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INTEGRATED ADSORPTION, OXIDATION AND BIODEGRADATION
FOR TREATING EMERGING CONTAMINANTS IN WASTEWATER AND WATER
A THESIS SUBMITTED TO THE GRADUATE DIVISION OF THE UNIVERSITY OF HAWAI‘I AT MĀNOA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
MASTER OF SCIENECE
IN
MOLECULAR BIOSCIENCES AND BIOENGINEERING
DECEMBER 2014
By
Jihyun R. Kim
Thesis Committee:
Eunsung Kan, Chairperson Jon-Paul Bingham
Soojin Jun
Keyword: Emerging contaminants, advanced oxidation, UV, biochar, activated carbon
i
DEDICATION
I dedicate this thesis to my family, whose support and love have provided me
with the foundation on which I will continue to build upon. Specifically, I would like to
thank my parents, uncle and aunt, my brothers and sister in-laws and my husband.
Without them my success would be impossible, and my accomplishments would just be
dreams.
ii
ACKNOWLEDGMENTS
I would like to express my appreciation to everyone who has supported and
encouraged me throughout my Master’s program.
First, I would like to thank Professor Eunsung Kan for providing guidance and
trust. I would like to thank my committee members: Dr. Soojin Jun, Dr. J-P Bingham for
their continuous support and helpful advice, and give a special thank you to Dr. Harry
Ako for his encouragement and positive feedback. Also, I would like to express my
gratitude for Dr. Hye-Ji Kim and Dr. Dominic Glover for valuable suggestions for my
research and career.
I would like to thank all of my lab mates: Stuart Watson, Vincent Cleveland and
Donghee Hoh for providing me with unconditional support and help whenever needed.
For all of those who helped and were not mentioned, please do not worry. You have not
been forgotten.
iii
ABSTRACT
Combination of adsorption and chemical or photocatalytic pretreatment with
subsequent biodegradation demonstrates high potential for treating wastewater
containing emerging contaminants (ECs). Removal of ECs, such as endocrine
disrupting compounds and pharmaceuticals in wastewater-treatment plants was found
to be rather low and persistent. Advanced oxidation processes (AOPs) are used widely
for the removal of organic pollutants including ECs not treated by conventional
techniques owing to their high chemical stability and/or low biodegradability. Although
AOPs for complete mineralization is usually costly, its combination with biological
treatment can reduce operating costs and achieve complete mineralization of organic
pollutants. The chemical or photocatalytic degradation of bisphenol A (BPA), a
representative endocrine disruptor and sulfamethoxazole (SMX), a representative
pharmaceutical, was carried out in aqueous suspension using either activated carbon
supported iron catalyst or biochar supported TiO2 photocatalyst under UV irradiation.
The photocatalytic pre-treatment using biochar-TiO2 led to partial oxidation of
biologically persistent part of BPA or SMX to produce biodegradable products.
iv
TABLE OF CONTENTS Page
DEDICATION .................................................................................................................... i ACKNOWLEDGMENTS .................................................................................................. ii ABSTRACT ..................................................................................................................... iii LIST OF TABLES .......................................................................................................... vii LIST OF FIGURES ........................................................................................................ viii CHAPTER 1. Introduction .............................................................................................. 1
1.1. Project overview ................................................................................................... 1
1.2. Objectives ............................................................................................................ 2
CHAPTER 2. Heterogeneous Oxidation of Methylene Blue with Surface-Modified Iron-Amended Activated Carbon .................................................................................. 3
2.1. Abstract ................................................................................................................ 3
2.2. Introduction .......................................................................................................... 3
2.3. Materials and methods ......................................................................................... 5
2.3.1. Adsorbent and adsorbate ............................................................................. 5
2.3.2. Preparation of Fe-amended activated carbon .............................................. 6
2.3.3. Batch adsorption studies .............................................................................. 6
2.3.4. Effect of pH .................................................................................................. 7
2.3.5. Adsorption kinetics ....................................................................................... 8
2.3.6. pH, point of zero charge ............................................................................... 9
2.3.7. Heterogeneous Fenton oxidation of the MB-saturated Fe-GAC .................. 9
2.3.8. Effect of semi-continuous feeding of H2O2 on oxidation of MB in the Fe-
GAC ............................................................................................................ 10
2.3.9. Analytical methods ..................................................................................... 10
2.4. Results and discussion ...................................................................................... 11
2.4.1. Adsorption of MB onto the GACs ............................................................... 11
2.4.2. Heterogeneous Fenton oxidation of the MB in the Fe-GAC ....................... 13
2.5. References ......................................................................................................... 15
CHAPTER 3. Heterogeneous Photo-Fenton Oxidation of Methylene Blue Using CdS-Carbon Nanotube/TiO2 under visible light ......................................................... 26
v
3.1. Abstract .............................................................................................................. 26
3.2. Introduction ........................................................................................................ 26
3.3. Materials and methods ....................................................................................... 29
3.3.1. Chemicals and reagents ............................................................................ 29
3.3.2. Preparation and characterization of CdS/MWCNT-TiO2 ............................ 29
3.3.3. Visible light mediated photo-Fenton oxidation of methylene blue using the
CdS/MWCNT-TiO2 photocatalyst ............................................................... 30
3.3.4. Analytical methods ..................................................................................... 32
3.4. Results and discussion ...................................................................................... 33
3.4.1. Characterization of the CdS/MWCNT-TiO2 ................................................ 33
3.4.2. Photo-Fenton oxidation of methylene blue using CdS/MWCNT-TiO2 under
visible light .................................................................................................. 33
3.5. Acknowledgement .............................................................................................. 37
3.6. References ......................................................................................................... 37
CHAPTER 4. Effect of Temperatures on Adsorption and Oxidative Degradation of Bisphenol A in an Acid-Treated Iron-Amended Granular Activated Carbon .......... 58
4.1. Abstract .............................................................................................................. 58
4.2. Introduction ........................................................................................................ 58
4.3. Materials and methods ....................................................................................... 60
4.3.1. Chemicals and reagents ............................................................................ 60
4.3.2. Preparation of acid-treated Fe-amended activated carbon ........................ 60
4.3.3. Batch adsorption of BPA onto the Fe-GAC ................................................ 61
4.3.4. Fenton oxidation-driven regeneration of the BPA-spent Fe-GAC .............. 63
4.3.5. Analytical methods ..................................................................................... 65
4.4. Result and discussion ........................................................................................ 66
4.4.1. Isotherm, kinetics and thermodynamic analysis for adsorption of BPA onto
the Fe-GAC ................................................................................................ 67
4.4.2. Fenton oxidation of BPA in the Fe-GAC .................................................... 69
4.4.3. Comparative analysis of temperature-dependent diffusion, desorption and
Fenton oxidation of BPA in the Fe-GAC .................................................... 71
4.5. Acknowledgements ............................................................................................ 72
vi
4.6. References ......................................................................................................... 72
CHAPTER 5. UV Photocatalytic Oxidation of Sulfamethoxazole Using TiO2
Supported on Biochar .................................................................................................. 91
5.1. Abstract .............................................................................................................. 91
5.2. Introduction ........................................................................................................ 91
5.3. Materials and method ......................................................................................... 94
5.3.1. Chemical and reagent ................................................................................ 94
5.3.2. Preparation and characterization of biochar-TiO2 ...................................... 94
5.3.3. UV light mediated photocatalytic oxidation of SMX using the biochar-TiO2
photocatalyst .................................................................................................. 95
5.3.4. Analytical methods ..................................................................................... 95
5.4. Results and discussion ...................................................................................... 96
5.4.1. Characterization of the biochar-TiO2 .......................................................... 96
5.4.2. UV photocatalytic oxidation of SMX ........................................................... 97
5.5. Acknowledgement .............................................................................................. 99
5.6. References ......................................................................................................... 99
Chapter 6. Conclusion ............................................................................................... 113
APPENDIX. Proof of publications ............................................................................. 116
vii
LIST OF TABLES
Table Page
2.1. Langmuir and Freundlich isotherm model coefficient for adsorption of MB
on GACs …...…………………………………………………………………………....18
2.2. Surface charge of the Fe-GAC and MB at various pH ..…………………………….19
2.3. pH, pzc of the virgin, acid-treated and acid-treated iron amended GAC ..………….20
2.4. Kinetic parameters for MB adsorption onto Fe-amended GAC …..………………..21
3.1. Summary of heterogeneous photo-Fenton and photo-Fenton like oxidation of
azo dyes under visible light irradiation ….……………………………………………46
3.2. EDX results of CdS/MWCNT-TiO2 before and after photo-Fenton reaction
under visible light irradiation ..………………………………………………………….47
3.3. Summary of optimal [H2O2]: [methylene blue] molar ratio in Fenton and
photo-Fenton treatment of methylene blue ….…….…………………………………48
4.1. Freundlich and Langmuir isotherm model coefficients for adsorption of BPA
on the GACs ..……………………………………………………………………………79
4.2. Adsorption kinetic parameters for BPA adsorption onto Fe-amended GAC ..…….80
4.3. BPA diffusivity at various temperatures …..…………………………………………..81
4.4. Thermodynamic parameters for BPA adsorption onto Fe-amended GAC ..………82
4.5. Thiele-modulus analysis for the Fenton oxidation of BPA-spent GAC at
increasing temperatures ……………..……………………………………………...…83
viii
LIST OF FIGURES
Figure Page
2.1. The molecular structure of methylene blue …...………………………………………22
2.2. Adsorption kinetics of MB on the Fe-GAC (a) pseudo-first-order kinetics and (b)
pseudo-second-order kinetics ..…..…………………………………………………....23
2.3. Effects of iron loading on oxidation of the MB-spent Fe-GAC with the fixed H2O2
dose ..…………………………………………………………………………..…………24
2.4. Effect of H2O2 on the oxidation of MB in the Fe-GAC ...……………………………...25
3.1. TEM (a) and SEM (b) images of the CdS/MWCNT-TiO2 …………………………….49
3.2. XRD patterns of (a) TiO2, (b) CdS/TiO2 and (c) CdS/MWCNT-TiO2 ..……………..50
3.3. Removal of methylene blue and TOC using photo-Fenton, dark Fenton oxidation
and photocatalysis ..…………….……………………..……...………………………..51
3.4. Effects of pH on adsorption and photo-Fenton oxidation of aqueous methylene
blue ……….……………………………………………………………………………...52
3.5. Effects of [Fe3+]/[H2O2] on photo-Fenton oxidation of methylene blue .…………...53
3.6. Effects of [H2O2]/[methylene blue] on photo-Fenton oxidation of methylene
blue …….………………………………………………………………………………...54
3.7. Reusability of the CdS/MWCNT-TiO2 in the photo-Fenton oxidation of methylene
blue …….………………………………………………………………………………...55
3.8. Proposed mechanisms of the photodegradation of methylene blue on
CdS/MWCNT-TiO2 composites under visible light irradiation ….………………….56
3.9. Effects of formate and ethanol on scavenging on photo-Fenton oxidation of
methylene blue ………………….………………………………………………………57
4.1. The proposed mechanisms in Fenton oxidation of BPA (contaminant) in the acid-
treated Fe-amended granular activated carbon (Fe-GAC) ……….…………….….84
4.2. Atomic chemical composition of Fe in acid-treated and untreated Fe-amended
GAC ……….…………………………………………………………………………..…85
4.3. Adsorption kinetics of BPA onto the Fe-GAC at various temperatures …..……….86
4.4. Effect of [H2O2]: [BPA] on BPA removal .……………………………………………..87
ix
4.5. Effect of temperature on Fenton oxidation of BPA in the GAC ..…………………...88
4.6. Proposed pathway for the Fenton oxidation of BPA ..……………………………….89
4.7. Temperature-dependent relative increase in calculated diffusion of BPA, measured
values of BPA desorption + diffusion, and Fenton-driven oxidation rate of BPA in
the Fe-GAC ……………………………………………………………………………..90
5.1. SEM images of (a) biochar and (b) biochar-TiO2 ..…………………………………105
5.2. XRD patterns of (a) TiO2 (b) biochar and (c) biochar-TiO2 ..………………………106
5.3. Removal of SMX and COD using photolysis and photocatalysis ..……………….107
5.4. Effect of the catalyst loading on removal of SMX and COD .……………………..108
5.5. Effect of sodium nitrate on removal of SMX ….…………………………………….109
5.6. Effect of sodium bicarbonate on removal of SMX .………………………………...110
5.7. Effect of reaction time on removal of SMX .………………………………………...111
5.8. The result of oxygen uptake rate of byproduct formed from SMX+UV and
SMX+ UV+Biochar-TiO2 …………………………………………...……...………….112
1
CHAPTER 1. Introduction 1.1. Project overview
The aim of this study was to develop a cost effective, highly efficient and
sustainable hybrid (adsorption and oxidation or biodegradation) treatment system to
eliminate emerging contaminants including but not limited to endocrine disrupting
compounds and antibiotics from wastewater and water. This system could be used as
an alternative tertiary wastewater treatment process.
Chapter 1 provides the overview and objective of the project. Chapter 2 and 3
present the proof of concept on the hybrid system of adsorption and chemical oxidation
or photocatalytic oxidation using an industrial dye compound, methylene blue. Prior to
elimination of emerging contaminants using (photo) catalyst, methylene blue is used to
confirm the treatment system. Chapter 4 focuses on adsorption and Fenton oxidation of
bisphenol A which is known as an endocrine disrupting compound using the acid
treated iron amended granular activate carbon. Chapter 5 presents adsorption and
photocatalytic oxidation of sulfamethoxazole (an antibiotic).
Chapter 2 has been published in a peer-reviewed journal, and chapter 3 and
chapter 4 is accepted for publication (articles in press). Chapter 5 will be submitted for
publication.
The following is a list of publications:
[1]. Jihyun R. Kim and Eunsung Kan (2014). Effects of temperature on adsorption and
oxidative degradation of bisphenol A in a surface modified iron-amended granular
activated carbon. Chemical Engineering Journal (In press).
[2]. Jihyun R. Kim and Eunsung Kan (2014). Heterogeneous Photo-Fenton oxidation of
methylene blue using CdS-Carbon nanotube/TiO2 under visible light in Journal of
Industrial and Engineering Chemistry (In press).
2
[3]. Jihyun R. Kim., Brenden Santiano., Hyosang Kim., and Eunsung Kan (2013).
Heterogeneous Oxidation of Methylene Blue with Surface-Modified Iron-Amended
Activated Carbon. Journal of Analytical Chemistry, 4, 115 – 122.
1.2. Objectives
1. Synthesis, modification and characterization of carbon based (photo) catalysts
2. To Understand adsorption characteristics – adsorption of emerging
contaminants onto carbon based (photo) catalysts
3. To Investigate adsorption and (photo) chemical oxidation or biodegradation of
emerging contaminants using carbon based (photo) catalysts
3
CHAPTER 2. Heterogeneous Oxidation of Methylene Blue with Surface-Modified Iron-Amended Activated Carbon
2.1. Abstract
The present study aims to develop effective adsorption and oxidation of synthetic
dye in wastewater by using the newly synthesized iron-amended activated carbon.
Recently synthetic dye-containing wastewater has gained more attention due to its
mass discharge, high toxicity and low biodegradation. For enhancing adsorption of dye
and oxidative regeneration of dye-exhausted activated carbon, the novel amendment of
iron-deposited granular activated carbon (GAC) was developed. It was to amend ferrous
ion onto the acid-pretreated GAC when pH of iron solution was higher than the pH at
point of zero charge (pH, pzc) of the GAC. Methylene blue (MB) in water was adsorbed
onto the acid-treated iron-amended GAC (Fe-GAC) followed by single or multiple
applications of H2O2. Batch experiments were carried out to study the adsorption
isotherm and kinetics indicating adsorption of MB onto the Fe-GAC followed Langmuir
isotherm and the pseudo-second order kinetics. The Fe-GAC showed the maximum
adsorption capacity (qm) of 238.1 ± 0.78 mg/g which was higher than the virgin GAC
with qm of 175.4 ± 13.6 mg/g at 20°C, pH 6 and the initial concentration of 20-200 mg/L.
The heterogeneous Fenton oxidation of MB in the Fe-GAC revealed that increasing the
H2O2 loading from 7 to 140 mmol H2O2/mmol MB led to enhancing the oxidation
efficiency of MB in the GAC from 62.6% to 100% due to the increased generation of
hydroxyl radicals. Further enhancement of oxidation of MB in the Fe-GAC was made by
the multiple application of H2O2 while minimizing OH radical scavenging often occurring
at high concentration of H2O2. Therefore, the acid-treated iron-amended GAC would
provide excellent adsorption capacity for MB and high oxidation efficiency of MB in the
GAC with multiple applications of H2O2 and optimum iron loading.
