar et al. and Philip et al. [188, 190]

in the synthesis of Ag nanoparticles using geranium leaf and mushroom

extract.

4.3.4.4 Stability of Nanoparticles

The stability of the AuNps synthesized by POME was examined by comparing

the UV-vis spectra of AuNps that were freshly prepared and stored at 4 °C for

more than 6 months (Figure 4.8). There is no obvious change of solution

colour and SPR peaks observed after 6 months of storage which would suggest

the AuNps synthesized have considerately good stability.

Figure 4.8 AuNps that was freshly prepared (a) and after stored at 4°C for

more than 6 months (b).

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4.3.5 Possible Functional Groups involved in Biosynthesis Process

(Please note that the experimental work included in this section has been

performed by both Ng Shi Han, an honors student in 2011 together with the

PhD candidate)

The involvement of surface functional groups in biosynthesis of AuNps was

studied by FTIR (Figure 4.9).

Figure 4.9 FTIR spectra of (a) POME powder before and (b) after gold

reduction and (c) AuNps synthesized.

The main difference between the spectrum of POME before and after gold

reduction is the disappearance of an absorption band at 2351 cm-1

. This band

can be related to the O-H stretching vibrations from strongly hydrogen-bonded

-COOH group. This phenomenon revealed that biomolecules in the POME

such as proteins or flavonoids that contain abundant of -COOH groups might

play an important role in the bioreduction and stabilization of AuNps during

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the synthetic process. FTIR analysis of AuNps shows the presence of five

bands at 1056, 1645, 2376, 2924 and 3453 cm-1

. The first two absorption

bands are attributed to the C-N stretching of aliphatic amines or

amines/phenols and amide I bands, respectively. The other two bands at 2376

and 2924 cm-1

are characteristic of the O-H stretch from strongly hydrogen-

bonded –COOH group and C-H stretching vibrations. The reappearance of

2376 cm-1

in the spectrum of AuNps suggested the attachment of strongly

hydrogen-bonded carboxylic O-H group onto AuNps during the biosynthesis

process. In addition, the band at 3453 cm-1

indicates the presence of

polyphenolic OH group. Hence, it was deduced that AuNps might be capped

and stabilized by functional groups on the POME surface such as proteins and

polyphenols through the interactions of strongly hydrogen-bonded carboxylic

O-H group.

4.3.6 Synthesis of AuNps using Different POME Extracts

(Please note that the experimental work included in this section has been

performed by Ng Shi Han, an honors student in 2011, under the supervision of

the PhD candidate)

In order to further investigate the compounds involved in the bioreduction of

gold ions mediated by POME, reducing capability of various polarity based

fractions of POME extract were analyzed. These solvents included n-hexane,

ethyl acetate, butanol, unboiled and boiled water, which cover the range from

low to high polarity. The polarity and dielectric constant of the solvents were

listed in Table 4.1 [191].

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Table 4.1 Various solvent of different polarity used for the extraction of

POME.

Type of solvent Polarity Dielectric constant

n-hexane Non-polar 1.89

Ethyl acetate Polar-aprotic 6

Butanol Polar-protic 17.8

Water Polar-protic 78.54

Boiled water Polar-protic 78.54

As found in most of the studies, the extraction procedures for plant extract

generally involve the step of boiling. However, Tripathy et al. indicated that

reducing components of neem leaves were destroyed after drying and caused

the loss of its biosynthesis ability [192]. Hence, except using boiling water, we

have also explored the use of unboiled water in the extraction of POME. The

different fractions of POME extract were then used for the reduction of gold

chloride and their UV-vis spectra were recorded in Figure 4.10.

Figure 4.10 UV-vis spectra of AuNps produced using various solvent

extracts after 48 h of reaction.

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According to the UV-vis spectra, characteristic SPR peaks of AuNps centered

at 550 nm were found for the AuNps synthesized using both boiled and

unboiled water extracts of POME. A weak absorbance peak was also observed

for synthetic process using ethyl acetate extract of POME whereas no

formation of AuNps was found for the butanol and n-hexane extracts. This

observation suggests that most of the bioorganic compounds involved in the

synthesis of AuNps are polar and water-soluble molecules, such as flavonoids,

proteins, reducing sugars, and water-soluble alkaloids present in plant systems.

This is consistent with the observations of C=O, N-H and O-H groups

observed in the FTIR spectrum.

Ethyl acetate as a medium polar-aprotic solvent can be used to extract some

polyphenols and alkaloids. The weak absorbance peak observed in the UV-vis

spectra might be due to the formation of minor amount of AuNps mediated by

some of the polyphenols and alkaloids extracted. Butanol as a polar-protic

solvent could be used for the extraction of glycosides, some polyphenols such

as saponins and flavonoids. Nevertheless, no formation of AuNps was found

by using butanol extract of POME. According to the dielectric constant,

butanol is more polar than ethyl acetate, which should be able to dissolve more

polyphenols and participate in the biosynthesis process. One possible

explanation is that the bioactive polyphenols present in the POME might be

more soluble in polar-aprotic solvent such as ethyl acetate rather than polar-

protic solvent such as butanol. Hexane is an excellent solvent system for the

fatty materials such as lipids, fats and chlorophyll. However, alkaloids,

polyphenols and some other medium-polar compounds are not soluble in

hexane and hence results in the loss of ability for the bioreduction of gold ions.

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4.3.7 Interaction of Biosynthesized AuNps with Mercury Ions

AuNps could form alloys with mercury with varying composition such as

Au3Hg, AuHg, AuHg3 [193]. The potential use of AuNps for the removal of

mercury from water through amalgamation of these two metals have been

demonstrated [194]. The interaction of these biosynthesized AuNps with

mercury was examined, which will correspond to its potential use in water

purification. For the treatment of AuNps with mercury, 50 mg of HgCl2 was

added into 5 ml of the as prepared AuNps. According to the UV-vis spectra,

shift in peak position was observed for AuNps before and after mercury

treatment, which suggested that AuNps were consumed to form Au-Hg

amalgam with different morphology (Figure 4.11). A rapid colour change of

AuNps solution from wine red to greyish after the addition of Hg(II) was

observed within seconds. The interaction of AuNps with Hg(II) shows the

chemical reactivity of the biosynthesized AuNps and its potential use for

mercury removal in water purification.

Figure 4.11 The UV-vis spectra of AuNps (A) before and (B) after mercury

treatment. Inset is the visual observation of AuNps before and after treated

with mercury.

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4.4 Concluding Remarks

The present study reports the synthesis of AuNps from gold precursor using

(POME) without adding external surfactant, capping agent or template. The

biosynthesized AuNps were characterized by using UV-vis spectroscopy,

TEM, XRD, and FTIR. According to the image analysis performed on a

representative TEM micrograph by counting 258 particles, the obtained

AuNps are predominantly spherical with an average size of 18.75 ± 5.96 nm.

In addition, some triangular and hexagonal nanoparticles were also observed.

The influence of various reaction parameters such as reaction pH,

concentration of gold precursor and interaction time to the morphology and

size of biosynthesized AuNps was investigated. This study shows the

feasibility of using agro waste material for the biosynthesis of AuNps which is

potentially more scalable and economic due to its lower cost.

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Chapter 5 Enhancement of Catalytic Performance of a Rice

hull-based Silica Supported Iron Catalyst by Biosynthesized

AuNps for Fenton-like Degradation of Rhodamine B in Water

5.1 Introduction

Rhodamine B (RhB) is chemically stable under various pH. It has

considerably high resistance to photo and oxidative degradation [115, 116].

Toxic and carcinogenic effects upon exposure to RhB have been

experimentally proven [117, 195]. Conventional methods to treat the dye

waste effluents include flocculation, activated carbon adsorption, and

biological treatment. However, these methods usually do not work efficiently

as they are non-destructive, costly and merely involve the transfer of

pollutants from one phase to another. Thus, it is important to remove RhB

from wastewater using a destructive method that is environmentally and

economically viable. A number of advanced oxidation processes (AOPs) have

been used as destructive approaches for the degradation of RhB. These include

Fenton-based oxidation [119-121], photocatalytic degradation [117, 122-124],

sonochemical degradation [37, 125-127], ozonation [128-130], etc.

In our previous work, the catalytic activity of an iron-loaded rice hull-based

silica catalyst (RHSi-Fe) on heterogeneous Fenton-like degradation of RhB

has been assessed. The results obtained show that efficient removal of RhB

could be achieved by using RHSi-Fe in the presence of H2O2. Unlike

conventional Fenton process which only works well at acidic pH (usually pH

3.0) with the addition of considerably large amount of H2O2, this RHSi-Fe

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catalyst is also able to work at circumneutral pH with small amount of H2O2.

However, there are still some areas of improvement needed to further enhance

its catalytic performance. This includes further improvement of the non-

irradiation-assisted catalysis at room temperature and photocatalytic activity

under UV-irradiation.

In addition to the active iron species, the silica support has also been shown to

exhibit noticeable activities towards some catalytic reactions [196-198]. The

photocatalytic activity of several silica-based catalyst such as silica-alumina

[199, 200], silica-supported zirconia [201], and silica-alumina-titania [202]

have been reported under UV-irradiation at room temperature. Nevertheless,

the photocatalytic activity of semiconductor compound is always limited by

the rapid recombination of electron-hole pair. To overcome this problem,

some strategies have been used such as immobilization on a transparent

support or doping with small concentration of metal ions.

It has been proposed that the photoactive sites were formed on the surface of

SiO2 prepared by sol-gel method [203]. As sol-gel method has also been

applied in our synthesis of RHSi-Fe, the silica matrix might also possess some

photoactive sites that could be utilized for photocatalysis. Moreover, the IR

symmetric stretching vibration of the Si-O- non-bridging bond at 950 cm

-1 has

been observed in the FTIR spectrum of RHSi-Fe. The presence this absorption

band is related to the structural change of SiO2 in size of nano-scale from SiO2

to SiO4 [204]. The SiO2 nanoparticles can be photo-excited under UV light

below ~390 nm, which corresponds to a charge transfer from bonding orbital

of Si-O to 2p non-bonding orbital of non-bridging oxygen [55].

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AuNps supported on a range of oxides were found to exhibit enhanced activity

in many reactions [205-208]. Moreover, the role of AuNps core as electron

scavenger has been demonstrated to be useful in the reduction of charges

recombination. In addition, AuNps grafted on nanoparticulate diamond

(Au/HO-npD) have also been reported to be a highly efficient catalyst in a

Fenton process that uses exclusively H2O2 without any of sources of

irradiation. The authors proposed that the catalytic generation of •OH radical

involves a swing between positive and neutral gold states in which gold act as

an electron relay from the oxidation to the reduction semi-reaction [5].

Hence, to further improve the catalytic degradation of RhB, the incorporation

of AuNps to the iron-loaded silica matrix is of particular interest. To make the

preparation of catalyst a greener procedure, biosynthesized AuNps were used.

Our previous study has shown that AuNps can be synthesized by an agro

waste, palm oil mill effluent (POME) [209]. Hence, these POME-assisted

biosynthesized AuNps were utilized in this work. An interesting feature of this

work is the use of natural agro waste as sources of raw material in the

preparation of catalyst.

