table of contents - corecore.ac.uk/download/pdf/48676506.pdfdeclaration page i declaration page i...
Post on 03-Aug-2020
0 Views
Preview:
TRANSCRIPT
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
Chapter 4
108
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
Chapter 4
109
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.
Chapter 4
110
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
Chapter 4
111
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.
Chapter 4
112
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.
Chapter 4
113
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).
Chapter 4
114
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
Chapter 4
115
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].
Chapter 4
116
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.
Chapter 4
117
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.
Chapter 4
118
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.
Chapter 4
119
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.
Chapter 5
120
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
Chapter 5
121
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].
Chapter 5
122
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
Chapter 5
123
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
Chapter 5
124
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).
Chapter 5
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-
Chapter 5
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
Chapter 5
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.
Chapter 5
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.
Chapter 5
129
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.
Chapter 5
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).
Chapter 5
131
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
Chapter 5
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
top related