2.2. Introduction Synthetic dyes are widely used for textile, pulp and paper, plastic, food, cosmetic
and pharmaceutical industries [1]. With mass production of synthetic dyes and mass
discharge of synthetic dye-containing wastewater, effective treatment of dye-containing
4
wastewater has gained more attention for the past decades [2]. Degradation of synthetic
dyes is difficult because of their complex aromatic structure and nature of the lead
compounds. Besides, the intermediate and byproducts of most of dye chemicals are
considered to be harmful to aquatic organisms due to their toxicity and carcinogenic
activity [3]. Therefore, efficient and cost-effective treatment processes of dye-containing
wastewater need to be developed [4]. Adsorption using granular activated carbon as one of most reliable methods has
been used to remove dyes in wastewater. It is simple, effective and independent from
toxicity of dye chemicals. However, regeneration of dye-exhausted activated carbon
determines the overall operating costs because the dye-spent activated carbon needs
to be regenerated to achieve its re-adsorptive capacity. For most of cases involving
GAC regeneration, the spent GAC is thermally regenerated either on-site or transported
to a thermal regeneration facility and regenerated off-site. The current thermal
regeneration method often leads to significant deterioration of the carbon pore structure,
specific surface area and functionality which influences on re-adsorptive capacity of
contaminant-spent activated carbon [5]. Alternative to the thermal regeneration, Fenton oxidation-driven regeneration of
the spent GAC is a treatment option under development to regenerate the GAC on-site
and in situ. It has shown excellent performance to oxidize various contaminants
including Methyl Tertiary Butyl Ether (MTBE), bezene, toluene, ethylenbenzene and
xylene (BTEX), trichloroethylene (TCE) and chlorobenzene from the activated carbon
[6,7]. The Fenton oxidation-driven regeneration relies on heterogeneous Fenton
oxidation which oxidizes contaminants at the GAC by using OH radicals generated by
reaction between H2O2 and iron immobilized at the GAC [7]. Thus, it leads to effective
oxidation of contaminants at surface of the GAC with little generation of iron sludge
which is different from mass production of iron sludge by homogeneous Fenton
oxidation. For heterogeneous Fenton oxidation, iron amendment onto the GAC is of
prime importance to determine oxidation efficiency of contaminants in the GAC since
more uniform iron amendment onto the GAC minimizes intraparticle diffusion limitation,
the rate-limiting step during heterogeneous Fenton oxidation in the GAC [7].
5
For this study, the acid-treated iron-amended GAC (Fe-GAC) was synthesized by
amending Fe2+ onto the acid-pretreated GAC at the conditions that the pH of Fe2+
solution was higher than the pH at point of zero charge (pH, pzc) of the GAC. As
reported by Kan and Huling [8], the acid-treatment of the GAC was conducted to
increase acidic surface groups containing oxygen and to lower pH, pzc of the GAC.
When the acid-treated GAC was suspended in the Fe2+ solution at the conditions that
the pH of ferrous solution was higher than the pH, pzc of the acid-treated GAC, the more
negatively charged surface of the acid-treated GAC enhanced electrostatic attraction
with Fe2+ for making more uniform Fe deposition into the GAC [7]. The more uniform Fe
deposition into the GAC would help effective oxidation of the contaminants in the GAC
with steady diffusion and reaction of H2O2 into the pores of the GAC.
Therefore, the present study characterized the adsorption capacity and the
oxidation efficiency of the acid-treated Fe-GAC for treating MB-contaminated water. The
adsorption isotherm and kinetics using the Fe-GAC were analyzed for understanding
the major mechanisms associated with adsorption of MB onto the Fe-GAC. The
oxidation efficiency of MB in the Fe-GAC by heterogeneous Fenton oxidation was
investigated at various iron and H2O2 loading.
2.3. Materials and methods 2.3.1. Adsorbent and adsorbate
The GAC (URV, 8 × 30 mesh, Calgon Carbon Corp., Pittsburgh, PA) was derived
from bituminous coal and activated in a manner to minimize H2O2 reactivity (Harrision,
Calgon Carbon, personal communication). The GAC was rinsed with deionized (DI)
water, dried in an oven at 105 °C, and stored in a desiccator until used. The surface
area and pore volume of the GAC as received was 1290 m2/g and 0.64 mL/g,
respectively [6]. Methylene blue (Figure 2.1, Fisher Scientific, Waltham MA), a cationic
dye, which is difficult to be degraded in a natural environment, was chosen as the model
adsorbate. A standard stock solution of 1000 mg/L was prepared and suitably diluted to
the required initial concentration.
6
2.3.2. Preparation of Fe-amended activated carbon A stock solution of ferrous chloride was prepared immediately before the Fe
solution was amended to the GAC by dissolving ferrous chloride (FeCl2 ·4H2O) into the
DI water (50 mg/L as Fe2+). The acid-treated GAC was prepared by treating the 10 g
virgin GAC with nitric acid at pH 3 for 4 d for increasing the acidic surface oxides and
lowering its pH, pzc (pH at point of zero charge). A pH, pzc is the pH at which positive and
negative surface charges are equal and GAC surface has a net charge of zero.
The acidic treatment of the GAC was to enhance the electrostatic interaction
between the GAC and Fe2+ for nearly uniform iron deposition in the GAC. The 6.3-187.5
mL of 50 mg L-1 Fe2+ was added to the 0.5 g the acid-treated GACs for iron amendment
onto the GAC. After the iron amendment onto the acid-treated GACs, the GACs were
rinsed three times with DI water to eliminate the chloride residual in the GACs and dried
in an oven at 80 °C for 24 h.
2.3.3. Batch adsorption studies The batch experiments for adsorption isotherm were conducted by suspending
25 mg of the Fe-GAC in a series of flasks containing 50 mL of methylene blue solutions
at the concentration of 0.02 – 0.2 g/L. The initial pH of MB solution (pH = 6.0) was
adjusted by adding 0.1 M HNO3 or 0.1 M NaOH. The initial concentrations of MB were
obtained by measuring O.D. at 663 nm (λmax) using BioSpec-mini UV-Vis
spectrophotometer (Shimadzu, Carlsbad CA). The Fe-GAC used for the adsorption
isotherm experiments contained 6 mg Fe/g GAC. The adsorption process was carried
out by agitating at the constant speed (200 rpm) at room temperature (20°C) for 72 h to
ensure equilibrium was reached. The post-sorption MB solution was sampled after
equilibrium (>3 d) in replicate and analyzed. The Differences between initial and final
concentrations were used to calculate the mass of MB adsorbed to the GAC. The MB
uptake at equilibrium, qe (mg/g), was calculated by Eq. (1):
𝑞! = !!!!! !
! (1)
7
where C0 and Ce (mg/L) are the liquid-phase concentration of dye at initial and at
equilibrium, respectively. V (L) is the volume of the solution, and W (g) is the mass of
dry adsorbent used. All samples were centrifuged prior to analysis to minimize the
interference of carbon fines with the analysis.
For understanding the adsorption mechanism and capacity of the GAC, the Langmuir
and Freundlich isotherm models were used to interpret the batch isotherm data. The
Langmuir equation [9] is valid for monolayer adsorption on a surface with a finite
number of identical sites while Freundlich model [10] is an empirical equation based on
adsorption on a heterogeneous surface. The linear form of Freundlich and Langmuir
isotherms are shown in Eqs. (2) and (3), respectively:
Freundlich isotherm: 𝑙𝑛𝑞! = 𝑙𝑛𝑘! + (1 𝑛)𝑙𝑛𝐶! (2)
Langmuir isotherm: 𝐶! 𝑞! = 1 𝐾! 𝑞!"# + 𝐶! 𝑞!"# (3)
where lnkf is roughly a measure of the maximum adsorption capacity and 1/n is an
indicator of adsorption effectiveness; qmax is the maximum adsorption capacity (mg/g)
corresponding to complete monolayer coverage of the surface, qe is the amount of dye
adsorbed per unit mass of adsorbent (mg/g), Ce is the liquid-phase concentration of dye
at equilibrium and KL is the Langmuir constant (L/g) related to the sorption/desorption
energy.
2.3.4. Effect of pH The effect of pH on the color removal and adsorption capacity of MB was
analyzed over the pH range from 3 to 11. The pH was adjusted using 0.1 M HCl and 0.1
M NaOH solutions. The dye solution (60 mL) at concentration of 1000 mg/L with 0.5 g of
acid-treated Fe amended GAC was shaken for 72 h at room temperature (20°C) on a
shaker at a constant speed of 200 rpm. The samples were then centrifuged and
analyzed using a UV spectrophotometer.
8
2.3.5. Adsorption kinetics Adsorption kinetic experiments were carried out by suspending 0.5 g of the Fe-
GAC in 60 mL of each MB solution at initial concentrations of 80 mg/L to 1000 mg/L, a
pH 6 and 20℃ for 24 h under mixing. Aliquots of 0.1 mL to 1 mL depending on the
solution of concentration were withdrawn at different time intervals, centrifuged and
analyzed for the MB concentration using a UV spectrophotometer. The kinetic of the
adsorption were analyzed using two different kinetic models: the pseudo-first order and
pseudo-second order. The pseudo-first order kinetic model is given by Lagergren [11]:
!!!!"= 𝑘!(𝑞! − 𝑞!) (4)
where q and qe represent the amount of dye adsorbed (mg/g) at any time t and at
equilibrium time, respectively, and k1 represents the sorption rate constant (min-1).
Integrating Eq (4). with respect to boundary conditions q=0 at t=0 and q=qt at t=t, then
Eq. (4) becomes:
ln 𝑞! − 𝑞! = 𝑙𝑛𝑞! − 𝑘!𝑡 (5)
where qe and qt (mg/g) are the amounts of adsorbate adsorbed at equilibrium and at any
time, t (min), respectively, and k1 (min-1) is the adsorption rate constant. Thus the rate
constant k1 (min-1) and qe can be calculated from the plot of ln (qe-qt) versus time t.
The pseudo-second-order equation [12] based on equilibrium adsorption is expressed
as:
!!!!"= 𝑘!(𝑞! − 𝑞!)! (6)
where k2 (g/mg min) is the rate constant of second-order adsorption, qe and q represent
the amount of dye adsorbed (mg/g) at equilibrium and at any time t. Separating the
variables in Eq (6). gives:
9
!!!(!!!!!)!
= 𝑘!𝑑𝑡 (7)
Integrating Eq (7). with respect to boundary conditions t=0 to t=t and q=0 and q=qe
gives
!!!= !
!!!!!+ !
!!𝑡 (8)
The rate constant of second-order adsorption (k2) and qe will be evaluated from the
linear plot t/qt versus t.
2.3.6. pH, point of zero charge
The pH at point of zero charge (pH, pzc) of the GAC was determined using the pH
drift method [7]. 50 mL of DI water amended with 0.01 M NaCl was placed in the 100
mL amber vials and sparged with N2 (200 mL/min; 10-15 min) to eliminate CO2 and to
stabilize pH. The pH was adjusted from pH 2 to pH 11 in a series of vials by adding
either HCl or NaOH while purging the headspace with N2. 0.15 g of the GAC was added
and the vial was capped immediately. The final pH (pH, final) was measured in each of
the vials after 48 h and plotted versus the initial pH (pH, initial). The pH, pzc was
determined graphically at the intersection of pH, final and the line pH, final= pH, initial.
2.3.7. Heterogeneous Fenton oxidation of the MB-saturated Fe-GAC The effect of various iron loadings (0.5 to 15 mg Fe/g GAC) on the oxidation
efficiency of MB in the Fe-GAC was investigated. The oxidation of the MB-saturated Fe-
GAC was performed with a single injection of 3.0 mL of 30% H2O2 (52.9 mmol H2O2/g
GAC; Sigma Aldrich, MO) at pH of 3.0-3.5. The pH of solution before adding H2O2 was
between 3.0-3.5. Since ·OH scavenging by high concentrations of H2O2 is a probable
source of overall oxidation inefficiency [13], the multiple applications of H2O2 were made
to minimize the initial H2O2 concentration and ·OH scavenging. H2O2 concentrations
were monitored over time under complete mixing condition (H2O2, initial of 14.3 g/L). The
oxidation efficiency of MB in the GAC was evaluated by measuring the MB in the GAC
10
after Fenton oxidation. Similarly, the effect of H2O2 concentration on oxidation of MB in
the Fe-GAC was investigated when various load of H2O2 (7 mmol to 140 mmol H2O2/
mmol MB) at the fixed iron content in the Fe-GAC (5 mg Fe/g GAC) for determining the
optimum load of H2O2.
The residual amount of MB in the Fe-GAC after Fenton oxidation was evaluated
by extracting the MB in the Fe-GAC with 10-20 mL methanol for 72 h. The MB
desorption and diffusion rate from MB-spent GAC was evaluated using the fill and draw
method. This involved post-sorption applications of the DI water (60 mL) to MB-spent
GAC (0.5 g) and measuring MB in solution at a different time interval. The rates of
desorption was calculated as the mass of MB desorbed from the GAC divided by the
mass of GAC and time period of desorption. In all cases tested, most of MB in the GAC
was not released to the aqueous phase and the concentration of MB in the aqueous
phase was far below the equilibrium MB concentration.
2.3.8. Effect of semi-continuous feeding of H2O2 on oxidation of MB in the Fe-GAC In order to investigate the effect of the semi-continuous feeding of H2O2 on
increasing MB oxidation efficiency in the Fe-GAC, one fourth of H2O2 amount was
added at four different points of time at every 12h. The initial H2O2 concentration (28
mmol H2O2/ mmol MB) was selected from the previous experiment with the same iron
load which was 6 mg Fe/g GAC. The single point H2O2 injection from the previous
section and semi-continuous feeding of H2O2 was evaluated the effectiveness of the
H2O2 injections for oxidation of the MB in the Fe-GAC.
2.3.9. Analytical methods
MB concentration in the supernatant solution after and before adsorption was
determined using UV-Vis spectrophotometer at 663 nm. The filtered samples using 0.2
µm membrane filters were measured for H2O2 (n=3) using a modified peroxytitanic acid
colorimetric procedure with a detection limit of 0.1 mg/L [7]. Iron was measured in the
GAC slurry solution using the Phenanthroline Method [7].
11
2.4. Results and discussion 2.4.1. Adsorption of MB onto the GACs 2.4.1.1. Adsorption isotherm analysis
In order to estimate the adsorption capacity of GACs for MB and to optimize the
design of adsorption system, it is important to analyze the adsorption equilibrium data.
In the present study, the Freundlich and Langmuir isotherms that are the most
commonly used isotherms were applied to investigate the overall adsorption efficacy
and to characterize the MB adsorption process in the system using Fe-GAC. The
isotherm parameters of the Freundlich and Langmuir models for the adsorption of MB
on virgin GAC and the Fe-GAC obtained at temperature of 20 °C after 72 h, are
summarized in Table 2.1. The results indicated that the maximum adsorption capacity
of the Fe-GAC (qm = 238.1m ± 0.78 mg/g) in this study was higher compared with the
virgin GAC (qm = 175.4 ± 13.6 mg/g). The result indicates the maximum adsorption
capacity of the activated carbon prepared in this study was comparable with the works
done by other studies on adsorption of MB in GAC (qm= 240-299 mg/g) [14,15]. Also,
the analysis of the batch isotherm data supported that Langmuir isotherm model is the
appropriate model for adsorption of MB onto the GAC indicating homogeneous
monolayer adsorption of MB onto the site of the GAC.
2.4.1.2. Effect of initial MB solution pH on adsorption of methylene blue on Fe-GAC
The pH of solution at which adsorption occurs is found to influence the extent of
adsorption. The pH of MB solution affects adsorption in that it governs the degree of
ionization of the acidic and basic functional groups in methylene blue (see Figure 2.1
for chemical structure of MB) [16]. The effect of initial pH of the MB solution on the
amount of MB adsorbed was studied by varying the initial pH with constant process
parameters. The experimental results indicated that the pH of initial MB solution did not
possess any significant influence on the adsorption of methylene blue. The adsorption
of MB at the initial pH of 3 to 11 showed nearly complete uptake of methylene blue with
the adsorption capacity of 123.9 to 135.1mg MB/g GAC (data not shown). This finding is
different from others’ investigation that initial pH value may enhance or depress
12
adsorption of dye on activated carbon [17]. As reported by others, MB adsorption
capacities are significantly improved at a higher solution pH [17,18]. This can be
explained by the electrostatic attraction (Table 2.2) between the positively charged MB
(solution pH > pKa, MB) and the negatively charged surface of activated carbon (solution
pH > pH, pzc, Table 2.3). Increasing solution pH increases the number of hydroxyl
groups thus, increases the number of negatively charged sites and enhances the
attraction between dye and adsorbent surface [19]. However, the nearly constant
adsorption capacity of Fe-GAC for MB over the pH range 3 to 11 in our study was an
indication that the adsorption of MB on the Fe-GAC depends on both the electrostatic
interaction and non-electrostatic such as van der Waals forces, hydrophobic interaction
and hydrogen bonding in this system. It can be explained by the fact that adsorption of
dye from aqueous solution on carbon is a complex interplay between non-electrostatic
and electrostatic interactions [20].