5.2 Material and Methods

5.2.1 Solutions and Reagents

RhB (90%) and Chloroauric acid (HAuCl4 ∙ 3H2O) were procured from

Sigma-Aldrich and used as received. H2O2 (30% w/w) was obtained from

Scharlau (Barcelona, Spain). Ferric nitrate, Fe(NO3)3∙9H2O (98%) was

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purchased from Fluka (Buchs, Switzerland). 3-aminopropyl-trimethoxysilane

(APTMS) were purchased from Aldrich and used as received. Rice hull waste

was collected from a rice processing mill located at Bangkok, Thailand

whereas POME was collected from a palm oil industry (Kluang Palm Oil

Processing Mill Sdn. Bhd.) located at Kluang, Johor, Malaysia, through a

one-time collection All the other chemicals were of analytical grade and used

without further purification. The RhB solutions were prepared by using

deionized (DI) water (resistivity ≈ 18.2 MΩ cm) obtained from a Millipore

(Billerica, MA) Direct Q purification unit.

5.2.2 Synthesis of Silica-coated AuNps (Au@SiO2)

The AuNps were prepared by the method as stated in section 4.2.3 (1 mM

HAuCl4, 0.1g/L POME powder, 3h reaction time, pH 3.0). This results in a

stable dispersion of gold particles with an average diameter of around 19 nm.

Each 2.5 mL of freshly prepared aqueous solution of 1 mM APTMS is added

to 500 mL of colloidal gold sol under vigorous magnetic stirring. The mixture

was stirred for 30 min. The sodium silicate solution extracted from rice hull

(as mentioned in section 3.2.2) was concentrated by removing the excess

solvent under vacuum suction. The pH of the sodium silicate solution was

adjusted to pH 10-11 using a cation-exchange resin, Amberlite IR 120. The

silica solution was then added to the stock solution containing APTMS under

vigorous magnetic stirring. The resulting dispersion with pH ≈ 9 is then

allowed to stand at room temperature for 1 week, so that the active silica

polymerizes onto the primed gold particle surface. After this, the synthesis of

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gold-silica-iron composite material (Au@RHSi-Fe) began via sol-gel

transition.

5.2.3 Synthesis of gold-silica-iron composite material (Au@RHSi-Fe)

In batches of 500 mL, the Au@SiO2 colloid was concentrated between 25 and

50 mL by applying vacuum suction under stirring at 50°C. To every 5 mL of

the concentrated Au@SiO2 dispersion, 0.15 mL of sodium silicate solution

was added under magnetic stirring. Once silicate addition was completed, the

pH was adjusted to 3 by titration with 3.0 M HNO3 containing appropriate

mass of Fe(NO3)3∙9H2O to obtain 3 wt.% of Fe3+

. The sol-gel transition occurs

over several hours and the resulting gelation was allowed to stand for 1 week.

After gelation, a brown gold-silica-iron composite gel settled under a head of

clear liquid. The gel was recovered by centrifugation at 4000 rpm for 10 min

followed by washing with distilled water under sonication for another 10 min.

To wash the gel thoroughly, the steps of centrifugation and sonication were

repeated 5 times. The gel after washing was dried in an oven at 383 K to

obtain the Au@RHSi-Fe.

5.2.4 Characterization of Catalyst

The formation of Au@RHSi-Fe from Au@SiO2 via sol-gel transition were

characterized by using a Shimadzu UV2450 UV-vis Spectrophotometer

operated at a resolution of 1 nm by using deionized (DI) water as blank for

each of the sample sets. The morphology of the AuNps was analyzed using

high resolution images obtained with a JEOL 3010 transmission electron

microscope (TEM).

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125

In order to analyze the crystallinity of Au@RHSi-Fe, X-ray powder diffraction

(XRD) was carried out using Bruker-AXS Smart Apex CCD single-crystal

diffractometer (Karlsruhe, Germany). The X-ray source was CuKα radiation (λ

= 0.1542 nm). The diffractogram was recorded in 10-80° 2θ range, with a

0.025° step size and a collecting time of 1s per point. The iron content of

catalyst was determined using a Perkin-Elmer Dual-view Optima 5300 DV

ICP-OES system.

5.2.5 Degradation Experiments

10 mg of Au@RHSi–Fe catalyst was added to 10 mL of 10 mg/L RhB

solution to obtain 10 mL of 1g/L catalyst. Then, to establish the adsorption

equilibrium between the RhB and the catalyst, the resulting solution was

stirred in dark for 1 h. After this pretreatment, the zero time reading was taken

and the degradation experiment has begun after the addition of H2O2 (0.196

mmol). Photocatalytic experiments were conducted using a Blak-Ray B-

100AP/R UV lamp (365 nm, 100 W) in a self-made black wooden box at

room temperature. Distance between the source of UV light and test solution

was about 30 cm. No temperature control was applied during the course of

reaction. Non-irradiation-assisted degradation experiments were conducted

under natural light at our experimental conditions.

For the determination of RhB concentration, all the aliquots withdrawn were

centrifuged at 14000 rpm for 5 min followed by subsequent UV-vis analysis of

the supernatant solutions. The UV-vis spectrum of RhB solutions were

recorded from 190 to 700 nm using Hach (Loveland, CO) DR 5000 UV-

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126

vis spectrophotometer. Each experiment was performed in triplicates and

presented as a mean value of 3 runs with a standard deviation of ≤ 5%.

For purpose of comparison, degradation experiments using rice hull-based

silica supported iron catalyst with about the same amount of iron loaded

(RHSi-Fe, 3.2% Fe) was conducted as well.

5.3 Results and discussions

5.3.1 Characterization of Catalyst

UV-vis absorption spectral is a very sensitive tool to monitor the formation of

AuNps and its composite in different sol-gel matrices (i.e. in APTMS). The

formation of of Au@RHSi-Fe from Au@SiO2 via sol-gel transition was

characterized by measuring the surface plasmon resonance (SPR) band using

UV-vis spectrophotometer. The exact position of SPR band is extremely

sensitive to particle size and shape and also the optical and electronic

properties of the medium surrounding the particles. The UV-vis absorbance

spectra of the Au@SiO2 dispersion and Au@RHSi-Fe gel are shown in Figure

5.1. According to the spectra, there is a red shift in the position of SPR band

from 536 to 540 nm with a slight increase of the intensity at the absorption

maximum. As reported by Mulvancy et al., the intensity of the SPR band

increases with a red shift in the position of the absorption maximum when the

shell thickness of the silica layer is increased [210]. This behaviour is due to

the increase in the local refractive index around the particles which is in

agreement with modified Mie’s theory for core-shell particles [211]. Thus, it

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127

is reasonable to assume that the formation of an iron-loaded silica layer around

the APTMS-modified AuNps after the sol-gel transition has caused a red shift

in SPR band due to the increase in the local refractive index around the

Au@SiO2. The slight increase of intensity measured for the SPR band of

Au@RHSi-Fe gel might due to the formation of large-sized silica shell which

makes significant scattering.

Figure 5.1 UV-vis absorbance spectra of the Au@SiO2 dispersion and

Au@RHSi-Fe gel.

TEM micrographs of AuNps before and after its incorporation into the

Au@RHSi-Fe composite are shown in Figure 5.2. The approximate size of

biosynthesized-AuNps before any modification is around 15-20 nm (Figure

5.2a and b). In the Au@RHSi-Fe composite, the AuNps are embedded in the

iron loaded-silica matrix with an increase of particle size to about 40-50 nm

(Figure 5.2 c and d). The increase of particle size is probably induced by the

formation of iron-silica coating on its surface.

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128

Figure 5.2 TEM micrographs of AuNps before (a and b) and after its

incorporation into the Au@RHSi-Fe composite (c and d) at different

magnifications.

The corresponding XRD-diffractogram of Au@RHSi-Fe was analyzed in

Figure 5.3a. As compared to the XRD-diffractogram of AuNps in Figure 5.3b,

the diffraction peaks observed at 38.1°, 44.1°, 64.7° and 77.8° can be indexed

to the (111), (200), (220) and (311) Bragg’s reflections of cubic structure of

metallic gold respectively (JCPDS No. 04-0784). However, some of the peaks

have been masked which is possibly caused by the layer of iron-silica coating

that has shield the crystalline structure of AuNps.

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Figure 5.3 XRD-diffractogram of Au@RHSi-Fe (a) and AuNps (b).

The results from ICP-OES analysis reveal that Au@RHSi-Fe catalyst contains

about 3 to 4% of iron and 0.5% of Au. EDX analysis of the Au@RHSi

composite shows the existence of both Au and Fe which further confirms the

elemental composition of this composite. It should be noted that the Cu peaks

are mainly originated from the copper grid which is used as a holder to support

the sample during the analysis.

Figure 5.4 EDX analysis of Au@RHSi-Fe.

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130

5.3.2 Catalytic Performance of Au@RHSi-Fe under UV-irradiation

The temporal evolution of the spectral changes during the process of RhB

degradation in the Au@RHSi-Fe/H2O2 system after its exposure to UV

irradiation is shown in Figure 5.5a. A comparison between degradation of

RhB in the presence of RHSi-Fe/H2O2 and Au@RHSi-Fe/H2O2 systems as a

function of reaction time was shown in Figure 5.5b. The results show that the

photocatalytic activity of Au@RHSi-Fe has improved significantly after the

incorporation of the AuNps in the silica matrix. Thus, this suggests that

AuNps have a significant effect on the photocatalytic properties of the iron-

loaded silica composite. Although AuNps have been embedded in the silica

matrix, SPR band from AuNps which were expected to appear around 536 nm

was not observed. This might be due to their small intensities as compared

with the dye absorption band which have masked the SPR peak from AuNps.

Figure 5.5 The temporal evolution of the spectral changes during the

process of RhB degradation in the Au@RHSi-Fe/H2O2 system after its

exposure to UV irradiation (a) and a comparison between degradation of

RHSi-Fe and Au@RHSi-Fe as a function of reaction time (b).

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The rate of the heterogeneous photocatalytic degradation of RhB was

described by the pseudo-first-order kinetic model. The pseudo-first-order

equation was given as below:

Ct = C0 exp(-kappt) (5.1)

where C0 and Ct are the initial concentration and the concentration at any time

t of RhB whereas kapp is the pseudo-first order apparent rate constant. The kapp

constants were obtained from the slopes of the straight lines by plotting –

ln(Ct/C0) as a function of reaction time, t, through regression. The kinetic rate

constants kapp were determined in both Au@RHSi-Fe/H2O2 and RHSi-Fe/H2O2

with corresponding regression coefficients R2 ranges from 0.96 to 0.99 (Figure

5.6). kapp for Au@RHSi-Fe/H2O2 and RHSi-Fe/H2O2 systems was determined

to be 0.0256 and 0.0078 min-1

, respectively.

Figure 5.6 The plot of –ln(Ct/C0) as a function of reaction time, t and the

corresponding kapp for both RHSi-Fe/H2O2 and Au@RHSi-Fe/H2O2.

The value of apparent rate constants, kapp, gives an indication for the activity

of the photocatalyst. The activity is directly proportional to the reaction rate

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132

constant. Thus, the incorporation of AuNps has increased the rate of RhB

degradation for significantly under our experimental conditions.

5.3.3 Catalytic Performance of Au@RHSi-Fe at Room Temperature

In addition to photocatalytic degradation, it was found in our previous study

that the degradation of RhB in the RHSi-Fe/H2O2 system for non-irradiation-

assisted process slows down significantly at room temperature (Batch

degradation experiments using RHSi-Fe/H2O2 were carried out at 50°C). To

evaluate whether there is any improvement after the incorporation of AuNps

in the iron-silica matrix, the catalytic performance of Au@RHSi-Fe over RhB

solutions with pH 5.0 was examined at room temperature (Figure 5.7).