2.4.1.3. Adsorption kinetics In order to investigate the mechanisms of MB adsorption on the Fe-GAC, the
pseudo-first order and pseudo-second order kinetics models were used to fit the
experimental data. The kinetic parameters of the pseudo-first order and the pseudo-
second order models were calculated from the linearized eq. (5) and (6), respectively.
MB adsorption of the pseudo-second order kinetic fitting results is shown in Figure 2.2.
The kinetic parameters acquired from fitting the results are summarized in Table 2.4.
The correlation coefficient values for the pseudo-first order and pseudo-second order
kinetic models were found to be close to one; however, the experimental qe values for
the first-order kinetic model do not agree with the calculated ones whereas the pseudo-
second order show a good agreement between experimental and calculated qe values.
The applicability of the kinetic model to describe the adsorption process was further
validated by the normalized standard deviation, ∆𝑞!(%), which is defined as:
∆𝑞! = 100×∑[!!,!"# !!!,!"#
!!,!"#]!
(!!!) (9)
13
where N is the number of data points, qe, exp and qe, cal (mg/g) are the experimental and
calculated equilibrium adsorption capacity value, respectively [21]. The ∆𝑞! obtained for
the pseudo-first-order kinetic model was 38.53%, which is relatively high as compared
to the ∆𝑞! values of 7.04% obtained for the pseudo-second order kinetic model. Base
on the high correlation coefficient values and the low ∆𝑞! value, the pseudo second
order model adequately described the experimental data of the adsorption of MB onto
Fe-GAC. This suggested that the overall rate of the adsorption process was controlled
by chemical sorption or chemisorption, which involved valency forces through sharing or
exchange of electrons between the adsorbent and adsorbate [12]. The similar
phenomena have been observed in the adsorption of MB on the bamboo-based
activated carbon [22], the activated carbon prepared from rattan sawdust [23] and on
the activated carbon from waste biomass by sulfuric acid activation [24].
2.4.2. Heterogeneous Fenton oxidation of the MB in the Fe-GAC
Fenton and Fenton-like reaction for the treatment of wastewater are very efficient
in a pH range between 2.8 and 3 [25,26]. In the present study, adjusting the pH of the
MB-saturated Fe-GAC in the solution was not needed because the pH of solution after
MB saturation onto Fe-GAC was approximately 3. As the major reactions in Fenton
oxidation describe in Eq. (10) - (13), the excessive H2O2 and Fe2+ scavenge !OH that is
the major oxidant in Fenton oxidation used for non-specific oxidation of broad ranges of
contaminants.
Fe2+ + H2O2 → Fe3+ + OH- + !OH (10)
Contaminant + !OH → R! + H2O
→ further oxidation to final products (11)
!OH + H2O2 → HO2! + H2O (12)
!OH + Fe2+ → Fe3+ + OH- (13)
Thus, the heterogeneous Fenton oxidation of MB in the Fe-GAC was carried out at
various iron or H2O2 loading for figuring the optimum conditions for oxidation of MB in
the Fe-GAC. To use a heterogeneous catalytic system in industrial practice it is
important to evaluate the loss of catalyst from the support [26]. In this study, there was
no leaching of iron in the solution after all oxidation process at pH 3 and 20 ℃.
14
2.4.2.1. Effect of Fe loading on oxidation of MB in the Fe-GAC In order to investigate the effect of the Fe loading on the MB oxidation in the Fe-
GAC, the experiments were performed at the different Fe2+ loading of 0.5 to 15 mg Fe/g
GAC at the fixed concentration of H2O2 of 140 mmol H2O2/ mmol MB. Figure 2.3 shows
the effect of Fe loading on oxidation of MB in the Fe-GAC to analyze the 1st order rate
constant and half-life of H2O2. It indicated enhancing the rate constant for H2O2
consumption while reducing the half-life of H2O2 at increasing iron loading in the GAC.
Besides, the oxidation efficiency of MB in the Fe-GAC was close to 100% in all the
conditions of Fe loading (0.5 mg/g GAC to 15 mg/g GAC) due to the excessive initial
concentration of H2O2. Note that 36 moles of H2O2 are required to completely oxidize 1
mole of MB based on the balanced stoichiometric equation of oxidation of MB.
Therefore, the effect of H2O2 concentration on the oxidation of MB in the Fe-GAC was
investigated at the fixed loading of Fe (6 mg Fe/g GAC).
2.4.2.2. Effect of single and multiple H2O2 injection on MB oxidation Two sets of experiments with the single and the multiple application of the H2O2
to the MB-saturated Fe-GAC were designed to evaluate the effectiveness of the H2O2
application [27]. The initial hydrogen peroxide concentration was varied between 7 to
140 mmol H2O2/ mmol MB with 6 mg Fe/g GAC from the previous section. The effect of
the single application of H2O2 on MB oxidation is shown in Figure 2.4. The increase of
the H2O2 concentration from 7 to 140 mmol H2O2/ mmol MB led to an increase in the
MB removal efficiency from 62.6% to 100%, respectively. The selection of this optimum
concentration is important from the commercial point of view due to the high cost of
H2O2 [28].
For the multiple H2O2 application, one fourth of H2O2 amount was added at four
different points of time. The initial H2O2 loading of 28 mmol H2O2/ mmol MB was
selected from the previous experiment with 6 mg Fe/g GAC. It was lower than the
stoichiometric demand of H2O2 for the complete mineralization of MB (36 mmol of H2O2
per mmol of MB). The results showed that the MB oxidation efficiency of multiple H2O2
application was 84.1%, which was higher than that by a single H2O2 application which
15
was 71.6%. The multiple additions of H2O2 are more effective than the single addition of
H2O2 because scavenging of OH radicals due to the excessive H2O2 could be reduced
with the staged application of H2O2 [27].
2.5. References [1] S. Khorramfar, N. M. Mahmoodi, M. Arami and K. Gharanjig, Equilibrium and
kinetic studies of the cationic dye removal capability of a novel biosorbent
Tamarindus indica from textile wastewater, Coloration Technology, 126 (2010)
261-268.
[2] S. Khorramfar, N. M. Mahmoodi, M. Arami and H. Bahrami, Oxidation of dyes
from colored wastewater using activated carbon/hydrogen peroxide,
Desalination, 279 (2011) 183-189.
[3] H. J. Fan, H. Y. Shu and K. Tajima, Decolorization of acid black 24 by the
FeGAC/H2O2 process, Journal of Hazardous Materials, 128 (2006) 192-200.
[4] V. P. Santos, M. F. R. Pereira, P. C. C. Faria and J. J. M. Orfao, Decolourisation
of dye solutions by oxidation with H2O2 in the presence of modified activated
carbons, Journal of Hazardous Materials, 162 (2009) 736-742.
[5] S. G. Huling, P. K. Jones, W. P. Ela and R. G. Arnold, Fenton-driven chemical
regeneration of MTBE-spent GAC, Water research, 39 (2005) 2145-2153.
[6] S. G. Huling, P. K. Jones and T. R. Lee, Iron optimization for Fenton-driven
oxidation of MTBE-spent granular activated carbon, Environmental science and
technology, 41 (2007) 4090-4096.
[7] E. Kan and S. G. Huling, Effects of Temperature and Acidic Pre-Treatment on
Fenton-Driven Oxidation of MTBE-Spent Granular Activated Carbon,
Environmental Science and Technology, 43 (2009) 1493-1499.
[8] S. G. Huling, E. Kan and C. Wingo, Fenton-driven regeneration of MTBE-spent
granular activated carbon-Effects of particle size and iron amendment
procedures, Applied Catalysis B-Environmental, 89 (2009) 651-658.
[9] I. Langmuir, The adsorption of gases on plane surfaces of glass, mica and
platinum, Journal of American Chemical Society, 40 (1918) 1361-1403.
16
[10] H. Freundlich, Over the adsorption in solution, Journal of Physical Chemistry, 57
(1906) 385-470.
[11] S. Lagergren, About the theory of so-called adsorption of soluble substances,
Kungliga Svenska Vetenskapsakademiens Handlingar, 24 (1898) 1-39.
[12] Y. S. Ho and G. McKay, Pseudo-second order model for sorption processes,
Process Biochemistry, 34 (1999) 451-465.
[13] T. A. Kurniawan and W. H. Lo, Removal of refractory compounds from stabilized
landfill leachate using an integrated H2O2 oxidation and granular activated carbon
(GAC) adsorption treatment, Water Research, 43 (2009) 4079-4091.
[14] F. Raposo, M. A. De La Rubia and R. Borja, Methylene blue number as useful
indicator to evaluate the adsorptive capacity of granular activated carbon in batch
mode: Influence of adsorbate/adsorbent mass ratio and particle size, Journal of
Hazardous Materials, 165 (2009) 291-299.
[15] G. G. Stavropoulos and A. A. Zabaniotou, Production and characterization of
activated carbons from olive-seed waste residue, Microporous and Mesoporous
Materials, 82 (2005) 79-85.
[16] W. J. Weber, Physiochemical Properties for Water Quality Control, John Wiley
and Sons Inc, New York, (1972).
[17] E. N. El Qada, S. J. Allen and G. M. Walker, Adsorption of basic dyes from
aqueous solution onto activated carbons, Chemical Engineering Journal, 135
(2008) 174-184.
[18] S. B. Wang SB, Z. H. Zhu, A. Coomes, F. Haghseresht and G. Q. Lu, The
physical and surface chemical characteristics of activated carbons and the
adsorption of methylene blue from wastewater, Journal of Colloid and Interface
Science, 284 (2005) 440-446.
[19] C. Lai and C. Y. Chen, Removal of metal ions and humic acid from water by iron-
coated filter media, Chemosphere, 44 (2001) 1177-1184.
[20] C. Moreno-Castilla, Adsorption of organic molecules from aqueous solutions on
carbon materials, Carbon, 42 (2004) 83-94.
17
[21] B. H. Hameed, I. A. W. Tan and A. L. Ahmad, Adsorption isotherm, kinetic
modeling and mechanism of 2,4,6-trichlorophenol on coconut husk-based
activated carbon, Chemical Engineering Journal, 144 (2008) 235-244.
[22] B. H. Hameed, A. M. Din and A. Ahmad, Adsorption of methylene blue onto
bamboo-based activated carbon: kinetics and equilibrium studies, Journal of
hazardous materials, 141 (2007) 819-825.
[23] B. H. Hameed, A. Ahmad and K. Latiff, Adsorption of basic dye (methylene blue)
onto activated carbon prepared from rattan sawdust, Dyes and Pigments, 75
(2007) 43-49.
[24] S. Karagöz S, T. Tay, S. Ucar and M. Erdem, Activated carbons from waste
biomass by sulfuric acid activation and their use on methylene blue adsorption,
Bioresource technology, 99 (2008) 6214-6222.
[25] M. Pérez M, F. Torrades, J. A. Garcıa-Hortal, X. Domènech and J. Peral,
Removal of organic contaminants in paper pulp treatment effluents under Fenton
and photo-Fenton conditions, Applied Catalysis B: Environmental, 36 (2002) 63-
74.
[26] J. H. Ramirez, C. A. Costa, L. M. Madeira, G. Mata, M. A.Vicente and M. Rojas-
Cervantes, et al, Fenton-like oxidation of Orange II solutions using
heterogeneous catalysts based on saponite clay, Applied Catalysis B:
Environmental, 71 (2007) 44-56.
[27] D. Kim D, J. K. C. Chen and T. F. Yen, Naval derusting wastewater containing
high concentration of iron, treated in UV photo-Fenton-like oxidation, Journal of
Environmental Sciences-China, 22 (2010) 991-997.
[28] K. Dutta K, S. Mukhopadhyay, S. Bhattacharjee and B. Chaudhuri, Chemical
oxidation of methylene blue using a Fenton-like reaction, Journal of hazardous
materials 84 (2001) 57-71.
18
Table 2.1. Langmuir and Freundlich isotherm model coefficient for adsorption of MB on GACs. Conditions: Initial concentration of MB, 20-120 mg/L; volume of solution, 60 mL; initial solution pH, 6.0; temperature, 20°C; contact time, 72 h.
Freundlich Langmuir GAC qmax n R2 qmax KL R2
Virgin GAC
121.0
5.62
0.784
175.4
(±13.6)
3.36
0.998
Fe-GAC
133.9
4.19
0.892
238.1 (±0.78)
2.96
0.999
qmax= the maximum adsorption capacity (mg/g); n= an indicator of adsorption effectiveness; KL= the Langmuir constant (L/g)
19
Table 2.2. Surface charge of the Fe-GAC and MB at various pH
pH Fe-GAC (pH, pzc = 5.0) MB (pKa =3.8) Interactions
<3.8
Positive charge dominated
Non-charge
van der Waals or
hydrophobic
3.8<pH<5.0
Positive charge dominated
Positive charge
van der Waals or hydrophobic
pH>5
Negative charge dominated
Positive charge
Electrostatic
20
Table 2.3. pH, pzc of the virgin, acid-treated and acid-treated iron amended GAC
GAC pH, pzc
Virgin GAC
5.4
Acid-treated GAC 4.2
Fe-GAC
5.0
21
Table 2.4. Kinetic parameters for MB adsorption onto Fe-amended GAC
C0
(mg/L)
Experimental qe
(mg/g)
Pseudo first order Pseudo second order qe
(mg/g) k1
(min-1) R2 qe
(mg/g) K2
(g/mg/min) R2
90
10.3
12.8
0.028
0.986
10.6
0.004
0.998
450
54.3
73.8
0.015
0.974
57.7
0.0003
0.982
1050
125.7
167.6
0.013
0.979
136.3
0.00011
0.986
22
Figure 2.1. The molecular structure of methylene blue.
23
Figure 2.2. Adsorption kinetics of MB on the Fe-GAC. Conditions: Initial concentration of MB, 90, 450 and 1050 mg/L; volume of solution, 60 mL; initial solution pH, 6.0; temperature, 20°C; contact time, 300-500 min.
(a) Pseudo-first-order kinetics
(b) Pseudo-second-order kinetics
-5 -4 -3 -2 -1 0 1 2 3 4 5 6
0 100 200 300 400 500 ln (q
e-qt
)
Time (min)
90 mg/L 450 mg/L 1050 mg/L
0
5
10
15
20
25
30
35
0 100 200 300 400 500
t/qt (
min
.g/m
g)
Time (min)
90 mg/L
450 mg/L
1050 mg/L
24
Figure 2.3. Effects of iron loading on oxidation of the MB-spent Fe-GAC with the fixed H2O2 dose. Conditions: Initial concentration of MB, 1000 mg/L; volume of solution, 60 mL; initial solution pH, 6.0; temperature, 20°C.
0
10
20
30
40
50
60
70
80
0
0.01
0.02
0.03
0.04
0.05
0 2 4 6 8 10 12 14 16
Hal
f life
, H2O
2 (h)
Rat
e co
nsta
nt, H
2O2 (1
/h)
Fe load(mgFe/gGAC)
Rate constant (1/h)
Half life (h)
25
Figure 2.4. Effect of H2O2 on the oxidation of MB in the Fe-GAC. Conditions: Initial concentration of MB, 1000 mg/L; volume of solution, 60 mL; initial solution pH, 6.0; temperature, 20°C.