Figure 5.7 A comparison between degradation of RhB in the presence of

RHSi-Fe/H2O2 and Au@RHSi-Fe/H2O2 systems at pH 5.0 under room

temperature as a function of reaction time.

Based on the results in Figure 5.7, there is an improvement of the rate of RhB

degradation when Au@RHSi-Fe was used. As compared to the RHSi-Fe/H2O2

system which only removed about 40% of RhB in a reaction time of 90 min,

Au@RHSi-Fe/H2O2 system is able to achieve almost complete removal of

Chapter 5

133

RhB within the same reaction time. The ability of Au@RHSi-Fe to remove

RhB at room temperature is highly desirable as it eliminates the need of

heating during the degradation process which could also reduce the energy

consumption and operational costs.

5.3.4 Catalytic Performance of Au@RHSi-Fe at different pH

The catalytic performance of Au@RHSi-Fe at different pH using non-

irradiation-assisted process on the removal of RhB at room temperature was

evaluated and the results are shown in Figure 5.8.

Figure 5.8 Catalytic performances of Au@RHSi-Fe/H2O2 (a) and RHSi-

Fe/H2O2 (b) systems at different pH under room temperature as a function of

reaction time. A comparison of the performance of both catalyst at the 50th

min during the course of reaction was depicted in (c).

Chapter 5

134

As depicted in Figure 5.8, both catalysts work better when the solution pH was

adjusted to 3.0 at room temperature. In particular, Au@RHSi-Fe exhibits a

better efficiency with almost complete removal of RhB achieved within 1 h at

room temperature. This is much faster than the degradation of RhB catalyzed

by RHSi-Fe under the same experimental conditions in which there are still

about 36% of RhB remain in the solution after 50 min of reaction. The better

performance for both catalysts at pH 3.0 is expected as most of the Fenton-

based processes show optimum performance at this pH. This behavior could

be attributed to the favorable formation of more hydroxyl radicals and the

adsorption of RhB in cationic form to the negatively charged catalyst surface

at this pH.

When the pH is increased to higher values (i.e. pH 5.0, 7.0 and 9.0), the

reaction rate slows down in both catalyst systems. As compared to RHSi-Fe,

Au@RHSi-Fe is still able to work efficiently at higher pH despite the reaction

rate was decreased with increasing pH. A comparison of the performance of

both catalysts at the 50th

min during the course of reaction shows that the

incorporation of AuNps has increased the efficiency of the catalyst within the

range of pH tested at room temperature (Figure 5.8c). At higher pH, there is an

increased formation of RhB molecules in zwitterionic form which could result

in the aggregation of RhB and hinder the adsorption of dye molecules onto the

surface structure of catalyst. The degradation of RhB was slowed down

because the catalysis process is mainly occurred on the sorbed species. Since

dye waste effluent is always discharged at different pH, the improved

performance of Au@RHSi-Fe is highly desirable.

Chapter 5

135

5.3.5 Possible Mechanisms

Although the iron ions are regarded as the active catalyzing species which

break down the H2O2 into •OH radicals, the discussion here will be focused on

the enhancement of catalytic performance on RhB degradation due to the

incorporation of AuNps. Unlike the bulk counterpart, AuNps are able to

exhibit unique properties that could be effectively used in catalysis.

Photocatalytic degradation of organic dye catalyzed by silica nanoparticles has

been reported [55].

When a photon of UV light strikes the surface of silica, it can absorb the UV

light and transfer an electron from its valence band (vb) to the conduction

band (cb) upon excitation. This process generates a positively charged hole in

the valence band (hvb+) and a negative charge in the conduction band (ecb

-),

result in the formation of photocatalytic active centers on the surface of SiO2

according to equation 5.2:

SiO2 + hv → ecb- + hvb

+ (5.2)

The hvb+

can then react with the chemisorbed H2O molecules to form reactive

species such as •OH radicals that are useful for dye degradation (Equation 5.3)

(H2O → H+

+ OH-) + hvb

+ → •OH (5.3)

On the other hand, the ecb- could react with acceptor such as dissolved O2 and

transform it into super oxide radical anion (O2•

), leading to the additional

formation of •OOH as given in the following equation:

O2 + ecb- → O2

• + (H

+ + OH

-) → •OOH +

OH (5.4)

Chapter 5

136

Electron donors such as OH-

and •OOH can react with hvb+, forming •OH

radial (Equation 5.5)

•OOH + OH- + hvb

+ → •OH + •OH (5.5)

In addition, some of the ecb- and hvb

+ might be trapped in the surface states and

react with the adsorbed species.

However, the efficiency of silica catalyst is limited by the recombination of

ecb- and hvb

+. The recombination process can occur within few nanoseconds on

the surface of the particle and the resulting energy is dissipated as heat. The

presence of a AuNps can enhance the formation of •OH radical by reducing

the recombination of charges. During the reaction, the Au ions are able to trap

electrons in the conduction band of SiO2 and generate holes in which they can

act as electron scavengers as illustrated in Equation 5.6-5.8:

Au + ecb- → Au(ecb

-) (5.6)

Au(ecb-) + H

+ → Au + H2 (5.7)

Au(ecb-) + O2

→ Au + O2

• + (H

+ + OH

-) → •OOH +

OH

(5.8)

Each Au3+

ion is able to take up three (ecb-) and assist in the generation of three

•OH. Thus, the AuNps can actually act as an electron-hole separation centres

to retard the recombination process in the silica matrix, whereby the

photocatalytic efficiency is increased due to the formation of more •OH

radicals [56].

In addition to the enhancement of photocatalytic activity of silica material,

AuNps could also act as an efficient Fenton-catalyst. In particular, Navalon et

Chapter 5

137

al. have reported on the use Au/HO-npD in a Fenton process that use

exclusively H2O2 without using any of sources of irradiation [5]. The authors

proposed that the catalytic generation of •OH radical involves a swing

between positive and neutral gold states in which gold act as an electron relay

from the oxidation to the reduction semi-reaction. This suggests that the

observed improved performance of Au@RHSi could come from the catalytic

properties of the AuNps as well. In particular, AuNps might be catalytically

active in the degradation of RhB under both UV-assisted and non-irradiation-

assisted conditions. According to a review of literature and the clues obtained

from our experimental data, a possible mechanism of catalytic activity of

Au@RHSi-Fe is proposed in Figure 5.9.

Figure 5.9 Possible mechanisms proposed for the improved catalytic

activity of Au@RHSi-Fe.

It was deduced that the observed enhancement of catalytic performance of

Au@RHSi-Fe could be a synergic effect of the photocatalytic activity of silica

support enhanced by AuNps, catalytic properties of AuNps, and Fenton-like

degradation induced by the iron loaded in the silica matrix.

Chapter 5

138

5.4 Concluding Remarks

The present study shows that the incorporation of AuNps in a silica supported

iron catalyst could enhance the catalytic activity towards the degradation of

RhB at a wide range of pH (including moderate basic pH conditions) under

room temperature with or without the application of UV-irradiation. The role

of AuNps is mainly as an electron scavenger which could consume the

electrons released by the silica support upon the exposure to UV light. In

addition, the better catalytic performance observed in non-irradiation-assisted

process without applying additional heating could be attributed to the ability

of AuNps to act as an electron relay between the oxidation and the reduction

semi-reactions. Furthermore, we have taken the advantage of using natural

agro waste as source of raw materials during the preparation of AuNps and

silica solution which is the unique feature of this study.

Chapter 6

139

Chapter 6 Conclusion and Future Work

This dissertation has successfully demonstrated the potential of using eco-

friendly materials transformed from natural agro wastes for remediation of

pollutants in water. The main objective of these works is to provide a low-cost

and environmentally benign solution for the treatment of some persistent

pollutants in water by taking advantages of using low-cost agro waste as

sources of raw material. In particular, POME which is discharged in large

amount from oil palm industry has been demonstrated to be a potential

biosorbent for the adsorptive removal of toxic heavy metals such as Hg(II) and

Cd(II) and also it could be used for the biosynthesis of noble metallic

nanoparticles such as AuNps. In addition, another agro waste, rice hull has

been demonstrated to be a useful source of silica which could be utilized in the

synthesis of Fenton-like catalyst for destructive removal of dye pollutants such

as RhB. The POME-assisted biosynthesized AuNps were also used to

enhance the catalytic performance of this rice hull-based Fenton-like catalyst,

which can be one of the potential uses of the biosynthesized AuNps. In

conclusion, this dissertation has demonstrated the possibility of turning agro

waste to useful material and shed light on their applications in remediation of

persistent pollutants in water.

6.1 Summary of Results

In Chapter 2, NaOH modified-POME was shown to be effective for the rapid

removal of Cd(II) and Hg(II) from aqueous solutions at an optimum pH of 4.5.

The biosorption process involves several mechanisms, in which the pseudo-

Chapter 6

140

second-order kinetic appears to be the dominant mechanism. The kinetic data

is also found to follow the intra-particle diffusion and external mass transfer

mechanisms although they are not the rate-determining steps. According to

Langmuir isotherms, the maximum adsorption capacities for Cd(II) and Hg(II)

were determined to be 21.9 and 19.23 mgg-1

at 293 and 283 K, respectively.

Both adsorption processes are thermodynamically spontaneous in nature.

Good recovery and regeneration of biosorbent are achieved by using EDDS as

desorption solution. These results show the potential of using oil palm waste

as biosorbent for the removal of Cd(II) and Hg(II) in aqueous solutions. The

low cost of biosorbent as compared to commercially available adsorbent

makes it a more affordable general decontamination technology especially for

developing countries.

In Chapter 3, heterogeneous Fenton-like degradation of RhB has been

demonstrated over a catalyst based upon a readily available, particularly low

cost and abundant rice hull waste, which is easily and safely handled.

Influences of various reaction parameters such as initial dye concentration,

catalyst dosage, H2O2 concentration, pH and temperature on the catalytic

degradation of RhB have been investigated and the results show that this

catalyst is able to work with a wide range of dye concentrations with fast

degradation rate (in 10 min) achieved at pH 3.0. Study on the effects of

foreign salts shows that the existence of salts such as NaCl and Na2SO4 could

actually enhance the degradation rate instead of inhibiting it in this catalyst

system, which is highly favourable for practical applications. In addition, the

degree of mineralization could be further enhanced by using a stepwise

addition strategy, in which a more efficient use of H2O2 could be achieved

Chapter 6

141

without the requirement of additional reagent. Considerably good stability and

reusability of this catalyst have also been demonstrated. This work

demonstrates the great potential of using RHSi-Fe as a low cost and green

catalyst for heterogeneous Fenton-like degradation of hazardous dye, giving

an enhanced treatability of textile wastewater for disposal in municipal

wastewater plants.

In Chapter 4, biosynthesis of AuNps mediated by POME has been

demonstrated to be a simple, low cost, and environmentally friendly method to

synthesize AuNps at room temperature. The AuNps synthesized are

predominantly spherical with an average size of 18.75 nm. The morphology

and size of AuNps could be controlled by varying the reaction conditions such

as initial pH of the HAuCl4 solution and reaction temperature. Bioactive

compounds involved in the biosynthesis are most likely proteins and water

soluble polyphenols in POME that contains amine and carbonyl groups.

Interaction of biosynthesized AuNps with Hg(II) has been investigated which

demonstrates the chemical reactivity of these nanoparticles.

Lastly, in Chapter 5, the incorporation of POME-assisted biosynthesized

AuNps in a silica supported iron catalyst has been shown to be able to enhance

the catalytic activity towards the degradation of RhB. This catalyst shows a

better performance on RhB removal at room temperature over a range of

tested pH including moderate basic pH conditions. The enhancement of

catalytic activity after incorporation of AuNps has been observed in both UV-

assisted and non-irradiation-assisted processes. The role of AuNps has been

proposed to be mainly an electron scavenger which could consume the

Chapter 6

142

electrons released by the silica support upon the exposure to UV light. The

improved performance observed for non-irradiation-assisted process could be

attributed to the ability of AuNps to act as an electron relay between the

oxidation and the reduction semi-reactions. One interesting feature of this

study is the use of natural agro waste as source of raw materials during the

preparation of AuNps and silica solution which reduces the use of hazardous

synthetic chemicals with a greener preparation procedure.

6.2 Current Challenges and Directions for Future Work

Although the work presented in this dissertation has shown some of the

potential applications of agro waste materials in remediation of pollutants in

water, there are still some challenges that need to be addressed to facilitate

further advancement in future research work. Firstly, the lack of understanding

of the biosorption mechanism and also the lack of robustness of biosorbent

due to its inherent complex biological nature appear to be the main factors that

limit the practical exploitation of biosorption. Hence, future research direction

could be focused more on the mechanistic study of biosorbent, especially

studies for biosorption mechanism at molecular level which might help to

form a more well-defined relationship between biosorption performances with

those important process parameters such as pH, ionic strength, temperature

and biomass loading, etc. Pre-treatment of biomass and its immobilization on

a suitable support could also be applied to improve the performance and the

mechanical property of biosorbent.

Chapter 6

143

To scale up the biosorption process, column studies and pilot-scale studies are

strongly encouraged. In addition, more research should be performed based on

multi-element solutions and real industrial effluents to further evaluate the

applicability of this new technique before it is brought into actual practice.

Furthermore, the post-treatment of exhausted biosorbent and possible methods

to recover the sorbed metals for recycling are also some important aspects that

worth further exploration.

Secondly, despite the fact that degradation of dye pollutants using AOPs have

been extensively studied, their practical applications are limited by the high

operative costs and the interferences induced by dye additives and other

impurities that could hamper their performances. Although the use of low-cost

agro-based material and dosing strategy might be useful for cost reduction

with consumption of less reagents and synthetic chemicals, it is suggested to

combine AOPs such as Fenton-based treatment with biological processes so

that it could further reduce the operating cost. In general, AOPs are able to

transform recalcitrant compounds into more biodegradable by-products so that

the toxicity of wastewater should be greatly reduced before total

mineralization has been achieved. However, toxicity induced by some of the

degraded by-products should not be overlooked. Thus, further research work

on the assessment of toxicity and biodegradability of these pre-treated

effluents is of particular interest.

Also, if the biosynthesized nanoparticles are to be applied in actual

applications, either for biomedical application or water purification, its

toxicity profiles and possible environmental impacts as compared to

Chapter 6

144

chemically synthesized nanoparticles have to be established. Hence,

toxicological studies to compare the relative effect of biogenic and chemically

synthesized nanoparticles are important to confirm their advantages as claimed

by the inherent non-toxic nature. In addition, another significant challenge of

biosynthesis is to consistently reproduce the biosynthesized nanoparticles with

desired sizes and morphology. In this case, the identification of possible

compounds responsible for the bioreduction and biocapping process requires

further investigation by using more advanced analytical methods. Moreover,

the mechanisms and role of the biosynthesized AuNps in the enhancement of

catalytic performance of the rice hull-based iron silica catalyst in Fenton-like

degradation process needs more detailed studies to better elucidate the

mechanism.

The other challenge of using natural biomass materials is their inherent

variability which depends on geographical location, variety, climate conditions

and harvest method. Hence, changes of performance by using agro waste

obtained from different sources could be another problem. In order to

overcome this problem, comparative studies by using biomass of different

sources could be another important research direction. In particular, more

research should be focused on those countries with abundant resources of low-

cost agricultural waste.

As a conclusion, the use of natural agro waste shows a huge potential in

turning the waste to useful resources. However, in order to overcome the

challenges and limitations faced at current research stage, continual research

work in relevant directions is always necessary to achieve its success in future.

References

145

References

[1] M.A. Rauf, S.S. Ashraf, Radiation induced degradation of dyes - An

overview, J. Hazard. Mater., 166 (2009) 6-16.

[2] D. Sud, G. Mahajan, M.P. Kaur, Agricultural waste material as potential

adsorbent for sequestering heavy metal ions from aqueous solutions - A

review, Bioresour. Technol., 99 (2008) 6017-6027.

[3] C. Fernández, M.S. Larrechi, M.P. Callao, An analytical overview of

processes for removing organic dyes from wastewater effluents, Trends. Anal.

Chem., 29 (2010) 1202-1211.

[4] A.N. Soon, B.H. Hameed, Heterogeneous catalytic treatment of synthetic

dyes in aqueous media using Fenton and photo-assisted Fenton process,

Desalination, 269 (2011) 1-16.

[5] S. Navalon, R. Martin, M. Alvaro, H. Garcia, Gold on Diamond

Nanoparticles as a Highly Efficient Fenton Catalyst, Angew. Chem. Int. Ed.,

49 (2010) 8403-8407.

[6] P.P. Gan, S.F.Y. Li, Potential of plant as a biological factory to synthesize

gold and silver nanoparticles and their applications, Rev. Environ. Sci.

Biotechnol., 11 (2012) 169-206.

[7] J.O. Duruibe, M.O.C. Ogwuegbu, J.N. Egwurugwu, Heavy metal pollution

and human biotoxic effects Int. J. Phys. Sci., (2007) 112-118.

[8] P. Almeida, L.B. Stearns, Political opportunities and local grassroots

environmental movements: The case of Minamata, Soc. Probl., 45 (1998) 37-

60.

References

146

[9] R. Bruggemann, E. Halfon, Ranking for environmental hazard of the

chemicals spilled in the Sandoz accident in November 1986, Sci. Total

Environ., 97-98 (1990) 827-837.

[10] M. Kaji, Role of experts and public participation in pollution control: The

case of Itai-itai disease in Japan, Ethics. Sci. Environ. Polit., 12 (2012) 99-111.

[11] T.K. Burki, Nigeria's lead poisoning crisis could leave a long legacy, The

Lancet, 379 (2012) 792.

[12] B. Volesky, Z.R. Holan, Biosorption of heavy metals, Biotechnol. Prog.,

11 (1995) 235-250.

[13] T. Prior, D. Giurco, G. Mudd, L. Mason, J. Behrisch, Resource depletion,

peak minerals and the implications for sustainable resource management,

Global. Environ. Chang., (2012) 577-587.

[14] A. Shukla, Y.-H. Zhang, P. Dubey, J.L. Margrave, S.S. Shukla, The role

of sawdust in the removal of unwanted materials from water, J. Hazard. Mater.,

95 (2002) 137-152.

[15] F. Fu, Q. Wang, Removal of heavy metal ions from wastewaters: A

review, J. Environ. Manag., 92 (2011) 407-418.

[16] Q. Chang, G. Wang, Study on the macromolecular coagulant PEX which

traps heavy metals, Chem. Eng. Sci., 62 (2007) 4636-4643.

[17] U. Farooq, J.A. Kozinski, M.A. Khan, M. Athar, Biosorption of heavy

metal ions using wheat based biosorbents - A review of the recent literature,

Bioresour. Technol., 101 (2010) 5043-5053.

[18] E.V. Soares, H.M.V.M. Soares, Bioremediation of industrial effluents

containing heavy metals using brewing cells of Saccharomyces cerevisiae as a

green technology: A review, Environ. Sci. Pollut. Res., 19 (2012) 1066-1083.

References

147

[19] D. Kratochvil, B. Volesky, Advances in the biosorption of heavy metals,

Trends Biotechnol., 16 (1998) 291-300.

[20] G.M. Gadd, Biosorption: critical review of scientific rationale,

environmental importance and significance for pollution treatment, J. Chem.

Technol. Biotechnol., 84 (2009) 13-28.

[21] D. Kratochvil, P. Pimentel, B. Volesky, Removal of Trivalent and

Hexavalent Chromium by Seaweed Biosorbent, Environ. Sci. Technol., 32

(1998) 2693-2698.

[22] T.A. Davis, B. Volesky, A. Mucci, A review of the biochemistry of

heavy metal biosorption by brown algae, Water Res., 37 (2003) 4311-4330.

[23] B. Volesky, Detoxification of metal-bearing effluents: biosorption for the

next century, Hydrometallurgy, 59 (2001) 203-216.

[24] A. Haug, B. Larsen, O. Smidsrod, A Study of the Constitution of Alginic

Acid by Partial Acid Hydrolysis, Acta Chem. Scand., 20 (1966) 183-190.

[25] M.Z. Alam, N.A. Kabbashi, S.N.I.S. Hussin, Production of bioethanol by

direct bioconversion of oil-palm industrial effluent in a stirred-tank bioreactor,

J. Ind. Microbiol. Biotechnol., 36 (2009) 801-808.

[26] K.W. Chou, I. Norli, A. Anees, Evaluation of the effect of temperature,

NaOH concentration and time on solubilization of palm oil mill effluent

(POME) using response surface methodology (RSM), Bioresour. Technol.,

101 (2010) 8616-8622.

[27] T. Mumtaz, N.A. Yahaya, S. Abd-Aziz, N.A. Abdul Rahman, P.L. Yee,

Y. Shirai, M.A. Hassan, Turning waste to wealth-biodegradable plastics

polyhydroxyalkanoates from palm oil mill effluent - a Malaysian perspective,

J. Clean. Prod., 18 (2010) 1393-1402.

References

148

[28] T. Ahmad, M. Rafatullah, A. Ghazali, O. Sulaiman, R. Hashim, Oil palm

biomass-based adsorbents for the removal of water pollutantsa review, J.

Environ. Sci. Health, Part C: Environ. Carcinog. Ecotoxicol. Rev., 29 (2011)

177-222.

[29] G. Harman, R. Patrick, T. Spittler, Removal of heavy metals from

polluted waters using lignocellulosic agricultural waste products, Ind.

Biotechn., 3 (2007) 366-374.

[30] M. Betancur, P.R. Bonelli, J.A. Velásquez, A.L. Cukierman, Potentiality

of lignin from the Kraft pulping process for removal of trace nickel from

wastewater: Effect of demineralisation, Bioresour. Technol., 100 (2009) 1130-

1137.

[31] A. Demirbas, Heavy metal adsorption onto agro-based waste materials: A

review, J. Hazard. Mater., 157 (2008) 220-229.

[32] S.C. de Vries, G.W.J. van de Ven, M.K. van Ittersum, K.E. Giller,

Resource use efficiency and environmental performance of nine major biofuel

crops, processed by first-generation conversion techniques, Biomass.

Bioenerg., 34 (2010) 588-601.

[33] C.S. Goh, K.T. Tan, K.T. Lee, S. Bhatia, Bio-ethanol from lignocellulose:

Status, perspectives and challenges in Malaysia, Bioresour. Technol., 101

(2010) 4834-4841.

[34] W.G. Lan, M.K. Wong, N. Chen, Y.M. Sin, Orthogonal array design as a

chemometric method for the optimization of analytical procedures part 5.*

Three-level design and its application in microwave dissolution of biological

samples, The Analyst, 120 (1995) 1115-1124.