0
10
20
30
40
50
60
70
80
90
100
0 20 40 60 80 100 120 140
Oxi
datio
n ef
ficie
ncy
(%) o
f MB
in
the
Fe-G
AC
mMol H2O2/mMol MB
26
CHAPTER 3. Heterogeneous Photo-Fenton Oxidation of Methylene Blue Using CdS-Carbon Nanotube/TiO2 under visible light
3.1. Abstract
Heterogeneous photo-Fenton oxidation using CdS/multi-walled carbon nanotube-
TiO2 (CdS/MWCNT-TiO2) under visible light relied on combination of the photocatalytic,
photo-Fenton and photo-sensitizing oxidation. The photo-Fenton reactions resulted in
much faster and higher removal of methylene blue and total organic carbon than the
dark Fenton oxidation and the photocatalytic degradation alone at the selected
conditions. The optimum molar ratio of [methylene blue]:[H2O2]:[Fe3+] (1:12:3.4) in the
photo-Fenton oxidation indicated cost-effectiveness of the process. The scavenging
tests of the hydroxyl radicals and the valence holes suggested that the hydroxyl radical-
driven oxidation was the major step among the multiple reactions in the photo-Fenton
oxidation.
3.2. Introduction Synthetic dyes have been widely used for textile, pulp and paper, plastic, food,
cosmetic and pharmaceutical industries [1]. With mass production of synthetic dyes and
mass discharge of synthetic dye-containing wastewater, effective treatment of dye-
containing wastewater has been actively studied for several decades [2]. Conventional
biological and chemical oxidation methods are not efficient for the degradation of dyes
due to their complex aromatic structure and recalcitrant nature [3-5]. Adsorption using
activated carbon and low cost adsorbents is simple and efficient for the removal of dyes
while requiring expensive regeneration or generating significant amounts of solid
wastes. [6-8].
Advanced oxidation using the non-selective and strong oxidation capacity of
hydroxyl radicals (•OH) exhibited excellent oxidation of various dyes in wastewater [9].
Advanced oxidation processes include ozone (O3), H2O2, O3/H2O2, H2O2/UV, O3/UV,
Fenton or photo-Fenton oxidation, photocatalytic oxidation, electrochemical oxidation
and sonochemical oxidation [10-16]. Among various advanced types of oxidation,
Fenton oxidation was found to be the most efficient and cost-effective method for
27
treating various synthetic dyes [13, 16-18]. However, homogenous Fenton oxidation to
generate •OH using H2O2 and Fe2+ is effective at only narrow pH conditions (i.e., < pH
4) while it generates iron sludge after treatment [15, 19, 20]. The various types of
heterogeneous Fenton oxidation using iron oxides or iron immobilized onto solid
supports have been studied for overcoming the limitations of homogenous Fenton
oxidation [21-27]. Although the heterogeneous Fenton reaction can oxidize aqueous
contaminants at broader pH and reduce release of iron in water after treatment, it has
low reaction rate and treatment efficiency because of the low reactivity of the initial
reaction between Fe3+ and H2O2 to produce •OH (please see Equation 1-3; [26, 28]).
Most of the heterogeneous Fenton oxidation using Fe3+ (main valence state of Fe
catalyst) is approximately 5000 times slower than homogeneous Fenton oxidation with
Fe2+ as a catalyst (Eq. 1 and 3).
Fe2+ + H2O2→ Fe3+ + •OH + OH− k1 =51 M−1 s−1 (1)
•OH + M (contaminant) → M (intermediates) →→ final products (2)
Fe3+ + H2O2→ Fe2+ + HO2• + H+ k3 = 0.001–0.01 M−1 s−1 (3)
This limitation of heterogeneous Fenton oxidation can be overcome by taking
advantage of heterogeneous photo-Fenton oxidation that UV light irradiation (λ < 360
nm) quickly converts Fe3+ to Fe2+, which is reacted with H2O2 for immediately producing
•OH for efficient oxidation of organic contaminants (Eq. 4, [29-32]).
Fe(OH)2+ + hν → Fe2+ + •OH (4)
However, UV light-based photo Fenton oxidation also has practical limitations
that are difficult to utilize sunlight as an irradiation source since UV light only consists of
approximately 5% of sunlight [33, 34]. Consequently, significant efforts to develop
various photocatalysts have been made to utilize visible light as the major proportion of
the solar light source [35]. Recent studies reported visible light mediated photo-Fenton
28
oxidation by employing TiO2 doped with the metals (i.e., Fe, Co, Bi), non-metal (i.e.,
diethanolamine, triethylamine) and narrow band gap semiconductors (i.e., CdS, WO3)
[5, 33, 36-46]. Photo-Fenton oxidation using the modified TiO2, Fe3+ and H2O2 under
visible lights led to significant enhancement of •OH production and degradation of
synthetic dyes, phenol and pesticides due to synergistic effects of photocatalysis and
Fenton oxidation [5, 33, 47, 48]. The oxidation of synthetic dyes by visible light mediated
photo-Fenton reactions using TiO2 are listed in Table 3.1.
For the present study, the multi-walled carbon nanotube-supported CdS with
TiO2 (CdS/MWCNT-TiO2) was used as the photocatalyst for photo-Fenton oxidation of
methylene blue under visible light. Since CdS-TiO2 composites show spectral response
in visible area owing to the photosensitization of CdS with narrow energy band gap (2.4
eV), the photogenerated electrons from CdS in CdS-TiO2 composites are transferred
into the TiO2 particles while the holes remain in the CdS particle, which diminishes the
recombination of the electron-hole pair [49-52]. Thus, the CdS-TiO2 composites
exhibited high efficiency for the decomposition of pollutants in water and air under
visible light irradiation [49, 50, 53, 54].
However, CdS-TiO2 composite alone has been shown its limited practical
application due to severe aggregation, difficult recovery and low utilization rate [52, 55].
Various carbonaceous materials, such carbon nanotubes, activated carbon,
graphene and graphene oxide have been used as supports in order to overcome the
above disadvantages of CdS, TiO2 or CdS-TiO2 composites alone [52, 55, 56]. It has
been reported that carbonaceous materials are promising co-catalysts in photocatalysis
due to their large surface area and high electrical conductivity [57, 58]. Activated carbon
has its pore structure (high proportion of micropores, < 2nm) which often cause severe
pore blockage and reduce surface area for catalyst loading and adsorption of large
molecules [59]. Heterogeneous surface functionality of activated carbon also make it
difficult to have uniform dispersion of catalysts in its pores [59]. Compared with
activated carbon, carbon nanotube has high surface area, high proportion of interstitial
space (mainly mesopores) and easy functionalization, which resulted in well-dispersed
photocatalysts and high adsorption of target contaminants [60-62]. Application of
graphene as a support for photocatalysts is at early stage and needs to be further
29
investigated despite its high potential as a support for catalysts and an adsorbent [63,
64].
The results from this study and others’ reports [65, 66] indicated that the
photocatalytic degradation by the CdS/MWCNT-TiO2 led to high decolorization, slow
oxidation rate (i.e., 3-4 h) and low mineralization of synthetic dyes.
Therefore, the objective of the study is to achieve high decolorization and
mineralization of methylene blue in a short time by the photo-Fenton oxidation using the
CdS/MWCNT-TiO2 with Fenton’s reagents (Fe3+ and H2O2) under visible light. To the
best of our knowledge, the visible light mediated photo-Fenton oxidation of methylene
blue using the CdS/MWCNT-TiO2 photocatalyst have not been studied up to date. The
effects of various conditions on the photo Fenton oxidation of methylene blue were
assessed while the major mechanism associated with the photo-Fenton oxidation was
suggested.
3.3. Materials and methods 3.3.1. Chemicals and reagents
Multi-walled carbon nanotubes (MWCNTs) were purchased from Sigma-Aldrich
(St. Louis, MO, USA). The as-received MWCNTs have the inside and outside
diameters of 5-10 nm and 60-100 nm with the length of 0.5-500 µm [58]. The surface
area of the MWCNT was 40-300 m2/g ([58]). Methylene blue (Fisher Scientific, Waltham
MA), a cationic dye, which is difficult to degrade in the natural environment, was chosen
as the model pollutant. Cadmium acetate dihydrate (Cd(CH3COO)2·2H2O), benzene
(anhydrous, 99.8%), m-chloroperbenzoic acid (≥ 99.8%), sodium sulfide nonahydrate
(Na2S·9H2O), sodium formate (NaCOOH), ethanol, tetrabutyl titanate (C16H36O4Ti,
97.0%), ferric chloride hexahydrate (FeCl3∙6H2O) and hydrogen peroxide (30%) were
purchased from Sigma-Aldrich (St. Louis, MO, USA).
3.3.2. Preparation and characterization of CdS/MWCNT-TiO2
The CdS/MWCNT-TiO2 photocataylst was prepared through the sol-gel method
modified from Zhu et al.[55]. The oxidized MWCNTs were prepared by suspending 1 g
of MWCNT powder to a solution (2 g m-chloroperbenzoic acid in 80 mL benzene) under
30
vigorous stirring at 70℃ for 6h. The resulting mixture was continuously washed with
deionized water and followed by ethanol before drying in the oven at 90℃ to obtain
oxidized MWCNTs. To obtain the CdS/MWCNT, 0.693 g of cadmium acetate powder
was dissolved in 30 mL of water, and the solution was magnetically stirred for 10 min.
0.3 g of the oxidized MWCNTs were added to the solution and stirred for 30 min. 0.626
g of Na2S·9H2O dissolved in 30 mL of water was prepared separately, and added a
drop wise to the solution with constant stirring for 6 h at 70 ℃. The solution was vacuum
filtered, then washed with deionized water and ethanol before drying in the oven at 80
℃ for 12 h. The dried sample was milled and heated at 300 ℃ for 1h to get a
CdS/MWCNT. To yield CdS/MWCNT-TiO2, 0.4 g of the powdered CdS/MWCNT was
added to 5 mL of 97% tetrabutyl titanate containing 15 mL of benzene. After stirring the
mixture for 5 h at 70 ℃, the reacted CdS/MWCNT was then filtered and dried at 80 ℃
for 12 h. The resulting sample was heated at 350 ℃ for 1 h.
The scanning electron microscopy (SEM) images and X-ray energy dispersive
spectroscopy data (EDX) were obtained with Hitachi HT7700 field emission scanning
electron microscope (10 kV) with an Oxford INCA PentaFET-x3 Si (Li) EDX detector
system (Oxford instruments, Oxford, United Kingdom) for EDX analysis. Transmission
electron microscopy (TEM) images were obtained with Hitachi S-4800 (Hitachi High
Technologies America Inc., Schaumburg, IL, USA) at 100kV. Powder X-ray diffraction
(XRD) patterns of the powered sample were performed on a Rigaku MiniFlex II
diffractometer (Rigaku, The Woodlands, TX, USA) with a Cu Kα radiation source.
3.3.3. Visible light mediated photo-Fenton oxidation of methylene blue using the CdS/MWCNT-TiO2 photocatalyst
All of the experiments were conducted in a glass reactor (9 cm diameter, 1.5 cm
height) filled with 40 mL of 50 µM methylene blue solution containing 30 mg of the
photocatalyst (CdS/MWCNT-TiO2) at 20±2oC under visible light irradiation. A 8 W white
fluorescent lamp (Philips Co., USA) was used as simulated visible light sources to
irradiate horizontally into the glass reactor open to the atmosphere. The lamp was
placed at 10 cm from the glass reactor. The visible light intensity in the photoreactor
31
(63±2 µmol m-2 s-1) was measured with a dual radiation meter (Spectrum Technologies,
Inc., Aurora, IL, USA) with a measurable wavelength range of 400 – 700 nm. A
fluorescent lamp (wavelength, 𝜆 = 380 - 700 nm) can emit a small amount of UV light
[68-70]. However, compared to the visible light intensity, the negligible UV light intensity
(1.2 – 2.3 µmol m-2 s-1) in the photoreactor was detected by the UV light meter (250 -
400nm, Spectrum Technologies, Inc., Aurora, IL, USA)
Before turning on the lamp, the suspension was magnetically stirred for 30 min
to establish the adsorption-desorption equilibrium without visible light exposure. The
adsorption capacity of methylene blue onto the CdS-MWCNT/TiO2 was analyzed over
the pH range from 3 to 9.
The photo-Fenton oxidation of methylene blue was initiated by turning on the
lamp after a certain amount of FeCl3∙6H2O and H2O2 solution was added to the solution.
The effect of pH, [Fe3+]:[H2O2] and [H2O2]:[methylene blue] on the oxidation of
methylene blue were conducted for finding out the optimum conditions in the photo-
Fenton oxidation of methylene blue. For the photo-Fenton oxidation of methylene blue
at various molar ratio of [Fe3+]:[H2O2] and [methylene blue]:[H2O2], the solution pH was
adjusted to 3.5 before adding H2O2 since adding various iron concentration was
changed to pH 3.2-4.2. For evaluating the reusability and stability of the CdS/MWCNT-
TiO2 photocatalyst, the resulting suspension was centrifuged and dried at 60 ℃ at the
end of the photo-Fenton reaction. The dried photocatalyst was reused for the repeated
photo-Fenton oxidation tests in the next cycle with the same initial conditions.
Besides, the photo-Fenton oxidation (CdS/MWCNT-TiO2 photocatalyst, H2O2,
Fe3+, visible light), the photocatalysis alone (CdS/MWCNT-TiO2 photocatalyst, visible
light) and the dark Fenton oxidation (CdS/MWCNT-TiO2 photocatalyst, H2O2, Fe3+) were
performed at the selected conditions (50 µM methylene blue, 308.6 µM Fe3+, 617.2 µM
H2O2, visible light irradiation, initial pH of 3.5, temperature of 20±2oC). The scavenging
tests with sodium formate and ethanol were also carried out for finding out the major
mechanism associated with the photo-Fenton oxidation under visible light.
32
3.3.4. Analytical methods The filtered samples using 0.2 µm membrane filters were measured for H2O2
(n=3) using a modified peroxytitanic acid colorimetric procedure with a detection limit of
0.1 mg/L [71]. Iron concentration was measured using the phenanthroline Method [71].
The aqueous concentration of methylene blue was measured at regular interval
by monitoring the absorbance of the aqueous sample using a Biospec-mini UV-Vis
spectrophotometer (Shimadzu, Torrance, CA, USA) at 663 nm. The decolorization
efficiencies (removal of methylene blue) were calculated by the following expression:
Decolorization efficiency (%) = 1− !" !!" !
× 100 (5)
where [MB]0 and [MB]t are the aqueous concentration of methylene at reaction time 0
and t, respectively. The residual amount of methylene in the CdS/MWCNT-TiO2 was
found to be negligible after measuring its methanol-extraction samples. Total organic
carbon (TOC) of the initial and irradiated samples was determined with a Shimadzu
TOC-L combustion analyzer (Shimadzu, Torrance, CA, USA). Mineralization efficiencies
were calculated by following expression:
Mineralization efficiency (%) = 1− !"# !!"# !
× 100 (6)
where [TOC]0 and [TOC]t are the TOC values at reaction time 0 and t, respectively.
The pH at point of zero charge (pH, pzc) of the CdS/MWCNT-TiO2 was
determined using the pH drift method [71]. 40 mL of DI water amended with 0.01 M
NaCl was placed in 40 mL amber vials and sparged with N2 (200 mL/min; 10-15 min) to
eliminate CO2, and stabilize pH. The pH was adjusted from pH 2 to pH 11 in a series of
vials by adding either HCl or NaOH while purging the headspace with N2. 0.04 g of
CdS/MWCNT-TiO2 was added and the vial was capped immediately. The final pH (pH,
final) was measured in each of the vials after 48h and plotted versus the initial pH (pH,
33
initial). The pH, pzc was determined graphically at the intersection of pH, final and the line
pH, final = pH, initial.
3.4. Results and discussion 3.4.1. Characterization of the CdS/MWCNT-TiO2
The TEM and SEM images of the CdS/MWCNT-TiO2 in Figure 3.1A and Figure 3.1B present the structure of the CdS/MWCNT-TiO2. The images exhibited the
MWCNTs attached with the CdS (6-10 nm, average: 8.35 nm, regular black dots) and
the TiO2 (6.0-8.5 nm, average: 7.18 nm, bright dots) while indicating well-dispersed
mixture of CdS, TiO2 and MWCNT with little agglomeration. The XRD patterns of the
TiO2, CdS/MWCNT and CdS/MWCNT-TiO2 are presented in Figure 3.2. The diffraction
peaks at 2θ=25.3, 37.8, 48.1 and 54.1 in the CdS/MWCNT-TiO2 were attributed to
anatase-TiO2 [52] while the diffraction peaks at 26.3, 28.2, 43.7 and 47.5 in the
CdS/MWCNT-TiO2 were attributed to the hexagonal CdS phase [72]. The XRD results
confirmed that the anatase TiO2 and hexagonal CdS coexisted in the CdS/MWCNT-
TiO2. The EDX analysis of the CdS/MWCNT-TiO2 revealed strong peak for TiO2 and
small peaks for the CdS and the CNT, which are consistent to the relative composition
of these compounds in the photocatalyst (Table 3.2).