References

149

[35] W.G. Lan, M.K. Wong, N. Chen, Y.M. Sin, Orthogonal array design as a

chemometric method for the optimization of analytical procedures. Part 1.

Two-level design and its application in microwave dissolution of biological

samples, The Analyst, 119 (1994a) 1659-1667.

[36] F. Ahmedchekkat, M.S. Medjram, M. Chiha, A.M.A. Al-Bsoul,

Sonophotocatalytic degradation of Rhodamine B using a novel reactor

geometry: Effect of operating conditions, Chem. Eng. J., 178 (2011) 244-251.

[37] S. Merouani, O. Hamdaoui, F. Saoudi, M. Chiha, C. Pétrier, Influence of

bicarbonate and carbonate ions on sonochemical degradation of Rhodamine B

in aqueous phase, J. Hazard. Mat. 175 (2010) 593-599.

[38] Q. Yang, H. Choi, S.R. Al-Abed, D.D. Dionysiou, Iron-cobalt mixed

oxide nanocatalysts: Heterogeneous peroxymonosulfate activation, cobalt

leaching, and ferromagnetic properties for environmental applications, Appl.

Catal., B, 88 (2009) 462-469.

[39] C.L. Yap, S. Gan, H.K. Ng, Fenton based remediation of polycyclic

aromatic hydrocarbons-contaminated soils, Chemosphere, 83 (2011) 1414-

1430.

[40] S.Q. Liu, S. Cheng, L.R. Feng, X.M. Wang, Z.G. Chen, Effect of alkali

cations on heterogeneous photo-Fenton process mediated by Prussian blue

colloids, J. Hazard. Mater., 182 (2010) 665-671.

[41] S. Gazi, A. Rajakumar, N.D.P. Singh, Photodegradation of organic dyes

in the presence of [Fe(III)-salen]Cl complex and H2O2 under visible light

irradiation, J. Hazard. Mater., 183 (2010) 894-901.

[42] C. Walling, Fenton's reagent revisited, Acc. Chem. Res., 8 (1975) 125-

131.

References

150

[43] S. Navalon, M. Alvaro, H. Garcia, Heterogeneous Fenton catalysts based

on clays, silicas and zeolites, Appl. Catal., B, 99 (2010) 1-26.

[44] A.A. Burbano, D.D. Dionysiou, M.T. Suidan, T.L. Richardson, Oxidation

kinetics and effect of pH on the degradation of MTBE with Fenton reagent,

Water Res., 39 (2005) 107-118.

[45] M.F. Hou, L. Liao, W.D. Zhang, X.Y. Tang, H.F. Wan, G.C. Yin,

Degradation of rhodamine B by Fe(0)-based Fenton process with H 2O 2,

Chemosphere, 83 (2011) 1279-1283.

[46] S. Navalon, A. Dhakshinamoorthy, M. Alvaro, H. Garcia, Heterogeneous

Fenton catalysts based on activated carbon and related materials,

ChemSusChem, 4 (2011) 1712-1730.

[47] X. Hu, B. Liu, Y. Deng, H. Chen, S. Luo, C. Sun, P. Yang, S. Yang,

Adsorption and heterogeneous Fenton degradation of 17α-methyltestosterone

on nano Fe3O4/MWCNTs in aqueous solution, Appl. Catal., B, 107 (2011)

274-283.

[48] D.I. Petkowicz, S.B.C. Pergher, C.D.S. da Silva, Z.N. da Rocha, J.H.Z.

dos Santos, Catalytic photodegradation of dyes by in situ zeolite-supported

titania, Chem. Eng. J., 158 (2010) 505-512.

[49] M.M. El-Moselhy, Photo-degradation of acid red 44 using Al and Fe

modified silicates, J. Hazard. Mater., 169 (2009) 498-508.

[50] F. Adam, J.N. Appaturi, A. Iqbal, The utilization of rice husk silica as a

catalyst: Review and recent progress, Catal. Today, 190 (2012) 2-14.

[51] H. Yoshida, Silica-based quantum photocatalysts for selective reactions,

Curr. Opin. Solid. St. M., 7 (2003) 435-442.

References

151

[52] U. Kalapathy, A. Proctor, J. Shultz, Silica xerogels from rice hull ash:

Structure, density and mechanical strength as affected by gelation pH and

silica concentration, J. Chem. Technol. Biotechnol., 75 (2000) 464-468.

[53] D. An, Y. Guo, Y. Zhu, Z. Wang, A green route to preparation of silica

powders with rice husk ash and waste gas, Chem. Eng. J., 162 (2010) 509-514.

[54] Y. Inaki, H. Yoshida, T. Yoshida, T. Hattori, Active Sites on Mesoporous

and Amorphous Silica Materials and Their Photocatalytic Activity: An

Investigation by FTIR, ESR, VUV-UV and Photoluminescence

Spectroscopies, J. Phys. Chem. B., 106 (2002) 9098-9106.

[55] Y. Badr, M.G. Abd El-Wahed, M.A. Mahmoud, Photocatalytic

degradation of methyl red dye by silica nanoparticles, J. Hazard. Mater., 154

(2008) 245-253.

[56] Y. Badr, M.A. Mahmoud, Photocatalytic degradation of methyl orange

by gold silver nano-core/silica nano-shell, J. Phys. Chem. Solids, 68 (2007)

413-419.

[57] L. Lin, W. Wang, J. Huang, Q. Li, D. Sun, X. Yang, H. Wang, N. He, Y.

Wang, Nature factory of silver nanowires: Plant-mediated synthesis using

broth of Cassia fistula leaf, Chem. Eng. J., 162 (2010) 852-858.

[58] D.V. Leff, P.C. Ohara, J.R. Heath, W.M. Gelbart, Thermodynamic

control of gold nanocrystal size: Experiment and theory, J. Phys. Chem., 99

(1995) 7036-7041.

[59] M.P. Pileni, Nanosized particles made in colloidal assemblies, Langmuir,

13 (1997) 3266-3276.

References

152

[60] N.H.H. Abu Bakar, J. Ismail, M. Abu Bakar, Synthesis and

characterization of silver nanoparticles in natural rubber, Mater. Chem. Phys.,

104 (2007) 276-283.

[61] M. Adlim, M. Abu Bakar, K.Y. Liew, J. Ismail, Synthesis of chitosan-

stabilized platinum and palladium nanoparticles and their hydrogenation

activity, J. Mol. Catal. A: Chem., 212 (2004) 141-149.

[62] N.E. Kotelnikova, G. Wegener, M. Stoll, V.N. Demidov, Comparative

Study of Intercalation of Zero-Valent Silver into the Cellulose Matrix by

Raster and Transmission Microscopy, Russ. J. Appl. Chem., 76 (2003) 117-

123.

[63] E.M. Egorova, A.A. Revina, Synthesis of metallic nanoparticles in

reverse micelles in the presence of quercetin, Colloids. Surf., A., 168 (2000)

87-96.

[64] K. Vijayaraghavan, S.P.K. Nalini, Biotemplates in the green synthesis of

silver nanoparticles, Biotechnol. J., 5 (2010) 1098-1110.

[65] S.S. Shankar, A. Rai, A. Ahmad, M. Sastry, Rapid synthesis of Au, Ag,

and bimetallic Au core-Ag shell nanoparticles using Neem (Azadirachta indica)

leaf broth, J. Colloid Interface Sci., 275 (2004) 496-502.

[66] S. Kundu, S. Panigrahi, S. Praharaj, S. Basu, S.K. Ghosh, A. Pal, T. Pal,

Anisotropic growth of gold clusters to gold nanocubes under UV irradiation,

Nanotechnology, 18 (2007).

[67] K. Okitsu, Y. Mizukoshi, T.A. Yamamoto, Y. Maeda, Y. Nagata,

Sonochemical synthesis of gold nanoparticles on chitosan, Mater. Lett., 61

(2007) 3429-3431.

References

153

[68] M.E. Meyre, M. Tréguer-Delapierre, C. Faure, Radiation-induced

synthesis of gold nanoparticles within lamellar phases. Formation of aligned

colloidal gold by radiolysis, Langmuir, 24 (2008) 4421-4425.

[69] T. Tsuji, T. Kakita, M. Tsuji, Preparation of nano-size particles of silver

with femtosecond laser ablation in water, Appl. Surf. Sci., 206 (2003) 314-320.

[70] F. Mafuné, J.Y. Kohno, Y. Takeda, T. Kondow, Full physical preparation

of size-selected gold nanoparticles in solution: Laser ablation and laser-

induced size control, J. Phys. Chem. B, 106 (2002) 7575-7577.

[71] N. Ahmad, S. Sharma, M.K. Alam, V.N. Singh, S.F. Shamsi, B.R. Mehta,

A. Fatma, Rapid synthesis of silver nanoparticles using dried medicinal plant

of basil, Colloids Surf., B, 81 (2010) 81-86.

[72] J.L. Gardea-Torresdey, K.J. Tiemann, J.G. Parsons, G. Gamez, I. Herrera,

M. Jose-Yacaman, XAS investigations into the mechanism(s) of Au(III)

binding and reduction by alfalfa biomass, Microchem. J., 71 (2002) 193-204.

[73] L. Castro, M.L. Blázquez, F. González, J.A. Muñoz, A. Ballester,

Extracellular biosynthesis of gold nanoparticles using sugar beet pulp, Chem.

Eng. J., 164 (2010) 92-97.

[74] A.I. Lukman, B. Gong, C.E. Marjo, U. Roessner, A.T. Harris, Facile

synthesis, stabilization, and anti-bacterial performance of discrete Ag

nanoparticles using Medicago sativa seed exudates, J. Colloid Interface Sci.,

353 (2011) 433-444.

[75] O. Fiehn, J. Kopka, R.N. Trethewey, L. Willmitzer, Identification of

uncommon plant metabolites based on calculation of elemental compositions

using gas chromatography and quadrupole mass spectrometry, Anal. Chem.,

72 (2000) 3573-3580.

References

154

[76] P. Rajani, K. SriSindhura, T.N.V.K.V. Prasad, O.M. Hussain, P.

Sudhakar, P. Latha, M. Balakrishna, V. Kambala, K. Raja Reddy, Fabrication

of biogenic silver nanoparticles using agricultural crop plant leaf extracts, in:

AIP Conf. Proc., 2010, pp. 148-153.

[77] K.B. Narayanan, N. Sakthivel, Phytosynthesis of gold nanoparticles using

leaf extract of Coleus amboinicus Lour, Mater. Charact., 61 (2010) 1232-1238.

[78] V. Kumar, S.K. Yadav, Plant-mediated synthesis of silver and gold

nanoparticles and their applications, J. Chem. Technol. Biotechnol., 84 (2009)

151-157.

[79] K.B. Narayanan, N. Sakthivel, Biological synthesis of metal

nanoparticles by microbes, Adv. Colloid Interface Sci., 156 (2010) 1-13.

[80] R.G. Haverkamp, A.T. Marshall, The mechanism of metal nanoparticle

formation in plants: Limits on accumulation, J. Nanopart. Res., 11 (2009)

1453-1463.

[81] N. Wattanapenpaiboon, M.L. Wahlqvist, Phytonutrient deficiency: The

place of palm fruit, Asia. Pac. J. Clin. Nutr, 12 (2003) 363-368.

[82] M.A. Oliver, Soil and human health: A review, Eur. J. Soil. Sci, 48 (1997)

573-592.

[83] M.F. García, R.P. García, N.B. García, A. Sanz-Medel, On-line

preconcentration of inorganic mercury and methylmercury in sea-water by

sorbent-extraction and total mercury determination by cold vapour atomic

absorption spectrometry, Talanta, 41 (1994) 1833-1839.

[84] D.T. Rein, M. Breidenbach, D.M. Nettelbeck, Y. Kawakami, G.P. Siegal,

W.K. Huh, M. Wang, A. Hemminki, G.J. Bauerschmitz, M. Yamamoto, Y.

References

155

Adachi, K. Takayama, P. Dall, D.T. Curiel, Evaluation of tissue-specific

promoters in carcinomas of the cervix uteri, J. Gene. Med, 6 (2004) 1281-1289.

[85] N. Chubar, J.R. Carvalho, M.J.N. Correia, Heavy metals biosorption on

cork biomass: effect of the pre-treatment, Colloids Surf., A, 238 (2004) 51-58.

[86] T.A. Davis, F. Llanes, B. Volesky, A. Mucci, Metal selectivity of

Sargassum spp. and their alginates in relation to their α-L-guluronic acid

content and conformation, Environ. Sci. Technol., 37 (2003b) 261-267.

[87] T.A. Davis, F. Llanes, B. Volesky, G. Diaz-Pulido, L. McCook, A. Mucci,

1H-NMR study of Na alginates extracted from Sargassum spp. in Relation to

metal biosorption, Appl. Biochem. Biotechnol., 110 (2003a) 75-90.

[88] D. Park, Y.S. Yun, J.M. Park, The past, present, and future trends of

biosorption, Biotechnol. Bioprocess Eng., 15 86-102.

[89] A.Z. Abdullah, B. Salamatinia, A.H. Kamaruddin, Application of

response surface methodology for the optimization of NaOH treatment on oil

palm frond towards improvement in the sorption of heavy metals, Desalination,

244 (2009) 227-238.

[90] H.B. Wan, W.G. Lan, M.K. Wong, C.Y. Mok, Orthogonal array designs

for the optimization of liquid chromatographic analysis of pesticides, Anal.

Chim. Acta, 289 (1994) 371-380.

[91] Y.S. Ho, D.A.J. Wase, C.F. Forster, Batch nickel removal from aqueous

solution by sphagnum moss peat, Water Res., 29 (1995) 1327-1332.

[92] F.A. Abu Al-Rub, Biosorption of zinc on palm tree leaves: Equilibrium,

kinetics, and thermodynamics studies, Sep. Sci. Technol., 41 (2006) 3499-

3515.

References

156

[93] S.S. Gupta, K.G. Bhattacharyya, Adsorption of Ni(II) on clays, J. Colloid

Interface Sci., 295 (2006) 21-32.

[94] M. Argun, S. Dursun, A new approach to modification of natural

adsorbent for heavy metal adsorption, Bioresour. Technol., (2008) 2516-2527.

[95] O. Altin, H.Ȍ. Ȍzbelge, T. Doğu, Use of general purpose adsorption

isotherms for heavy metal-clay mineral interactions, J. Colloid Interface Sci.,

198 (1998) 130-140.

[96] S.K. Das, J. Bhowal, A.R. Das, A.K. Guha, Adsorption behavior of

rhodamine B on Rhizopus oryzae biomass, Langmuir, 22 (2006) 7265-7272.

[97] P. Miretzky, C. Munoz, A. Carrillo-Chavez, Cd (II) removal from

aqueous solution by Eleocharis acicularis biomass, equilibrium and kinetic

studies, Bioresour. Technol., 101 (2010) 2637-2642.

[98] M. Kilic, H. Yazici, M. Solak, A comprehensive study on removal and

recovery of copper(II) from aqueous solutions by NaOH-pretreated

Marrubium globosum ssp globosum leaves powder: Potential for utilizing the

copper(II) condensed desorption solutions in agricultural applications,

Bioresour. Technol., 100 (2009) 2130-2137.

[99] M.A. Atieh, O.Y. Bakather, B. Al-Tawbini, A.A. Bukhari, F.A.

Abuilaiwi, M.B. Fettouhi, Effect of Carboxylic Functional Group

Functionalized on Carbon Nanotubes Surface on the Removal of Lead from

Water, Bioinorg. Chem. Appl., 2010 (2010) 603978.

[100] S. Saygideger, O. Gulnaz, E.S. Istifli, N. Yucel, Adsorption of Cd(II),

Cu(II) and Ni(II) ions by Lemna minor L.: Effect of physicochemical

environment, J. Hazard. Mater., 126 (2005) 96-104.

References

157

[101] M.A. Wahab, S. Jellali, N. Jedidi, Ammonium biosorption onto sawdust:

FTIR analysis, kinetics and adsorption isotherms modeling, Bioresour.

Technol., 101 5070-5075.

[102] Y.S. Al-Degs, M.I. El-Barghouthi, A.A. Issa, M.A. Khraisheh, G.M.

Walker, Sorption of Zn(II), Pb(II), and Co(II) using natural sorbents:

Equilibrium and kinetic studies, Water Res., 40 (2006) 2645-2658.

[103] Y.S. Ho, Review of second-order models for adsorption systems, J.

Hazard. Mater., 136 (2006) 681-689.

[104] G. McKay, THE ADSORPTION OF BASIC DYE ONTO SILICA

FROM AQUEOUS-SOLUTION SOLID DIFFUSION-MODEL, Chem. Eng.

Sci., 39 (1984) 129-138.

[105] N.F. Cardoso, E.C. Lima, I.S. Pinto, C.V. Amavisca, B. Royer, R.B.

Pinto, W.S. Alencar, S.F.P. Pereira, Application of cupuassu shell as

biosorbent for the removal of textile dyes from aqueous solution, J. Environ.

Manage., 92 (2011) 1237-1247.

[106] Z. Bekci, Y. Seki, L. Cavas, Removal of malachite green by using an

invasive marine alga Caulerpa racemosa var. cylindracea, J. Hazard. Mater.,

161 (2009) 1454-1460.

[107] G. Bayramoglu, M.Y. Arica, Construction a hybrid biosorbent using

Scenedesmus quadricauda and Ca-alginate for biosorption of Cu(II), Zn(II)

and Ni(II): Kinetics and equilibrium studies, Bioresour. Technol., 100 (2009)

186-193.

[108] X.S. Wang, F.Y. Li, W. He, H.H. Miao, Hg(II) Removal from Aqueous

Solutions by Bacillus subtilis Biomass, CLEAN, 38 (2010) 44-48.

References

158

[109] L. Deng, Y. Su, H. Su, X. Wang, X. Zhu, Sorption and desorption of

lead (II) from wastewater by green algae Cladophora fascicularis, J. Hazard.

Mater., 143 (2007) 220-225.

[110] W. Yin, W. Wang, L. Zhou, S. Sun, L. Zhang, CTAB-assisted synthesis

of monoclinic BiVO4 photocatalyst and its highly efficient degradation of

organic dye under visible-light irradiation, J. Hazard. Mater., 173 (2010) 194-

199.

[111] D. Wang, X. Li, J. Chen, X. Tao, Enhanced photoelectrocatalytic

activity of reduced graphene oxide/TiO 2 composite films for dye degradation,

Chem. Eng. J., (2012).

[112] W. Dong, Y. Sun, Q. Ma, L. Zhu, W. Hua, X. Lu, G. Zhuang, S. Zhang,

Z. Guo, D. Zhao, Excellent photocatalytic degradation activities of ordered

mesoporous anatase TiO 2-SiO 2 nanocomposites to various organic

contaminants, J. Hazard. Mater., 229-230 (2012) 307-320.

[113] B.S. Kadu, R.A. Limaye, A.D. Natu, R.C. Chikate, Potentiality of fe-ni

nanocomposites towards environmental abetment of magenta dye, Environ.

Prog. Sustain. Energy., 27 (2012).

[114] S. Navalon, M. De Miguel, R. Martin, M. Alvaro, H. Garcia,

Enhancement of the catalytic activity of supported gold nanoparticles for the

fenton reaction by light, J. Am. Chem. Soc., 133 (2011) 2218-2226.

[115] A.K. Mittal, C. Venkobachar, Uptake of Cationic Dyes by Sulfonated

Coal: Sorption Mechanism, Ind. Eng. Chem. Res., 35 (1996) 1472-1474.

[116] Z. Ai, L. Lu, J. Li, L. Zhang, J. Qiu, M. Wu, Fe@Fe2O3 Core-shell

nanowires as iron reagent. 1. Efficient degradation of rhodamine by a novel

sono-fenton process, J. Phys. Chem. C, 111 (2007) 4087-4093.

References

159

[117] R. Jain, M. Mathur, S. Sikarwar, A. Mittal, Removal of the hazardous

dye rhodamine B through photocatalytic and adsorption treatments, J. Environ.

Manag., 85 (2007) 956-964.

[118] J. Rochat, P. Demenge, J.C. Rerat, Toxicologic study of a fluorescent

tracer: rhodamine B, Toxicol. Eur. Res., 1 (1978) 23-26.

[119] X. Xue, K. Hanna, N. Deng, Fenton-like oxidation of Rhodamine B in

the presence of two types of iron (II, III) oxide, J. Hazard. Mater., 166 (2009)

407-414.

[120] L. Zhang, Y. Nie, C. Hu, J. Qu, Enhanced Fenton degradation of

Rhodamine B over nanoscaled Cu-doped LaTiO3 perovskite, Appl. Catal., B,

125 (2012) 418-424.

[121] Y. Fan, Z. Ai, L. Zhang, Design of an electro-Fenton system with a

novel sandwich film cathode for wastewater treatment, J. Hazard. Mater., 176

(2010) 678-684.

[122] K. Byrappa, A.K. Subramani, S. Ananda, K.M. Lokanatha Rai, R.

Dinesh, M. Yoshimura, Photocatalytic degradation of rhodamine B dye using

hydrothermally synthesized ZnO, Bull. Mater. Sci., 29 (2006) 433-438.

[123] S.Q. Liu, S. Cheng, L. Luo, H.Y. Cheng, S.J. Wang, S. Lou,

Degradation of dye rhodamine B under visible irradiation with Prussian blue

as a photo-Fenton reagent, Environ. Chem. Lett., 9 (2011) 31-35.

[124] T. Li, L. Zhao, Y. He, J. Cai, M. Luo, J. Lin, Synthesis of g-

C3N4/SmVO4 composite photocatalyst with improved visible light

photocatalytic activities in RhB degradation, Appl. Catal., B, 129 (2013) 255-

263.

References

160

[125] S. Merouani, O. Hamdaoui, F. Saoudi, M. Chiha, Sonochemical

degradation of Rhodamine B in aqueous phase: Effects of additives, Chem.

Eng. J., 158 (2010) 550-557.

[126] P.R. Gogate, M. Sivakumar, A.B. Pandit, Destruction of Rhodamine B

using novel sonochemical reactor with capacity of 7.5 l, Sep. Purif. Technol.,

34 (2004) 13-24.

[127] L. Zhu, T. Ghosh, C.-Y. Park, Z.-D. Meng, W.-C. Oh, Enhanced

Sonocatalytic Degradation of Rhodamine B by Graphene-TiO2 Composites

Synthesized by an Ultrasonic-Assisted Method, Chin. J. Catal., 33 (2012)

1276-1283.