3.4.2. Photo-Fenton oxidation of methylene blue using CdS/MWCNT-TiO2 under visible light
The photo-Fenton oxidation of methylene blue under visible light led to nearly
complete removal of methylene blue in 30 min (97.7%) compared with the dark Fenton
oxidation (75%) and the photocatalytic degradation alone (63%, Figure 3.3). The
pseudo first-order rate constant of the photo-Fenton oxidation also exhibited about 5.6 –
6.9 times as fast as the dark Fenton oxidation and photocatalysis alone (Figure 3.3).
The TOC removal by the photo-Fenton oxidation (83%) was also higher than those by
the dark Fenton oxidation and the photocatalysis alone indicating its effective
mineralization. Such the results were consistent to those as reported by Kim et al. and
others ([5, 33, 48]). The dark Fenton and the photocatalytic oxidation alone showed
34
slower and lower removal of methylene blue and TOC mainly due to the slow reaction
with Fe3+ or the limited oxidation capacity.
Figure 3.4 presents the adsorption of methylene blue onto the CdS/MWCNT-
TiO2 under dark condition at pH 3-7. The adsorption experiments indicated quite similar
adsorption capacity for methylene blue (32-35%) at various pH although it included
multiple mechanisms such as electrostatic interaction and uncharged adsorption based
on the pKa of methylene blue (~3.8) and the pH,pzc of the photocatalyst (~3.3). At pH 3-
3.3, the non-electrostatic interactions such as van der Waals and hydrophobic
interaction would be the major interactions between the slightly negatively charged
photocatalyst (solution pH < pH, pzc) and the non-charged methylene blue (solution pH <
pKa, MB). At pH 3.3-3.8, van der Waals and hydrophobic interaction would be also the
major interactions between the negatively charged photocatalyst (solution pH > pH, pzc)
and the non-charged methylene blue (solution pH < pKa, MB). On the other hand, at the
pH 3.8-7.0, the electrostatic interactions interaction would be the major interaction
between the negatively charge photocatalyst (solution pH > pH, pzc) and the positively
charged methylene blue (solution pH > pKa, MB).
Therefore, the adsorption of methylene blue onto the photocatalyst at pH 4-7 was
mainly mediated by the electrostatic interaction while the adsorption at pH 3 was driven
by van der Waals/hydrophobic interaction. The results indicated there was no significant
difference of adsorption of methylene blue on the photocatalyst at pH 3-7.
On the other hand, the removal of methylene blue by the photo-Fenton oxidation
at pH 3-7 resulted in highest removal efficiency of methylene blue at pH 3 followed by
pH 4, 5 and 7 (Figure 3.4). It is consistent to the optimum pH for the photo-Fenton
oxidation reported by others (Figure 3.4; [5, 16, 33]).
Determination of the optimum concentration of the Fenton’s reagent (H2O2, Fe2+
or Fe3+) is highly important for enhancing treatment efficiency and minimizing operating
costs [73, 74]. The increase of the [Fe3+]:[H2O2] from 0.05 to 0.5 enhanced the
methylene blue removal efficiency from 75% to 100% and the TOC removal efficiency
from 61% to 84% (Figure 3.5). The results agreed with the optimum [Fe3+]:[H2O2] ratio
of 0.29. Besides, both methylene blue and TOC removal efficiency increased from 32%
to 99% and from 2 % to 83 % when the molar ratio of [H2O2]:[methylene blue]
35
increased from 3 to 12 (Figure 3.6). Therefore, the combined results from Figure 3.5
and Figure 3.6 led to the optimum molar ratio of [methylene blue]:[H2O2]:[Fe3+] which
was 1:12:3.42 at pH 3.5. Surprisingly the optimum molar ratio of [H2O2]:[methylene blue]
for the photo-Fenton reaction in this study (12 mol H2O2/mol methylene blue) was found
to be much lower than those for the heterogeneous dark-Fenton reaction (28 mol
H2O2/mol methylene blue, pH 3 [75]), the UV light-activated photo-Fenton reaction (32
mol H2O2/mol methylene blue, pH 5 [76]) and the visible light-mediated photo-Fenton
reaction (183 mol H2O2/mol methylene blue, pH 6.5 [40]) reported by others. Table 3.3
summarizes the molar ratio of [H2O2]:[methylene blue] used for the photo-Fenton and
Fenton oxidation of methylene blue. The results proved that the photo-Fenton process
with CdS/MWCNT-TiO2 would be a highly efficient and cost-effective solution to
degrade methylene blue at even low H2O2 concentration.
Reusability of catalysts is of prime importance in assessing the practical
application of photocatalysts in water and wastewater treatment. It can contribute
significantly to lowering the operational cost of the process [5]. The six cycles of the
photo-Fenton oxidation using the same catalyst resulted in 98 - 99% removal of
methylene blue and 96 - 97% of COD, which confirms high stability of the photocatalyst
over the multiple photo-Fenton reactions (Figure 3.7). The EDX data in Table 3.2
indicated the iron content (0.5 %) in the CdS/MWCNT-TiO2 after the photo-Fenton
oxidation. The result can support that most of Fe3+ was immobilized onto the surface of
the CdS/MWCNT-TiO2 during the photo-Fenton oxidation. The measurement of iron
concentration during the photo-Fenton reaction confirmed the immediate adsorption of
iron onto the surface of the CdS/MWCNT-TiO2 in 10 min (the data not shown).
A possible reaction mechanism of photo-Fenton oxidation of methylene blue
using CdS/MWCNT-TiO2 under visible light irradiation is summarized in Figure 3.8
based on the mechanisms reported by others [48, 55, 77-82]. First, the photocatalytic
degradation of organic pollutants under visible light irradiation (> 400 nm) using the
CdS/TiO2 composites was already explored by other studies [55, 81]. The photocatalytic
oxidation is based on absorption of visible light for exciting the electrons on the valence
band of CdS and transferring the electrons to the conduction bands of TiO2. The
photogenerated electron (ecb-) at the surface of TiO2 is scavenged by electron acceptors
36
such as O2 while the photogenerated hole (hvb+) can react with OH- or H2O at the
surface of CdS. Both reactions can generate !O2- or !OH, respectively, which can
destroy organic pollutants. Second, the photogenerated electron is transferred to
adsorbed Fe3+ at the TiO2 surface, leading to a conversion of Fe3+ to Fe2+. The
regenerated Fe2+ rapidly reacts with H2O2 to produce !OH while the regeneration of Fe3+
from Fe2+ by Fenton reaction enables an easy cycle of Fe3 +/Fe2+ thus maintaining a
considerable concentration of Fe3+ as an electron acceptor. In addition, scavenging
photogenerated electron by Fe3+ at the TiO2 surface possibly prevents direct electron-
hole recombination, which enhances the photocatalytic activity [77]. The synergistic
combination of TiO2 photocatalysis and Fenton-like reaction (Fe3+/H2O2) under UV
irradiation showed significant kinetic enhancement on oxidation of organic pollutant [48].
Third, several studies have been reported on the photo-Fenton oxidation of organic dye
pollutants under visible light irradiation in which visible light was also found enhance the
photocatalytic reaction [77-80, 82]. Similarly methylene blue as a photo-sensitizer would
be transformed to the excited form (MB*) with visible light irradiation in the present study.
Some researchers explained that the intermolecular electron transfer between the MB*
and Fe3+ regenerates the Fe2+ which accelerates the photo-Fenton reaction[79, 80, 83].
Therefore, the CdS/MWCNT-TiO2 in this study is suggested to combine possibly the
above three mechanisms.
Figure 3.9 presents the •OH scavenging with formate and ethanol. Formate can
scavenge •OH and electron holes while ethanol scavenges only •OH.
The scavenging efficiency was estimated by using the following equation[48]:
Scavenging efficiency (%) = 1− !"!!
× 100 (7)
where ko and ki are the pseudo first-order rate constants for methylene blue degradation
in the absence and the presence of the scavengers.
The photo-Fenton oxidation of methylene blue in the presence of formate or
ethanol showed similar scavenging effects (86-87%) compared to the photo-Fenton
oxidation of methylene blue without any scavengers. It clearly demonstrated that the
photo-Fenton oxidation of methylene blue using CdS/MWCNT-TiO2 under visible light
mainly relies on •OH-driven oxidation among the multiple reactions in the process. It
37
also emphasizes importance of the cycling of Fe3+/Fe2+ during the photo-Fenton
oxidation which influences on the •OH production for the rapid oxidation of
contaminants.
3.5. Acknowledgement
This work was supported by University of Hawaii at Manoa (Project number:
2300016).
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46
Table 3.1. Summary of heterogeneous photo-Fenton and photo-Fenton like oxidation of azo dyes under visible light irradiation
Target Dye
Catalyst/ Support
Operating conditions % Decolorization /% TOC removal
Reference
Methylene blue (MB)
Fe (Phen)3-Y
300 W UV-lamp (λ>420 nm); H2O2/MB = 183 (molar ratio); pH 6.5; 2 g catalyst; volume, 120 mL; reaction time, 120 min
RE%, MB = 77
[40]
Acid Blue 29 (AB29)
Fe-SiO2 150 W metal halide lamp (400<λ<700); H2O2/AB29 = 123 (molar ratio); 0.4 g/L catalyst; volume, 250 mL; 26± 2 oC; pH 3.0; reaction time, 100 min
RE%, AB29 = 98 RE%, TOC = 60
[41]
Reactive Brilliant orange X-GN (X-GN)
Fe-Mt
300 W halogen lamp (λ>420 nm); light intensity, 42300±200 lx; H2O2/X-GN = 28 (molar ratio); 0.6 g/L catalyst; volume, 200 mL; 30 oC ; pH 3.0; reaction time, 140 min
RE%, X-GN = 99 RE%, TOC = 53
[44]
Rhodamine B (RhB)
Fe-R
500 W halogen lamp (λ>420 nm); light intensity, 4.67 x104 lx; H2O2/X-GN = 75 (molar ratio); 0.4 g/L catalyst; volume, 25 mL; catalyst, 0.4 g/L; 25 oC ; pH 4.5; reaction time, 6 h
RE%, RhB = 75 RE%, TOC = 94
[39]
Rhodamine B (RhB)
Fe-R
300 W Dy lamp ( λ>420 nm); H2O2/RhB = 23.5 (molar ratio); 2 g/L catalyst; volume, 100 mL; 25 oC ; pH 4.25; reaction time, 100 min
RE%, RhB = 99 RE%, COD = 89
[46]
Rhodamine B (RhB)
H-Fe-S
300 W Dy lamp ( λ>420 nm); H2O2/RhB = 60 (molar ratio); 2 g/L catalyst; volume, 100 mL; 25 oC ; pH, 4.25; reaction time, 100 min
RE%, RhB = 100 RE%, COD = 86 RE%, TOC = 38
[38]
47
Table 3.2. EDX results of CdS/MWCNT-TiO2 before and after photo-Fenton reaction under visible light irradiation. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL; CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; [Fe3+], 180 µM; [H2O2], 600 µM; temperature,20°C; light intensity, 63±2 µmol m-2 s-1.
Element Before photo-Fenton oxidation Weight % Atomic %
After photo-Fenton oxidation Weight % Atomic %
C 13.2 21.6 13.1 21.6 O 52.6 64.6 51.4 63.8 Ti 32.1 13.2 33.8 14 Cd 1.1 0.19 0.8 0.14 S 0.42 0.25 0.47 0.29 Fe 0 0 0.46 0.16 Al 0.14 0.1 0 0
Total 100 100 100 100
48
Table 3.3. Summary of optimal [H2O2]: [methylene blue] molar ratio in Fenton and photo-Fenton treatment of methylene blue. Process Optimal
[H2O2]:[MB] molar ratio
% Decolorization /% Mineralization
Operating conditions Reference
Homogeneous Fenton oxidation
14 RE%, MB =98 RE%, COD =81
26 oC; pH 2.2 - 2.6; 1h
[73]
Heterogeneous Fenton oxidation
127 RE%, MB = 100 50 oC; pH 3; 1h; LiFe(WO4)2
[74]
Heterogeneous Fenton oxidation
28 RE%, MB = 84 20 oC; pH 3; Fe-GAC
[64]
Homogeneous photo-Fenton (UV) oxidation
51 RE%, MB = 100 RE%, TOC = 100
pH 3; 20 min
[75]
Heterogeneous photo-Fenton (UV) oxidation
32 RE%, MB = 94 25 oC; pH 5; 1h; LiFe(WO4)2
[65]
Heterogeneous photo-Fenton (visible) oxidation
183 RE%, MB = 77 pH 6.5; 2h; Fe(Phen)3
+Zeolite [40]
Heterogeneous photo-Fenton (visible) oxidation
12 RE%, MB = 99 RE%, TOC = 83
20 oC; pH 3.5; 1h; CdS/MWCNT-TiO2
Our work
49
Figure 3.1. TEM (a) and SEM (b) images of the CdS/MWCNT-TiO2
(a)
(b)
50
Figure 3.2. XRD patterns of (a) TiO2, (b) CdS/TiO2 and (c) CdS/MWCNT-TiO2 (a)
(b)
(c)
51
Figure 3.3. Removal of methylene blue and TOC using photo-Fenton, dark Fenton oxidation and photocatalysis. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL; CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; temperature, 20°C; light intensity, 63±2 µmol m-2 s-1 ; [H2O2], 600 µM; [Fe3+], 180 µM (only for dark and photo-Fenton oxidation). (a)
(b)
0 10 20 30 40 50 60 70 80 90
100
-60 -30 0 30 60 90 120
Met
hyle
ne B
lue
Rem
oval
(%)
Time (min)
Dark Fenton Photocatalysis Photo-Fenton
k= 0.117 min-1
k= 0.017 min-1
k= 0.021 min-1
0
10
20
30
40
50
60
70
80
90
100
1 2 3
Rem
oval
% o
f met
hyle
ne b
lue
and
TOC
Photo-Fenton Dark Fenton Photocatalysis
MB RE%, 60 min MB RE%, 120 min TOC RE%, 120 min
52
Figure 3.4. Effects of pH on adsorption and photo-Fenton oxidation of aqueous methylene blue. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution, 40 mL; catalyst (CdS/MWCNT-TiO2) loading, 30 mg; reaction time, 1 h (adsorption), 1 h (photo-Fenton oxidation); temperature, 20°C.
0
10
20
30
40
50
60
70
80
90
100
3 4 5 7
Met
hyle
ne b
lue
rem
oval
(%)
pH
Adsorption Photo-Fenton
53
Figure 3.5. Effects of [Fe3+]/[H2O2] on photo-Fenton oxidation of methylene blue. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; [H2O2], 600µM; temperature, 20°C; light intensity, 63±2 µmol m-2 s-1.
0
10
20
30
40
50
60
70
80
90
100
0 0.1 0.2 0.3 0.4 0.5
Rem
oval
% o
f met
hyle
ne b
lue
and
TOC
[Fe3+]/[H2O2]
MB removal (%) TOC removal (%)
TOC removal
MB removal
54
Figure 3.6. Effects of [H2O2]/[methylene blue] on photo-Fenton oxidation of methylene blue. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL; CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; [Fe3+]/[H2O2], 0.35; temperature, 20°C; light intensity, 63±2 µmol m-2 s-1.
0
10
20
30
40
50
60
70
80
90
100
2 4 6 8 10 12 14
Rem
oval
% o
f met
hyle
ne b
lue
and
TOC
[H2O2]/[methylene blue]
MB removal (%) TOC removal (%)
TOC removal
MB removal
55
Figure 3.7. Reusability of the CdS/MWCNT-TiO2 in the photo-Fenton oxidation of methylene blue. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL; CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; [Fe3+], 180 µM; [H2O2], 600 µM; temperature, 20°C; light intensity, 63±2 µmol m-2 s-1.
0
10
20
30
40
50
60
70
80
90
100
0 1 2 3 4 5 6 7
Met
hyle
ne b
lue
rem
oval
(%)
The number of cycles
56
Figure 3.8. Proposed mechanisms of the photodegradation of methylene blue on CdS/MWCNT-TiO2 composites under visible light irradiation.
57
Figure 3.9. Effects of formate and ethanol on scavenging on photo-Fenton oxidation of methylene blue. Conditions: Initial concentration of methylene blue, 50 µM; volume of methylene blue solution; 40 mL; CdS/MWCNT-TiO2 catalyst, 30 mg; initial pH, 3.5; [Fe3+], 180 µM; [H2O2], 600 µM; temperature, 20°C; light intensity, 63±2 µmol m-2 s-1 .
0
10
20
30
40
50
60
70
80
90
100
1 2 3 4
Scav
engi
ng e
ffici
ency
(%)
[H2O2]:[Formate]=1:1 [H2O2]:[Formate]=1:10 [H2O2]:[Ethanol]=1:1 [H2O2]:[Ethanol]=1:10
58
CHAPTER 4. Effect of Temperatures on Adsorption and Oxidative Degradation of Bisphenol A in an Acid-Treated Iron-Amended Granular Activated Carbon
4.1. Abstract
The present study suggests a combined adsorption and Fenton oxidation using
an acid-treated Fe-amended granular activated carbon (Fe-GAC) for effective removal
of bisphenol A in water. When the Fe-GAC adsorbs, and is saturated with BPA in water,
Fenton oxidation of BPA occurs in the BPA-spent Fe-GAC for regeneration of the GAC.
Particularly this study showed temperature as an effective mean to enhance adsorption,
desorption, diffusion and Fenton oxidation associated with the adsorption and Fenton
oxidation of BPA on the Fe-GAC. The adsorption rates of BPA onto the Fe-GAC were
enhanced with increasing temperature mainly due to increase in diffusion of BPA. The
estimated thermodynamic parameters associated with adsorption of BPA indicated that
the adsorption of BPA onto the Fe-GAC was a spontaneous, exothermic and physical
adsorption process. On the other hand, the molar ratio of [H2O2]:[BPA] (36 to 108
H2O2/mol BPA) in the Fenton oxidation of the BPA-spent Fe-GAC led to 91 – 99%
removal of BPA with negligible aromatic products and some soluble organic acids. The
oxidation rates of BPA and H2O2 during the Fenton oxidation of the BPA-spent Fe-GAC
were drastically enhanced by the factor of 2.5 and 5 when the reaction temperature
increased from 293 – 331 K, respectively. The comparative analysis of temperature-
dependent enhancement in diffusion, desorption and oxidation rates of BPA in Fe-GAC
indicated that the BPA oxidation in the Fe-GAC is mainly controlled by diffusive
transport of BPA from the Fe-GAC. Besides, the Thiele-modulus analysis clearly
supported more significant pore diffusion limitation of H2O2 in the Fe-GAC with
increasing temperature and the Fe-GAC particle size.
4.2. Introduction Bisphenol A (BPA) is an endocrine disrupting compound which is used in the
production of polycarbonate plastics, epoxy resins, flame retardants and packaged food
[76-78]. Due to the release of a great quantity of BPA into the aquatic environment,
BPA has been detected at ppb levels in industrial and urban wastewater and sub-ppb
59
level in treated effluents and drinking water [79-82]. Since BPA at low concentrations
can cause hormonal disruption, infertility and breast cancer, it is necessary to develop
effective technology for the removal of BPA from wastewater [81, 83-88].
Adsorption and advanced oxidation are among various remediation options
considered to be effective treatment methods to remove BPA from wastewater and
water sources [89-95]. Adsorption is simple and effective, but it requires high operating
costs associated with the regeneration of contaminant-spent adsorbents (i.e., activated
carbon) [59, 96]. The current thermal regeneration method often leads to significant
deterioration of the carbon pore structure, specific surface area and functionality which
influence re-adsorptive capacity of activated carbon [97]. Advanced oxidation processes
(i.e., ozone, O3/H2O2, H2O2/UV, Fenton, ultrasound, photocatalytic and electrochemical
oxidation) can achieve effective degradation of BPA [10, 12-16]. However, advanced
oxidation processes with low concentrations of BPA would result in low reaction rates
and efficiencies while the short lifetime of the oxidants requires the use of large
amounts of oxidants to oxidize trace levels of BPA leading to increased operation costs
[82, 98].
This study suggests a combined adsorption and Fenton oxidation using an acid-
treated Fe-amended granular activated carbon (Fe-GAC) for effective removal of
bisphenol A in water. When the Fe-GAC adsorbs and is saturated with BPA in water,
Fenton oxidation of BPA occurs in the BPA-spent Fe-GAC for regeneration of the GAC.
The adsorption step of the treatment process allows BPA to undergo phase transfer
from a dilute solution onto the GAC surface where the BPA is immobilized and
concentrated. When Fenton oxidation is carried out in the iron-amended BPA-spent
GAC, the oxidation of BPA is much more efficient and effective relative to chemical
oxidative processes carried out in the dilute solution [59]
However, Fenton oxidation of BPA in BPA-spent GAC (Fenton oxidation-driven
regeneration of GAC) involves multiple steps: (1) desorption of BPA from solid (GAC) to
liquid phase (water), (2) diffusive transport of BPA within the pores involving pore and
surface diffusion, (3) diffusive transport of BPA through a quiescent film surrounding the
particle, and (4) advective transport into the bulk solution (Figure 4.1) [59]. Since the
previous studies reported that intraparticle diffusion of contaminant (i.e., diffusion of
60
desorbed BPA toward external surface of GAC) was the major limiting step in the
Fenton oxidation-driven regeneration of spent GAC, it needs to be enhanced for
effective oxidation of contaminants and regeneration efficiency of GAC [99]. To
overcome the intraparticle diffusion limitation in the Fenton oxidation-driven
regeneration of spent GAC, it was found that increasing temperature and reducing
particle size of GAC significantly enhanced the Fenton oxidation efficiency of methyl
tert-butyl ether in the Fe-GAC [59, 96].
Therefore, the objective of this study was to effectively remove BPA in water by a
combined adsorption and Fenton oxidation process using the Fe-GAC. To the best of
our knowledge, there have been no studies for the adsorption and heterogeneous
Fenton oxidation of BPA using the Fe-GAC. Furthermore this study focused on the
effects of temperature on adsorption, desorption, diffusion and Fenton oxidation of BPA
in the Fe-GAC to enhance overall treatment efficiency of BPA.
The possible mechanisms and oxidative products associated with adsorption and
oxidation of BPA in the Fe-GAC at elevated temperatures are addressed through the
execution of the experimental procedures, and the critical analysis and discussion of the
results.
4.3. Materials and methods 4.3.1. Chemicals and reagents
The GAC (URV, 8 × 30 mesh, Calgon Carbon Corp., Pittsburgh, PA) derived
from bituminous coal was used for this study. The GAC was rinsed with deionized (DI)
water, dried in an oven at 378 K, and stored in a desiccator until used. The surface area
and pore volume of the GAC as received were 1290 m2/g and 0.64 mL/g, respectively
[6]. Bisphenol A (BPA), ferrous chloride (FeCl2·4H2O) and hydrogen peroxide (30%,
(w/w)) were purchased from Sigma-Aldrich (St. Louis, MO, USA).
4.3.2. Preparation of acid-treated Fe-amended activated carbon The acid-treated GAC was prepared by the methods reported by Kan and Huling
[59]. The raw GAC (10 g) was suspended in nitric acid at pH 3 for 96h to increase the
acidic surface oxides and to lower the pH at point of zero charge (pHPZC) (i.e., the pH at
61
which positive and negative surface charges are approximately equal). The initial pHPZC
of the GAC was 5.0-5.1 and was lowered to 4.2-4.5 through the acid treatment process
[21]. Subsequently, the acid-treated, Fe-amended GAC (Fe-GAC) was prepared by
amending a ferrous iron solution (1.5 L; 50 mg/L) to the acid-treated GAC (10 g). Under
this condition, the pH of the ferrous solution amended to the GAC was nearer the pHPZC
of the acid-treated GAC. Consequently, repulsive forces between the GAC surface and
iron ions were limited allowing greater penetration depth of the Fe+2 in the GAC, and
more uniform Fe deposition on GAC surfaces [21, 22]. After the iron amendment onto
the acid-treated GAC, the GAC was rinsed (3×) with DI water to eliminate the chloride
residual in the GAC, and dried in an oven (353 K, 24 h).
4.3.3. Batch adsorption of BPA onto the Fe-GAC 4.3.3.1. Adsorption isotherm
In this study, the Fe-GAC was used for adsorption and Fenton oxidation of BPA
while the raw GAC’s function was limited to only adsorption of BPA. Thus, the
adsorption capacities of the raw GAC and the Fe-GAC were comparatively investigated.
The batch experiments for adsorption isotherm were conducted by suspending 25 -100
mg of the raw GAC or the Fe-GAC in a series of flasks containing 50 - 150 mL of BPA
solutions at the concentration of 0.10 – 0.25 g/L. The initial pH of BPA solution (pH =
6.0) was adjusted by adding 0.1 M HCl or 0.1 M NaOH. The adsorption process was
carried out by shaking at the constant speed (200 rpm) at room temperature (293 K) for
4 – 7 d to ensure equilibrium was reached. The post-adsorption BPA solution was
sampled after equilibrium (> 4 d) in replicate and analyzed. The differences between
initial and final concentrations were used to calculate the mass of BPA adsorbed to the
GAC.
For understanding the adsorption mechanism and capacity of the raw GAC and
the Fe-GAC, the Langmuir and Freundlich isotherm models were used to interpret the
batch isotherm data. The nonlinear form of Freundlich and Langmuir isotherms are
shown in Eq. 1 and 2, respectively:
Freundlich isotherm: (1) )(C kq 1/nef=e
62
Langmuir isotherm: (2)
where kf is a constant to indicate adsorption capacity, 1/n is an indicator of adsorption
effectiveness, qm is the maximum adsorption capacity (mg/g) corresponding to complete
monolayer coverage of the surface, qe is the amount of BPA adsorbed per unit mass of
adsorbent at equilibrium (mg/g), Ce is the liquid-phase concentration of BPA at
equilibrium (mg/L) and KL is the Langmuir constant (L/g) related to the
sorption/desorption energy. The nonlinear forms of Langmuir and Freundlich models
(Eq. 1 and 2) were used to estimate the model parameters. A trial and error procedure
was performed using the solver function in Microsoft Excel to minimize the sum of
squared errors which was the sum of squared errors ((q, measured – q, calculated)2).
4.3.3.2. Adsorption kinetics and thermodynamic analysis at various temperatures The adsorption kinetics of BPA on the Fe-GAC was studied by adding 0.1 g of
the Fe-GAC into 100 mL of 100 mg/L BPA solution at various solution temperatures
(293 – 333 K) and pH 6. The solution was then magnetically stirred at 200 ppm for 4 – 7
h while the samples were collected at a regular interval. All of the samples were
centrifuged prior to HPLC analysis.
The kinetic of the adsorption were analyzed using two different kinetic models:
the pseudo-first order and pseudo-second order. The pseudo-first order kinetic model is
expressed by Eq. 3 [100]:
q t = q!(1− e!!!!!) (3)
where q(t) and qe represent the amount of BPA adsorbed (mg/g) at any time t and at
equilibrium time, respectively, and k1 represents the adsorption rate constant (min-1).
The pseudo-second-order equation based on equilibrium adsorption is expressed by
Eq. 4 [101]:
q t = !!!!!!!!!!!!!
(4)
qe = qm KLCe
1+KLCe
63
where k2 (g/mg min) is the rate constant of second-order adsorption, qe and q(t)
represent the amount of BPA adsorbed (mg/g) at equilibrium and at any time t.
Similar to the adsorption isotherms, the model parameters in the pseudo 1st and 2nd
order kinetics were estimated using the nonlinear forms of the adsorption kinetic models
(Eqs. 3 and 4).
In addition, the thermodynamic parameters such as Gibb’s free energy (ΔG0), change in
enthalpy (ΔH0) and entropy (ΔS0) for the adsorption of BPA on the Fe-GAC have been
calculated by using the Eq. 5 and 6 [102]:
∆G° = ∆H° − T∆S (5)
log !!!!
= ∆!°!.!"!#
+ !∆!!!.!"! !"
(6)
where qe is the amount of BPA adsorbed per unit mass of the Fe-GAC (mg/g), Ce is
BPA concentration at equilibrium (mg/L), T is temperature in K, and R is the gas
constant (8.314 J/mol-K).
4.3.4. Fenton oxidation-driven regeneration of the BPA-spent Fe-GAC 4.3.4.1. Effect of H2O2 dose and temperature on oxidation of BPA in the Fe-GAC Batch experiments for heterogeneous Fenton oxidations of BPA at various molar
ratio of [H2O2]:[BPA] and temperature were carried out in a glass flask with a side-arm
fitted with a Teflon tube containing GAC that captured any volatile emission during the
Fenton oxidation. 0.1 g of the Fe-GAC was saturated with 100 mL of 150 – 180 mg/L of
BPA solution (pH 6) over three times in order to adsorb a sufficient amount of BPA onto
the Fe-GAC. Fresh BPA stock was added after each subsequent adsorption. The
summation of the difference between initial and final concentration for each adsorption
were used to calculate the mass of BPA adsorbed to the Fe-GAC for total adsorption.
The BPA saturated Fe-GAC was separated from BPA solution, and dried in an oven at
333 K. Each glass flask was filled with 10 mL nanopure water with 0.1 g of the BPA-
saturated Fe-GAC (320 ± 20 mg BPA/g Fe-GAC). The pH of the GAC slurry was
64
adjusted to 3 with 0.1 M nitric acid for effective Fenton oxidation. The pH for Fenton
oxidation was different from that for adsorption (pH 6) because Fenton oxidation for the
treatment of wastewater is very efficient at acidic pH (2.8-3). The Fenton oxidation at pH
4-7 showed much lower and poor efficiency to oxidize BPA.
The reaction was then initiated by adding H2O2 to the glass reactor under
magnetic stirring. The concentrations of BPA and H2O2 were monitored at regular
intervals.
First, the effect of H2O2 concentration on the Fenton oxidation of BPA in the Fe-
GAC was investigated at various molar ratio of [H2O2]:[BPA] (9 to 108 mol H2O2/mol
BPA) at pH 3 and 293 K. Then, the effects of temperature on Fenton oxidation of BPA
in the Fe-GAC were also examined at various reaction temperatures (293 – 333 K) at
pH 3, 0.1 g of the Fe-GAC and [H2O2]:[BPA] of 18. The overall removal efficiencies of
BPA in the Fe-GAC at various temperatures were evaluated after the H2O2 was
completely consumed. In parallel with the overall removal efficiency of BPA, the BPA
oxidation rate (mg BPA/g GAC-h) was also calculated by measuring the residual BPA in
the GAC per hour. Although the granular activated carbon (Calgon Carbon, PA) in the
sidearm traps of the flaks was analyzed for BPA, volatile BPA losses were negligible
over the reaction temperature. Negligible leaching of Fe from the Fe-GAC (0.29% Fe
out of total Fe in the GAC) during the Fenton oxidation at pH 3 and [H2O2]:[BPA] of 18,
which was the optimum condition in this study, was detected. Compared with the
[H2O2]:[BPA] of 18, the Fenton oxidation at high molar ratio of [H2O2]:[BPA] (36-108)
showed 4.6 – 9.2 % Fe leaching out of total Fe in the GAC. In addition, negligible
evaporation of water, changes in BPA concentration (< 1% for both) and thermal
degradation of BPA during the adsorption experiments at 293 – 333K were observed.
4.3.4.2. Effect of temperature on desorption of BPA The effect of temperature on desorption of BPA from the Fe-GAC was performed
at different reaction temperature of 293, 313 and 333 K. The BPA saturated Fe-GAC
(300-350 mg/g) was separated from the solution and washed with nanopure water to
removal any unadsorbed BPA then added into a flask containing 100mL of nanopure
water at pH 3. The solution containing the Fe-GAC was magnetically stirred during
65
desorption at 293 – 333 K, and the sample was taken at a regular interval, centrifuged
and then analyzed using HPLC.
4.3.5. Analytical methods H2O2 and iron concentration in the aqueous solution: Filtered samples (using
0.2 µm membrane filters) were measured for H2O2 (n=3) using a modified peroxytitanic
acid colorimetric procedure [103] with a detection limit of 0.1 mg/L. TiSO4 reagent was
from Pfaltz and Bauer Inc., and H2O2 (30% wt. solution in water, reagent grade) was
from Aldrich. Iron was measured in the GAC slurry solution using the Phenanthroline
Method (Method No. 3500-Fe D) [104].
Fe content in GAC: The Fe content in the GAC was measured by digesting
representative samples (0.4 g) of the Fe-amended GAC using 40 mL of 10% nitric acid
for 40 minutes in a microwave oven at 423 K and 1000 kPa. Extracts were analyzed for
metals by an inductively coupled plasma - optical emission spectrometer (Perkin Elmer,
Model Optima 3300 DV, Norwalk, CT).