[128] Ê. Machado, V. Sales Dambros, L. Kist, E. Alcayaga Lobo, S. Tedesco,

C. Moro, Use of Ozonization for the Treatment of Dye Wastewaters

Containing Rhodamine B in the Agate Industry, Water, Air, Soil Pollut., 223

(2012) 1753-1764.

[129] B. Cuiping, X. Xianfeng, G. Wenqi, F. Dexin, X. Mo, G. Zhongxue, X.

Nian, Removal of rhodamine B by ozone-based advanced oxidation process,

Desalination, 278 (2011) 84-90.

[130] W. Qin, X. Li, J. Qi, Catalytic Ozonation of Rhodamine B over CuO

Catalyst Confined in Multiwalled Carbon Nanotubes: An Experimental and

Theoretical Account, Langmuir, 27 (2009) 11729-11729.

[131] S. Singh, M. Das, S.K. Khanna, Biodegradation of Malachite Green and

Rhodamine B by Cecal Microflora of Rats, Biochem. Biophys. Res. Commun.,

200 (1994) 1544-1550.

References

161

[132] Y. Guo, J. Zhao, H. Zhang, S. Yang, J. Qi, Z. Wang, H. Xu, Use of rice

husk-based porous carbon for adsorption of Rhodamine B from aqueous

solutions, Dyes Pigm., 66 (2005) 123-128.

[133] S. Priscilla Prabhavathi, P. Shameela Rajam, S. Sivapriya, R. Vijayaraj,

A comparative study of the adsorption capacities of the adsorbents rice husk

and activated alumina in the removal of the dye-rhodamine B using adsorption

technique, Indian J. Environ. Prot., 31 (2011) 819-824.

[134] N. Kannan, M.M. Sundaram, Studies on the removal of rhodamine B by

adsorption using various carbons - A comparative study, Fresenius Environ.

Bull., 10 (2001) 814-822.

[135] N.K. Daud, B.H. Hameed, Decolorization of Acid Red 1 by Fenton-like

process using rice husk ash-based catalyst, J. Hazard. Mater., 176 (2010) 938-

944.

[136] T.-H. Liou, H.-S. Lin, Synthesis and surface characterization of silica

nanoparticles from industrial resin waste controlled by optimal gelation

conditions, J. Ind. Eng. Chem., (2012).

[137] F. Adam, J. Andas, Amino benzoic acid modified silica-An improved

catalyst for the mono-substituted product in the benzylation of toluene with

benzyl chloride, J. Colloid Interface Sci., 311 (2007) 135-143.

[138] M. Dükkanci, G. Gündüz, S. Yilmaz, R.V. Prihod'ko, Heterogeneous

Fenton-like degradation of Rhodamine 6G in water using CuFeZSM-5 zeolite

catalyst prepared by hydrothermal synthesis, J. Hazard. Mater., 181 (2010)

343-350.

[139] M. Hagman, E. Heander, J.L.C. Jansen, Advanced oxidation of

refractory organics in leachate - Potential methods and evaluation of

References

162

biodegradability of the remaining substrate, Environ. Technol., 29 (2008) 941-

946.

[140] F. Adam, J. Andas, I.A. Rahman, A study on the oxidation of phenol by

heterogeneous iron silica catalyst, Chem. Eng. J., 165 (2010) 658-667.

[141] J.H. Ramirez, C.A. Costa, L.M. Madeira, G. Mata, M.A. Vicente, M.L.

Rojas-Cervantes, A.J. López-Peinado, R.M. Martín-Aranda, Fenton-like

oxidation of Orange II solutions using heterogeneous catalysts based on

saponite clay, Appl. Catal., B, 71 (2007) 44-56.

[142] N.K. Daud, M.A. Ahmad, B.H. Hameed, Decolorization of Acid Red 1

dye solution by Fenton-like process using Fe-Montmorillonite K10 catalyst,

Chem. Eng. J., 165 (2010) 111-116.

[143] A. Chouket, H. Elhouichet, M. Oueslati, H. Koyama, B. Gelloz, N.

Koshida, Energy transfer in porous-silicon/laser-dye composite evidenced by

polarization memory of photoluminescence, Appl. Phys. Lett., 91 (2007).

[144] L. Zhang, Y. Nie, C. Hu, X. Hu, Decolorization of methylene blue in

layered manganese oxide suspension with H 2O 2, J. Hazard. Mater., 190

(2011) 780-785.

[145] T.M. Elmorsi, Y.M. Riyad, Z.H. Mohamed, H.M.H. Abd El Bary,

Decolorization of Mordant red 73 azo dye in water using H2O2/UV and

photo-Fenton treatment, J. Hazard. Mater., 174 (2010) 352-358.

[146] M.A. Behnajady, N. Modirshahla, S.B. Tabrizi, S. Molanee, Ultrasonic

degradation of Rhodamine B in aqueous solution: Influence of operational

parameters, J. Hazard. Mater., 152 (2008) 381-386.

References

163

[147] N. Shimizu, C. Ogino, M.F. Dadjour, T. Murata, Sonocatalytic

degradation of methylene blue with TiO2 pellets in water, Ultrason.

Sonochem., 14 (2007) 184-190.

[148] S. Basu-Modak, R.M. Tyrrell, Singlet oxygen: A primary effector in the

ultraviolet A/near-visible light induction of the human heme oxygenase gene,

Cancer Res., 53 (1993) 4505-4510.

[149] K. Yu, S. Yang, S.A. Boyd, H. Chen, C. Sun, Efficient degradation of

organic dyes by BiAg xO y, J. Hazard. Mater., 197 (2011) 88-96.

[150] A. Brunmark, Formation of electronically excited states during the

interaction of p-benzoquinone with hydrogen peroxide, J Biolumin

Chemilumin., 4 (1989) 219-225.

[151] F. Chen, W. Ma, J. He, J. Zhao, Fenton degradation of malachite green

catalyzed by aromatic additives, J. Phys. Chem. A, 106 (2002) 9485-9490.

[152] E. Neyens, J. Baeyens, A review of classic Fenton's peroxidation as an

advanced oxidation technique, J. Hazard. Mater., 98 (2003) 33-50.

[153] S.-Q. Liu, S. Cheng, L.-R. Feng, X.-M. Wang, Z.-G. Chen, Effect of

alkali cations on heterogeneous photo-Fenton process mediated by Prussian

blue colloids, J. Hazard. Mater., 182 (2010) 665-671.

[154] J.M. Lin, M. Liu, Chemiluminescence from the decomposition of

peroxymonocarbonate catalyzed by gold nanoparticles, J. Phys. Chem. B, 112

(2008) 7850-7855.

[155] J.E. Biaglow, A.V. Kachur, The Generation of Hydroxyl Radicals in the

Reaction of Molecular Oxygen with Polyphosphate Complexes of Ferrous Ion,

Radiat. Res., 148 (1997) 181-187.

References

164

[156] A. Mehrdad, B. Massoumi, R. Hashemzadeh, Kinetic study of

degradation of Rhodamine B in the presence of hydrogen peroxide and some

metal oxide, Chem. Eng. J., 168 (2011) 1073-1078.

[157] Y. Li, S. Sun, M. Ma, Y. Ouyang, W. Yan, Kinetic study and model of

the photocatalytic degradation of rhodamine B (RhB) by a TiO2-coated

activated carbon catalyst: Effects of initial RhB content, light intensity and

TiO2 content in the catalyst, Chem. Eng. J., 142 (2008) 147-155.

[158] J. Hong, N. Yuan, Y. Wang, S. Qi, Efficient degradation of Rhodamine

B in microwave-H2O2 system at alkaline pH, Chem. Eng. J., 191 (2012) 364-

368.

[159] D.D. Dionysiou, M.T. Suidan, I. Baudin, J.M. Laîné, Effect of hydrogen

peroxide on the destruction of organic contaminants-synergism and inhibition

in a continuous-mode photocatalytic reactor, Appl. Catal., B, 50 (2004) 259-

269.

[160] T.A. Gad-Allah, M.E.M. Ali, M.I. Badawy, Photocatalytic oxidation of

ciprofloxacin under simulated sunlight, J. Hazard. Mater., 186 (2011) 751-755.

[161] S.S. Lin, M.D. Gurol, Catalytic decomposition of hydrogen peroxide on

iron oxide: Kinetics, mechanism, and implications, Environ. Sci. Technol., 32

(1998) 1417-1423.

[162] M. Mohammadi, A.J. Hassani, A.R. Mohamed, G.D. Najafpour,

Removal of Rhodamine B from Aqueous Solution Using Palm Shell-Based

Activated Carbon: Adsorption and Kinetic Studies, J. Chem. Eng. Data, 55

(2010) 5777-5785.

References

165

[163] H.M.H. Gad, A.A. El-Sayed, Activated carbon from agricultural by-

products for the removal of Rhodamine-B from aqueous solution, J. Hazard.

Mater., 168 (2009) 1070-1081.

[164] I. Arslan, I.A. Balcioglu, D.W. Bahnemann, Heterogeneous

photocatalytic treatment of simulated dyehouse effluents using novel TiO2-

photocatalysts, Appl. Catal., B, 26 (2000) 193-206.

[165] K. Shakir, A.F. Elkafrawy, H.F. Ghoneimy, S.G. Elrab Beheir, M.

Refaat, Removal of rhodamine B (a basic dye) and thoron (an acidic dye) from

dilute aqueous solutions and wastewater simulants by ion flotation, Water Res.,

44 (2010) 1449-1461.

[166] J.J. Pignatello, Dark and photoassisted Fe3+-catalyzed degradation of

chlorophenoxy herbicides by hydrogen peroxide, Environ. Sci. Technol., 26

(1992) 944-951.

[167] M.C. Lu, Y.F. Chang, I.M. Chen, Y.Y. Huang, Effect of chloride ions

on the oxidation of aniline by Fenton's reagent, J. Environ. Manage., 75 (2005)

177-182.

[168] P.R. Gogate, A.B. Pandit, A review of imperative technologies for

wastewater treatment I: Oxidation technologies at ambient conditions, Adv.

Environ. Res., 8 (2004) 501-551.

[169] M. Alkan, O. Demirbaş, S. Çelikçapa, M. Doǧan, Sorption of acid red

57 from aqueous solution onto sepiolite, J. Hazard. Mater., 116 (2004) 135-

145.

[170] Y.J. Oh, T.C. Gamble, D. Leonhardt, C.H. Chung, S.R.J. Brueck, C.F.

Ivory, G.P. Lopez, D.N. Petsev, S.M. Han, Monitoring FET flow control and

wall adsorption of charged fluorescent dye molecules in nanochannels

References

166

integrated into a multiple internal reflection infrared waveguide, Lab on a

Chip - Miniaturisation for Chemistry and Biology, 8 (2008) 251-258.

[171] P.M. Dove, C.J. Nix, The influence of the alkaline earth cations,

magnesium, calcium, and barium on the dissolution kinetics of quartz,

Geochim. Cosmochim. Acta, 61 (1997) 3329-3340.

[172] Y.L. Pang, A.Z. Abdullah, S. Bhatia, Review on sonochemical methods

in the presence of catalysts and chemical additives for treatment of organic

pollutants in wastewater, Desalination, 277 (2011) 1-14.

[173] C. Catrinescu, C. Teodosiu, M. Macoveanu, J. Miehe-Brendlé, R. Le

Dred, Catalytic wet peroxide oxidation of phenol over Fe-exchanged pillared

beidellite, Water Res., 37 (2003) 1154-1160.

[174] F. Adam, T.S. Chew, J. Andas, A simple template-free sol-gel synthesis

of spherical nanosilica from agricultural biomass, J. Sol-Gel Sci. Technol., 59

(2011) 580-583.

[175] S.M. Taheri Otaqsara, Biosynthesis of quasi-spherical Ag nanoparticle

by Pseudomonas aeruginosa as a bioreducing agent, Eur. Phys. J. Appl. Phys.,

56 (2011).

[176] N. Pradhan, R.R. Nayak, A.K. Pradhan, L.B. Sukla, B.K. Mishra, In situ

synthesis of entrapped silver nanoparticles by a fungus-penicillium

purpurogenum, Nanosci. Nanotechnol. Lett., 3 (2011) 659-665.

[177] S.K. K, A. R, P. Arumugam, S. Berchmans, Synthesis of gold

nanoparticles: an ecofriendly approach using Hansenula anomala, ACS Appl.

Mater. Interfaces, 3 (2011) 1418-1425.

[178] J.G. Parsons, J.R. Peralta-Videa, J.L. Gardea-Torresdey, Chapter 21 Use

of plants in biotechnology: Synthesis of metal nanoparticles by inactivated

References

167

plant tissues, plant extracts, and living plants, in: Developments in

Environmental Science, 2007, pp. 463-485.

[179] Y. Zhou, W. Lin, J. Huang, W. Wang, Y. Gao, L. Lin, Q. Li, M. Du,

Biosynthesis of gold nanoparticles by foliar broths: Roles of biocompounds

and other attributes of the extracts, Nanoscale. Res. Lett., 5 (2010) 1351-1359.

[180] D. Philip, Rapid green synthesis of spherical gold nanoparticles using

Mangifera indica leaf, Spectrochim. Acta, Part A, 77 (2010) 807-810.

[181] V. Kumar, S.C. Yadav, S.K. Yadav, Syzygium cumini leaf and seed

extract mediated biosynthesis of silver nanoparticles and their characterization,

J. Chem. Technol. Biotechnol., 85 (2010) 1301-1309.

[182] D. Cruz, P.L. Falé, A. Mourato, P.D. Vaz, M. Luisa Serralheiro, A.R.L.

Lino, Preparation and physicochemical characterization of Ag nanoparticles

biosynthesized by Lippia citriodora (Lemon Verbena), Colloids Surf., B, 81

(2010) 67-73.

[183] J. Turkevich, COLLOIDAL GOLD. PART I, Gold Bull., 18 (1985a) 86-

91.

[184] J. Turkevich, COLLOIDAL GOLD. PART II: COLOUR,

COAGULATION, ADHESION, ALLOYING AND CATALYTIC

PROPERTIES, Gold Bull., 18 (1985b) 125-131.

[185] A.P.M. Antunes, G.M. Watkins, J.R. Duncan, Batch studies on the

removal of gold(III) from aqueous solution by Azolla filiculoides, Biotechnol.

Lett., 23 (2001) 249-251.

[186] N. Kuyucak, B. Volesky, Accumulation of gold by algal biosorbent,

Biorecovery, 1 (1989) 189–204.

References

168

[187] G.S. Ghodake, N.G. Deshpande, Y.P. Lee, E.S. Jin, Pear fruit extract-

assisted room-temperature biosynthesis of gold nanoplates, Colloids Surf., B,

75 (2010) 584-589.

[188] S.S. Shankar, A. Ahmad, M. Sastry, Geranium Leaf Assisted

Biosynthesis of Silver Nanoparticles, Biotechnol. Prog., 19 (2003) 1627-1631.

[189] E.C. Njagi, H. Huang, L. Stafford, H. Genuino, H.M. Galindo, J.B.

Collins, G.E. Hoag, S.L. Suib, Biosynthesis of Iron and Silver Nanoparticles at

Room Temperature Using Aqueous Sorghum Bran Extracts, Langmuir, (2011)

null-null.

[190] D. Philip, Biosynthesis of Au, Ag and Au-Ag nanoparticles using edible

mushroom extract, Spectrochim. Acta, Part A, 73 (2009) 374-381.

[191] A.I. Vogel, A.R. Tatchell, B.S. Furnis, A.J. Hannaford, P.W.G. Smith,

Vogel's Textbook of Practical Organic Chemistry, fifth ed, Wiley, New York,

1989.

[192] A. Tripathy, A.M. Raichur, N. Chandrasekaran, T.C. Prathna, A.

Mukherjee, Process variables in biomimetic synthesis of silver nanoparticles

by aqueous extract of Azadirachta indica (Neem) leaves, J. Nanopart. Res., 12

(2010) 237-246.

[193] T. Pradeep, Anshup, Noble metal nanoparticles for water purification: A

critical review, Thin Solid Films, 517 (2009) 6441-6478.

[194] K. Lisha, A. Anshup, T. Pradeep, Towards a practical solution for

removing inorganic mercury from drinking water using gold nanoparticles,

Gold Bull., 42 (2009) 144-152.

[195] J. Rochat, P. Demenge, J.C. Rerat, Toxicologic study of a fluorescent

tracer: rhodamine B, Toxicol. Eur. Res., 1 (1978) 23-26.

References

169

[196] J. Haber, K. Pamin, L. Matachowski, D. Mucha, Catalytic performance

of the dodecatungstophosphoric acid on different supports, Appl. Catal., A,

256 (2003) 141-152.

[197] R. Bal, B.B. Tope, T.K. Das, S.G. Hegde, S. Sivasanker, Alkali-Loaded

Silica, a Solid Base: Investigation by FTIR Spectroscopy of Adsorbed CO2

and Its Catalytic Activity, J. Catal., 204 (2001) 358-363.

[198] G. Fornasari, F. Trifirò, Oxidation with no-redox oxides: ammoximation

of cyclohexanone on amorphous silicas, Catal. Today, 41 (1998) 443-455.

[199] H. Yoshida, N. Matsushita, Y. Kato, T. Hattori, Active sites in sol-gel

prepared silica-alumina for photoinduced non-oxidative methane coupling,

Phys. Chem. Chem. Phys., 4 (2002) 2459-2465.

[200] H. Yoshida, Y. Kato, T. Hattori, Photoinduced non-oxidative methane

coupling over silica-alumina, Stud. Surf. Sci. Catal., 130 (2000) 659-664.

[201] H. Yoshida, M.G. Chaskar, Y. Kato, T. Hattori, Fine structural

photoluminescence spectra of silica-supported zirconium oxide and its

photoactivity in direct methane conversion, Chem. Commun. (Cambridge, U.

K.), (2002) 2014-2015.

[202] Y. Kato, N. Matsushita, H. Yoshida, T. Hattori, Highly active silica-

alumina-titania catalyst for photoinduced non-oxidative methane coupling,

Catal. Commun., 3 (2002) 99-103.

[203] H. Yoshida, C. Murata, T. Hattori, Screening study of silica-supported

catalysts for photoepoxidation of propene by molecular oxygen, J. Catal., 194

(2000) 364-372.

References

170

[204] T. Furukawa, K.E. Fox, W.B. White, Raman spectroscopic investigation

of the structure of silicate glasses. III. Raman intensities and structural units in

sodium silicate glasses, J. Chem. C., 75 (1981) 3226-3237.

[205] M. Chen, D.W. Goodman, Catalytically active gold on ordered titania

supports, Chem. Soc. Rev., 37 (2008) 1860-1870.

[206] J.A. Rodriguez, G. Liu, T. Jirsak, J. Hrbek, Z. Chang, J. Dvorak, A.

Maiti, Activation of Gold on Titania: Adsorption and Reaction of SO2 on

Au/TiO2(110), J. Am. Chem. Soc., 124 (2002) 5242-5250.

[207] C.M. Goodman, C.D. McCusker, T. Yilmaz, V.M. Rotello, Toxicity of

gold nanoparticles functionalized with cationic and anionic side chains,

Bioconjugate Chem., 15 (2004) 897-900.

[208] J.A. odr guez, J. Evans, J.s. Graciani, J.-B. Park, P. Liu, J. Hrbek, J.F.

Sanz, High Water-Gas Shift Activity in TiO2(110) Supported Cu and Au

Nanoparticles: Role of the Oxide and Metal Particle Size, J. Phys. Chem. C.,

113 (2009) 7364-7370.

[209] P.P. Gan, S.H. Ng, Y. Huang, S.F.Y. Li, Green synthesis of gold

nanoparticles using palm oil mill effluent (POME): A low-cost and eco-

friendly viable approach, Bioresour. Technol., 113 (2012) 132-135.

[210] L.M. Liz-Marzán, M. Giersig, P. Mulvaney, Synthesis of Nanosized

Gold-Silica Core-Shell Particles, Langmuir, 12 (1996) 4329-4335.

[211] A.L. Aden, M. Kerker, Scattering of Electromagnetic Waves from Two

Concentric Spheres, J. Appl. Phys., 22 (1951) 1242-1246.

Publications

171

Publications and Manuscripts in Preparation

1. P.P. Gan, S.H. Ng, Y. Huang, S.F.Y. Li, Green synthesis of gold nanoparticles

using palm oil mill effluent (POME): A low-cost and eco-friendly viable approach,

Bioresour. Technol., 113 (2012) 132-135.

2. P.P. Gan, S.F.Y. Li, Potential of plant as a biological factory to synthesize gold

and silver nanoparticles and their applications, Rev. Environ. Sci. BioTechnol. 11

(2012) 169-206.

3. P.P. Gan, S.F.Y. Li, Biosorption of elements, Element Recovery and

Sustainability, RSC Green Chemistry, Chapter 4, publication in progress.

4. P.P. Gan, S.F.Y. Li, Efficient removal of Rhodamine B using a rice hull-based

silica supported iron catalyst by Fenton-like process, Chem. Eng. J., accepted.

5. F. Liu, P.P. Gan, H. Wu, W.S. Woo, E.S. Ong, S.F. Li, A combination of

metabolomics and metallomics studies of urine and serum from

hypercholesterolaemic rats after berberine injection, Anal. Bioanal. Chem., 403

(2012) 847-856.

6. P.P. Gan, S.F.Y. Li, Biosynthesis of gold nanoparticles using palm oil mill

effluent (POME), 4th International Conference on Challenges in Environmental

Science & Engineering, Taiwan, Sep 2011.

7. P.P. Gan, S.F.Y. Li, Potential of plant as a biological factory to synthesize gold

and silver nanoparticles and their applications, 4th International Conference on

Challenges in Environmental Science & Engineering, Taiwan, Sep 2011.

Publications

172

8. P.P. Gan, S.F.Y. Li, Biosorption of Cd(II) and Hg(II) from aqueous solutions

using palm oil mill effluent (POME) as a low-cost Biosorbent, 7th

Singapore

International Chemical Conference, Singapore, Dec 2012.

Appendices

173

Appendix 1

Figure S1 Schematic diagram of oil extraction from oil palm and POME

generation (dashed line represents byproduct/waste stream). Reproduced with

permission from reference [27].

Appendices

174

Appendix 2

Table S1.1 Structures of various dye classes. Reproduced with permission from

reference [1].

Class Structure Representative

dye Structure

Acridine

Acridine O

Azo

Amido B

Diarylmethane

Auramine O

Anthraquinone

Carmine

Triarylmethane

Malachite green

Nitro

Naphthol Y

Xanthene

RhB

Quinone-imine

Safranin O