Scanning Electron Microscopy (SEM)/Energy Dispersive X-ray Spectrometry (EDS): Imaging and microanalysis of GAC particles were conducted
using scanning electron microscopy (SEM) with energy dispersive X-ray spectrometry
(EDS). GAC particles were examined with a JEOL scanning electron microscope
(Model JSM-6360, Montgomery, TX) with a beam current of 10 nA. The microscope is
equipped with an Oxford Instruments energy dispersive X-ray spectrometer (Model
6587) with high- resolution germanium detectors. GAC particles were coated with gold
for SEM and EDS analysis. Two particles were selected from test reactors and two sites
were randomly selected on each particle (n=4). For surface elemental composition
analyses, the surface area interrogated by EDS (47,000 µm2 per site) represented
approximately 0.3-4.2% of the external surface area of 8×30 GAC particles (i.e.,
assuming spherical GAC particles). The beam electron-specimen interactions are used
to derive information on the nature of the specimen [105]. The acceleration voltage was
66
varied (5-30 keV) to assess depth dependent elemental composition of GAC particles.
All SEM-EDS procedures have been previously reported [6].
pH,PZC of the GAC: The pH at point of zero charge (pH,PZC) of the GAC was
determined using the pH drift method [59]. De-ionized (DI) water (50 mL) amended with
NaCl (0.01 M) was placed in 100 mL amber vials and sparged with N2 (10-15 minutes)
to eliminate CO2 and to stabilize pH. The pH was adjusted (pH 2-11) in a series of vials
by adding either HCl or NaOH while purging the headspace with N2. GAC (0.1g) was
added and the vial was capped immediately. The final pH (pHFINAL) was measured in
each of the vials after 48 h and plotted versus the initial pH (pHINITIAL). The pH,PZC was
determined graphically at the intersection of pHFINAL and the line pHFINAL= pHINITIAL.
Analysis of BPA by HPLC: HPLC analysis for the BPA concentration in water
was carried out on a Waters 2690 separations module with a Waters 996 Photo Diode
Array at 235 nm using a Phenomenex Aries Peptide column (3.6 µm XB-C18, 150 x
4.60 mm). The mobile phase was 42% v/v nanopure water and 58% v/v HPLC grade
methanol applied as an isocratic run with a constant flow rate of 1 ml/min. The runs
lasted for ten minutes with a five-minute delay between injections to allow for the
column to equilibrate. Separate calibration curves were created for varying sample
methanol concentrations. To analyze the BPA content in the GAC, 0.1 g of the Fe-GAC
was extracted with 10 mL of methanol for 3 days with periodic shaking each day. The
BPA concentration in methanol was analyzed by the HPLC described as above.
4.4. Result and discussion In this study, the acid-treated Fe-amended GAC (Fe-GAC) was used for
adsorption and oxidation of BPA in water. As Kan and Huling reported [59], the acid-
treated Fe-amended GAC showed quite uniform deposition of Fe (a catalyst for Fenton
oxidation) into the GAC to broaden reactive zone for Fenton oxidation for effective
oxidation of contaminants in the GAC (Figure 4.2). The previous studies [59] showed
quite uniform Fe deposition in the acid-treated GAC).
67
4.4.1. Isotherm, kinetics and thermodynamic analysis for adsorption of BPA onto the Fe-GAC
The adsorption capacities of BPA onto the raw GAC and the Fe-GAC were
determined via equilibrium adsorption experiments. The chemical analysis showed that
the Fe-GAC used for the adsorption isotherm experiments contained 6 mg Fe/g GAC. In
the present study, the Freundlich and Langmuir isotherms were applied to investigate
the overall adsorption efficiency representing two of the most extensively adopted
models for describing adsorption phenomena in aqueous solutions [106]. The isotherm
parameters of the Freundlich and Langmuir models for adsorption of BPA on the raw
GAC and the Fe-GAC at 293 K are listed in Table 4.1. Table 4.1 supports that
Freundlich isotherm model is the better-fitted model than Langmuir model for adsorption
of BPA onto the raw GAC and the Fe-GAC indicating adsorption using a heterogeneous
surface with interaction between adsorbed molecules.. The kf and n for the Fe-GAC was
similar to those for the raw GAC. Others also reported that Freundlich was more
appropriate model for adsorption of BPA onto activated carbons made from coal,
coconut shell and bamboo [107, 108]. Furthermore, the maximum adsorption capacity
of the Fe-GAC for BPA (qm=505.96± 14.0 mg/g) was similar to the raw GAC
(qm=479.63±8.0 mg/g). The result indicates the maximum adsorption capacity of the
activated carbon prepared in this study is the highest among other activated carbon
based adsorbents (qm=62.5 mg/g – 360 mg/g) reported in literature [80, 107, 109, 110].
Temperature is a good measure to determine whether the adsorption mechanism
whether it is an exothermic or endothermic process [111]. When the dynamic adsorption
of BPA onto the Fe-GAC was conducted at various temperatures (293 – 333 K), the
pseudo-first order and pseudo-second order kinetic models in Eqs. 3 and 4 were used
to fit the experimental data (Table 4.2). The pseudo-second order kinetic model was
found to be better fitted to the experimental results, especially the experimental qe value
in Table 4.2. Both the pseudo-first and the pseudo-second order kinetic models showed
excellent fitting to the experimental results. However, the pseudo-second order kinetic
model was found to be better fitted to the experimental results, especially the
experimental qe value in Table 4.2. Besides, it is well-known that a pseudo-second-
order kinetic model includes all the steps of adsorption including external film diffusion,
68
adsorption, and internal particle diffusion which can represents well the adsorption of
BPA into porous media such as activated carbon [80, 107, 112]. Liu et al. (2009, [80])
reported that the pseudo-second order kinetic model represented the adsorption of BPA
onto the nitric acid or thermally treated activated carbon. Tsai et al. (2006, [107]) also
reported that the pseudo 2nd order kinetic model showed the best fit to the experimental
data for adsorption of BPA onto the granular activated carbon made from coconut and
coal. Thus, the experimental and theoretical analysis confirmed that Freundlich isotherm
and pseudo second order kinetic models can represent the adsorption of BPA onto the
acid-treated Fe-amended GAC well.
Figure 4.3 shows that the pseudo second order rate constant for BPA
adsorption rate on the Fe-GAC increased at increasing temperature (293 – 333 K)
because a higher temperature could decrease the solution viscosity and thus promoted
the mass transfer and diffusion of BPA molecules [113, 114]. The diffusivity of BPA in
water at various temperature (293 – 333 K) was calculated using the Wilke-Chang
equation in Eq. 7 [115] (Table 4.3).
DC,W = T×7.4×10-8 (ΦW MW)1/2 / (µW ×VC 0.6) (7)
where DC,W = diffusivity of chemical C in water (cm2/s at 1 atm), µW = viscosity of water
(centipoise), T = absolute temperature (K), MW = molecular weight of water, VC = molar
volume of a chemical at normal boiling point (266 and 27.2 cm3/g-mole for BPA and
H2O2, respectively), and ΦW association parameter for water (2.26, dimensionless)
[115].
The inset in Figure 4.3 indicates that enhancement of adsorption rate of BPA
was linearly proportional to increase in diffusion coefficient of BPA over the
temperatures. It is consistent with the fact that adsorption process is limited by diffusion
of adsorbate assuming that adsorption of a target contaminant onto the GAC is much
faster than diffusion of a target contaminant [59, 116].
The thermodynamic parameters such as changes in free energy (ΔG0), enthalpy
(ΔH0), and entropy (ΔS0) calculated using Eqs. 5 and 6 are listed in Table 4.4. The
negative values of ΔH0 indicated that the adsorption process of BPA is exothermic and,
69
occurs mainly by physical process. The positive values of ΔS0 suggested that increased
randomness at the solid-solution interface occurs in the internal structure of the
adsorption process of BPA. The negative values of ΔG0 indicated that the adsorption
process of BPA is spontaneous while the spontaneity increases at higher temperature.
Similar results for exothermic and physical adsorption of BPA were observed from other
studies on activated carbon [80, 110].
4.4.2. Fenton oxidation of BPA in the Fe-GAC In Fenton process, the optimum concentration of hydrogen peroxide is important
for determining operation costs as well as enhancing treatment efficiency [62, 63].
Figure 4.4 shows the effect of H2O2 concentration on the oxidation of BPA. An increase
of molar ratio of [H2O2]: [BPA] from 9 to 108 enhanced BPA oxidation efficiency from
62.3% to 99.3% at the fixed loading of Fe (6 mg Fe/g GAC). This trend can be
explained by higher oxidation of BPA with more production of ⋅OH at higher molar ratio
of [H2O2]:[BPA] [117]. The drastic enhancement of BPA removal occurred up to
[H2O2]:[BPA] of 36 which is the stoichiometric demand of H2O2 for complete oxidation of
BPA resulting in 91% removal (Figure 4.3).
In addition, the effects of temperature on Fenton oxidation of BPA in the Fe-GAC
were investigated when the reaction temperature was varied from 293 K to 333 K
(Figure 4.5). Negligible loss of BPA was observed over the reaction temperatures due
to poor vapor pressure and volatility of BPA even at high temperature. The BPA
oxidation rate (mg BPA removed per g of GAC per h) and H2O2 decay constant (min-1),
which were observed for 1 h reaction, were drastically enhanced by the factor of 2.5 and
5 when the reaction temperature increased from 293 K to 333 K. Thus, increasing
temperature clearly enhanced the H2O2 reaction to increase ⋅OH production and BPA
oxidation. These results are consistent with those reported by Kan and Huling [59] when
the Fenton oxidation of MTBE in the GAC was enhanced at increasing temperature.
Unlike the BPA oxidation rate, there were similar removal efficiencies of BPA (78% at
293 K, 88% at 313 K and 90% at 333 K) at various temperatures. Similar BPA removal
efficiencies at various temperatures were thought to be mainly due to different reaction
70
time at which most of H2O2 was consumed (please note that complete consumption of
H2O2 occurred 30 h for 293 K and 6 h for 333 K).
Figure 4.6 Information summarizes the Fenton oxidation pathway proposed by
Poerschmann et al. (2010) [118], Luo et al. and Hua et al. [119, 120]. It was reported
that the Fenton oxidation of BPA at low dose of H2O2 generated the aromatic products
such as hydroquinone, 4’-hydroxyacetophenone, 4’-isoproenylphenone, and catechol as
the major oxidative products. On the contrary, the Fenton oxidation of BPA at high dose
of H2O2 (i.e., above the stoichiometric demand of H2O2) led to generation of the highly
soluble organic acids such as oxalic acid, acetic acid, lactic acid, and fumaric acid.
Although a detailed analysis of intermediates and oxidative products of BPA was not the
scope of this study, our analysis showed that the Fenton oxidation of BPA at the molar
ratio of [H2O2]:[BPA] above 18 produced mostly the soluble organic acids and negligible
concentration of the aromatic products (data not shown).
Previously, an extensive investigation on the impact of iron amendment and
Fenton oxidative treatments on GAC surface area and pore volume was investigated
[6]. Specifically, this involved the same GAC as used in this study and correlations were
established between the concentration of iron amended to the GAC and the surface
area and pore volume of the GAC. Based on these quantitative correlations [6],
assuming approximately 6 g/kg iron amendment to the GAC resulted in approximately a
1.3% and 1.7% decline in surface area (m2/g GAC) and pore volume (mL/g GAC),
respectively. Given the 95% confidence interval (C.I.) for the pre-treated average
surface area 1190 m2/g (95% C.I. 1110-1270 m2/g), and the average pore volume 0.59
mL/g (95% C.I. 0.52-0.66 mL/g), these changes were not considered significant.
Further, post-oxidation changes in GAC surface area and the pore volume resulting
from two full cycles of aggressive Fenton oxidative treatment resulted in a 2% decline of
each parameter [97]. The loss in surface area and pore volume in this case was partially
attributed to the residual contaminant, methyl tert-butyl ether (MTBE), on the GAC.
Based on these results that involved similar GAC, Fe amendment methods, and Fe
concentration, it is reasonable to assume a negligible loss in surface area and pore
volume resulted from the oxidative treatment in this study.
71
4.4.3. Comparative analysis of temperature-dependent diffusion, desorption and Fenton oxidation of BPA in the Fe-GAC
To further understand the role of the intraparticle diffusion limitation as the major
limiting step in the Fenton oxidation of BPA in the Fe-GAC three values were compared:
the temperature-dependent relative increase in calculated diffusion of BPA; the
measured values of BPA desorption + diffusion; and Fenton-driven oxidation rate of
BPA (Figure 4.7). First, the calculated diffusivity of BPA in water using the Wilke-
Chang equation (Table 4.3) was varied by nearly a factor of 2.4 over the temperature
range (293 – 333 K) used in these experiments. BPA desorption + diffusion values
measured using the fill and draw procedures, also increased at the elevated
temperatures. However, the enhancement of BPA desorption + diffusion was higher
than that of BPA diffusion indicating that both diffusion and desorption of BPA increased
at high temperature. However, the BPA oxidation rate in the Fe-GAC exhibited a slope
quite similar to that of BPA diffusion (Figure 4.7). This suggests that the BPA oxidation
in the Fe-GAC is mainly controlled by diffusive transport of BPA from the Fe-GAC.
The dimensionless Thiele-modulus number (Φ, Eq. 8) was used to analyze the
degree of intraparticle H2O2 diffusion limitation associated with H2O2 reaction in GAC
[121].
Φ = RGAC × (kH2O2 / Deff)1/2 (8)
where RGAC = radius of GAC particle (cm), kH2O2 = H2O2 decay kinetic constant (s-1), and
Deff = effective diffusivity of H2O2 (cm2 s-1).
In Eq. 8, the diffusivity of H2O2 in the Fe-GAC (DH2O2,GAC) is calculated by
amplifying the molecular diffusion coefficient obtained from Wilke Chang equation (Eq.
7) with ε/τ (ε: the internal void fraction of the pellet, τ: the tortuosity). Tortuosity is the
ratio of the actual transport distance to the shortest (direct) pathway distance. In this
study the tortuosity value of 10 was taken which is a typical value in granular activated
carbons [121]. Table 4.5 shows the Thiele-modulus numbers (Φ) calculated by Eq. 8 for
small GAC particle (600 µm) and large GAC particle (2400 µm) since the GAC used for
72
this study ranges from 600 µm to 2400 µm (8 x 30 mesh size). The Thiele-modulus
analysis indicated: Φ = 0.25 and 0.36 for small GAC particles (dGAC = 600 µm, mesh
size of 30), and Φ = 1 and 1.43 for large GAC particles (dGAC = 2400 µm, mesh size of
8) at 293 K and 333 K. The effectiveness factor (η) is the ratio of the actual H2O2
reaction and the ideal H2O2 reaction (assuming no diffusional limitations) and is a
measure of the reduction in reaction rate attributed to H2O2 diffusion limitations [122].
The effectiveness factor is calculated as function of Thiele-modulus number
using Eq. 9 in case of a spherical porous particle with first order reactions [122].
η = 3 x (Φ coth Φ -1)/Φ2 (9)
For example, Φ « 1, η = 1 indicates no pore diffusion limitations and no reduction
in reaction, and Φ » 1, η = 1/Φ indicates strong pore diffusion limitations [122]. In our
study, it is estimated that negligible diffusion limitation of H2O2 occurred in small GAC
particles (dGAC = 600 µm; 293 – 333 K) while there was some pore diffusion limitation of
H2O2 and reduction in H2O2 reaction in large GAC particles (dGAC = 2400 µm; 293 – 333
K). This analysis presented that the reduction of H2O2 reaction associated with diffusion
limitation of H2O2 in the GAC increased with increasing the reaction temperature and
the GAC particle size (Table 4.5). These results are consistent with others’ results
which showed H2O2 diffusion limitation in large GAC particles [96]. Thus, enhanced BPA
oxidation and removal in GAC measured at higher solution temperature was mainly
attributed to an increase in BPA and H2O2 diffusion.
4.5. Acknowledgements This work was supported by University of Hawaii (Project number: 2300016).
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79
Table 4.1. Freundlich and Langmuir isotherm model coefficients for adsorption of BPA on the GACs. Conditions: Initial BPA (100-250 mg/L, 0.05-0.15 L); virgin GAC and Fe-GAC (0.025-0.1 g); initial solution pH, 6; 293 K.
Samples Freundlich Langmuir kf n R2 qmax
(mg/g) KL
(L/g)
R2
Virgin GAC
169
3.74
0.97
479
0.38
0.973
Fe-GAC
171
4.70
0.98
466
0.23
0.929
80
Table 4.2. Adsorption kinetic parameters for BPA adsorption onto Fe-amended GAC. Conditions: Initial BPA (120 mg/L, 0.1 L); 0.1 g Fe-GAC; initial solution pH, 6; 293K.
C0 (mg/L)
Experimental qe
(mg/g)
Pseudo first order Pseudo second order
qe (mg/g)
k1 (min-1)
R2 qe (mg/g)
k2 (g/mg/min)
R2
120
118
95
0.0073
0.993
128
4.8 x 10-5
0.992
81
Table 4.3. BPA diffusivity at various temperatures.
Temperature
(K)
Viscosity of
water
(centipoise)
BPA Diffusivity in water
(×10-5 cm2/s)1
Overall increase in BPA diffusivity 2
H2O2 Diffusivity
in water (×10-5 cm2/s)1
293
1.00 0.48
0.79
1.18
1
1.6 (1.6)
2.4 (1.5)
1.83
313
0.65
3.15
333
0.47
4.47
1 Diffusivity calculated from Eq. 9. 2 Relative increase in BPA diffusivity from 298 K in parenthesis.
82
Table 4.4. Thermodynamic parameters for BPA adsorption onto Fe-amended GAC. Conditions: Initial BPA (120 mg/L, 0.1 L); 0.1 g Fe-GAC; initial solution pH, 6. Temperature
(K) Qe
(mg/g) Ce
(mg/L) Qe/Ce (Keq)
ΔG (KJ/mole)
ΔH (KJ/mole)
ΔS (J/mol-K)
293
128
0.28
457
-46.1
-20.6
87.2
313
124
0.24
517
-47.9
333
116
0.19
611
-49.6
*Qe: Amount of BPA adsorbed per unit mass of adsorbent at equilibrium (mg/g) Ce: Liquid-phase concentration of BPA at equilibrium (mg/L) Keq: Equilibrium constant
83
Table 4.5 Thiele-modulus analysis for the Fenton oxidation of BPA-spent GAC at increasing temperatures.
GAC particle (µm)
Temperature (K)
Thiele-modulus number (Φ)
Effectiveness factor (η)
600
293
0.79
0.75
313 0.90 0.71 333 1.13 0.63
2400
293
3.16
0.32
313 3.59 0.28 333 4.53 0.22
84
Figure. 4.1. The proposed mechanisms in Fenton oxidation of BPA (contaminant) in the acid-treated Fe-amended granular activated carbon (Fe-GAC): (1) Desorption of BPA from the surface of GAC, (2) BPA diffusive mass transport (pore + surface), intraparticle diffusive transport of H2O2, (3) BPA diffusive transport outward through the quiescent film, H2O2 diffusive transport from the bulk solution through the quiescent film into the GAC pores, and (4) BPA and H2O2 mixing in bulk solution. Further, H2O2 reacts with immobilized Fe (Fenton-like reaction) resulting in the formation of ·OH and oxidation of BPA by ·OH. Fenton oxidation of BPA in the Fe-GAC involves the simultaneous occurrence of these multiple mechanisms [22].
85
Figure 4.2. Atomic chemical composition of Fe in acid-treated and untreated Fe-amended GAC. The radius of the GAC particles (300-1200 µm) represents the range in GAC particles passed through 8×30 sieves. Measurements were by SEM/EDS and depth-dependent Fe was derived by varying the accelerating voltages (5, 10, 20, 30 keV) [22].
86
Figure 4.3. Adsorption kinetics of BPA onto the Fe-GAC at various temperatures. Conditions: Initial BPA (120 mg/L, 0.1 L); 0.1 g Fe-GAC; initial solution pH, 6.
0.000
0.002
0.004
0.006
0.008
0.010
273 293 313 333
Ads
orpt
ion
kine
tic c
onst
ant
(g G
AC
/mg
BPA
-h)
Temperature (K)
Adsorption kinetic constant
0
1
2
3
273 293 313 333
Enha
ncem
ent (
-)
Temp. (K)
Diffusion Adsorption
87
Figure 4.4. Effect of [H2O2]: [BPA] on BPA removal. Conditions: 0.1 g BPA-saturated Fe-GAC (320 ± 20 mg BPA/g Fe-GAC); initial solution pH, 3; 293 K.
0
10
20
30
40
50
60
70
80
90
100
0 10 20 30 40 50 60 70 80 90 100 110
Oxi
datio
n ef
ficie
ncy
(%) o
f BPA
in F
e-G
AC
Mol H2O2/Mol BPA
88
Fig 4.5. Effect of temperature on Fenton oxidation of BPA in the GAC. Conditions: 0.1 g BPA-saturated Fe-GAC (320 ± 20 mg BPA/g Fe-GAC); initial solution pH, 3; [H2O2]/ [BPA], 18.
0
0.005
0.01
0.015
0.02
0.025
0
50
100
150
273 293 313 333 353
Oxi
datio
n ra
te o
f BPA
(mg
BPA
/g G
AC
-h)
H2O
2 dec
ay c
onst
ant (
min
-1)
Temperature (K )
BPA oxidation rate H2O2 reaction rate
89
Figure 4.6. Proposed pathway for the Fenton oxidation of BPA [1-3].
90
Fig 4.7. Temperature-dependent relative increase in calculated diffusion of BPA, measured values of BPA desorption + diffusion, and Fenton-driven oxidation rate of BPA in the Fe-GAC.
0
1
2
3
4
273 293 313 333
Incr
ease
rela
tive
to 2
93 K
Temperature (K)
BPA desorption + diffusion BPA diffusion BPA oxidation rate
91
CHAPTER 5. UV Photocatalytic Oxidation of Sulfamethoxazole Using TiO2
Supported on Biochar
5.1. Abstract The presence of drugs in wastewater has recently received high levels of
attention as contaminants of concern in the environment. Sulfamethoxazole (SMX), a
sulfonamide bacteriostatic antibiotic, is a chemical that has been extensively used for
the treatment and prevention of both human and animal diseases. Due to the antibiotic-
resistant pathogens and toxic effects on the aquatic ecosystem and human health
caused by these antibiotics at low levels, developing effective method to degrade these
compounds to non-toxic and pharmaceutically inactive byproducts. The photolytic and
photocatalytic degradation of SMX was carried out in aqueous suspension using
biochar supported TiO2 (biochar-TiO2) catalyst under UVC radiation (𝜆 = 254nm). The
photocatalytic oxidation using biochar-TiO2 resulted in similar removal percent of SMX
to photolytic oxidation; however, the higher removal of chemical oxygen demand (COD)
of SMX was achieved indicating further degradation of SMX byproducts. The impacts of
photocatalyst concentration, nitrate (0.1 M – 1 M) and bicarbonate (0.02 M – 1 M)
concentration and reaction time were examined in this study.
5.2. Introduction
Due to the widespread use of antibiotics as medicines and animal growth
promoters (~16,600 ton/yr in the U.S.), these chemicals have recently received high
levels of attention as contaminants of concern in animal farms, aquatic environments,
food safety, and human health [1, 2]. For instance, sulfamethoxazole (SMX), a
sulfonamide bacteriostatic antibiotic, is a chemical that has been extensively used for
the treatment and prevention of both human and animal diseases [17]. It has been
ubiquitously found in the high ng/L-µg/L range in discharges from WWTPs and in the
low ng/L range in rivers and groundwater [12, 15, 18]. SMX is relatively unreactive in
soils; thus, instead of biodegrading, it has high mobility in soils [19]. If released into
aquatic systems through discharges from WWTPs, SMX may cause toxic effects in
aquatic organisms and induce antibiotic-resistant pathogens [20, 21]. SMX has been
92
shown to cause acute and chronic responses in the range of 100 ppm and 7-10 ppm,
respectively [3]. Due to the antibiotic-resistant pathogens and toxic effects on the
aquatic ecosystem and human health caused by these antibiotics at low levels,
pharmaceutical compounds from wastewater and water need to be effectively removed.
To date, several methods such as biological treatment, advanced oxidation,
membrane separation, and adsorption have been studied for the removal of SMX in
wastewater and water [3-12]. Unfortunately, these conventional wastewater and water
treatment processes cannot achieve the complete or effective removal of SMX due to its
low concentrations and biorefractory properties [13, 14]. Current biological treatment
cannot be used for the effective degradation of SMX due to the compounds’ toxicity and
biorefractory properties [5, 9]. Membrane separation is a promising technology for the
excellent removal of micropollutants in water; however, high operation costs still limit its
full-scale use. Membranes easily suffer from (bio) fouling problems that could result in
unexpected interruptions during the treatment of aqueous contaminants [15, 16].
Adsorption is simple and effective, but it requires high operating costs associated with
the regeneration of contaminant-spent adsorbents (i.e., activated carbon) [17, 18]. The
current thermal regeneration method requires high operating costs, and leads to
significant reduction of re-adsorptive capacity of activated carbon due to significant
deterioration of carbon pore structure, specific surface area and functionality [17, 18].
Compared with other treatment methods, advanced oxidation processes (i.e., ozone,
O3/H2O2, H2O2/UV, Fenton, ultrasound, photocatalytic and electrochemical oxidation)
can achieve effective degradation of SMX [19-25]. However, advanced oxidation
process (AOP) often generates toxic intermediate/byproducts under suboptimum
conditions, and the AOP of low concentration of SMX would result in low reaction rate
and removal efficiency [26, 27]. The short lifetime of the oxidants requires use of large
amounts of oxidants to oxidize a trace level of SMX leading to increasing operation
costs [18].
Among various AOPs, the photocatalyst such as TiO2 have shown highly efficient
degradation of various SMX under UV light irradiation [28]. !OH and electron holes
generated from UV light-excited TiO2 effectively can oxidize various SMX via direct or
indirect routes [29]. However, current UV light-driven photocatalytic degradation of SMX
93
exhibited critical shortcomings to limit its practical application in the field such as: 1)
high operating/maintenance costs associated UV lamps, 2) low mineralization of the
SMX (degraded to CO2 and H2O), 3) low reactivity with low concentration of SMX in
water, 4) difficult to separate the photocatalyst after the reactions, and 5) photochemical
reactivity significantly influenced by water quality (i.e., turbidity, radical scavengers) [30].
Thus, a cost-effective photocatalytic degradation of SMX in reclaimed water needs to be
developed.
For the present study, biochar supported-TiO2 was used as photocatalyst for
photocatalytic oxidation of SMX under UVC light. Recent studies reported UV light
mediated photocatalytic oxidation using TiO2 [31-34]. Photocatalytic oxidation using
TiO2 under UV light led to 80 – 100% degradation of SMX. However, the results from
other studies indicated that the photocatalytic degradation by TiO2 showed low
mineralization of SMX, and generating more toxic byproducts than the parent compound
[31, 34]. In addition, TiO2 powder is difficult to recycle and recover from the solution and
easy to agglomerate [35]. For these reasons, several approaches have been examined
by immobilizing TiO2 on a supporting material such as carbon nanotube, activated
carbon, zeolite and alumina [36-39].
Therefore, the objective of this study was to effectively remove SMX in water by
photocatalytic oxidation process using the biochar-TiO2. To the best of our knowledge,
there have been no studies for photocatalytic oxidation of SMX using Biochar-TiO2.
Furthermore, this study focus on the effect of the catalyst loading, nitrate and carbonate
concentration and exposure time on photocatalytic oxidation and total organic carbon
(TOC) reduction rate of SMX in biochar-TiO2 to enhance overall treatment efficiency of
SMX.
The possible mechanisms and oxidative products associated with photolytic and
photocatalytic oxidation of SMX in the biochar-TiO2 at the different experimental
conditions is addressed through the execution of the experimental procedures, and the
critical analysis and discussion of the results.
94
5.3. Materials and method 5.3.1. Chemical and reagent
Biochar was obtained from Hawaii Natural Energy Institute in Honolulu, HI, and
used as supports of TiO2. The biochar was rinsed with deionized (DI) water, and dried in
an oven at 70 ℃ , and stored in a desiccator until used. Titanium isopropoxide
(Ti(OCH(CH3)2)4), 97%, Sigma-Aldrich) was used as the Ti precursor without any further
purification. Ethanol (C2H6O, 100%) and hydrochloric acid (HCl, 37% v/v) were
purchased from Fisher Scientific (Waltham, MA, USA). Sulfamethoxazole (SMX),
sodium bicarbonate (NaHCO3), and sodium nitrate (NaNO3) were purchased from
Sigma-Aldrich (St. Louis, MO, USA).
5.3.2. Preparation and characterization of biochar-TiO2
The TiO2/Biochar hybrids were prepared using modified methods [37]. Biochar
was washed with deionized water, and dried in an oven at 70 °C for 12 h. 5 g biochar
was treated by stirring in a flask that containing 0.3 M nitric acid for 24 h. After filtration,
the acid treated-biochar was washed with DI water multiple times. The acid treated-
biochar was dried in an oven at 70 °C for 24 h and stored in a desiccator for further use.
2.5 g acid treated-biochar was dispersed in 60 mL ethanol, and 20 mL titanium
isopropoxide was added. The mixture was stirred at room temperature for 1 h before
adding a solution containing 8 mL 37% (v/v) HCl and 20 mL ethanol with constant
stirring for 1 h. The solution was vacuum filtered, then washed with ethanol before
drying in the oven at 100 °C for 24 h. The dried sample was milled and heated at 325 °C
for 1 h to yield the biochar –TiO2.
The scanning electron microscopy (SEM) images and X-ray energy dispersive
spectroscopy data (EDX) were obtained with Hitachi HT7700 field emission scanning
electron microscope (10 kV) with an Oxford INCA PentaFET-x3 Si (Li) EDX detector
system (Oxford instruments, Oxford, United Kingdom) for EDX analysis. Powder X-ray
diffraction (XRD) patterns of the powered sample were performed on a Bruker D8
Discover (Bruker, Billerica, MA, USA) with a Cu Kα radiation source.
95
5.3.3. UV light mediated photocatalytic oxidation of SMX using the biochar-TiO2 photocatalyst
All of the experiments were conducted in a glass reactor (12.5 cm diameter, 6.5
cm height) filled with 100 mL of 10 mg/L SMX solution containing 0.5 g of the
photocatalyst (biochar-TiO2) at 20 ± 2 ℃ under UVC irradiation. Suspension pH was
adjusted by adding hydrochloric acid or sodium hydroxide. Unless otherwise stated, all
experiments were performed at pH 4. A 15 W germicidal UCV lamp (Nuaire, Plymouth,
MN, USA) was used as UVC light source to irradiate horizontally into the glass reactor.
The lamp was placed at 20 cm from the glass reactor. Before turning the lamp, the
suspension was magnetically stirred for 30 min to establish the adsorption-desorption
equilibrium under dark condition.
The photocatalytic oxidation of SMX was initiated by turning on the lamp, and the
mixture was magnetically stirred. The effect of the catalyst loading (0.25 – 1 g) and
reaction time (1 – 6 h) on the oxidation of SMX was conducted for finding out the
optimum condition in the photocatalytic of SMX.
Besides, the photocatalytic oxidation (biochar-TiO2 photocatalyst, UVC light) and
the photolytic oxidation (UCV light) were performed at the selected conditions (100 mg/L
SMX, UVC light irradiation, initial pH of 4, temperature of 20 ± 2 ℃). The scavenging
tests with sodium formate and ethanol were also carried out for finding out the major
mechanism associated with the photocatalytic oxidation under UVC light.
Oxygen uptake rate was performed by using 10 mL of byproducts from different
treatments and mixed bacteria culture (OD:2.2-2.5). The pH of 10 mL solution was
adjusted to 7, and oxygen was saturated for 30 sec. After O2 saturation, mixed bacteria
culture was injected into the solution, and magnetically stirred in a closed system. The
concentration of oxygen in the solution was measured using oxygen gas sensor
(Vernier).
5.3.4. Analytical methods
HPLC analysis for SMX concentration in water was carried out on a Waters 2690
separations module with a Waters 996 Photo Diode Array at 235 nm using a
Phenomenex Aries Peptide column (3.6 µm XB-C18, 150 x 4.60 mm). The mobile
96
phase was 67 % v/v nanopure water (0.1% v/v formic acid) and 33 % v/v HPLC grade
methanol which was applied as an isocratic run with a constant flow rate of 0.75 ml/min.
The runs lasted for ten minutes with a three-minute delay between injections to allow for
the column to equilibrate. Separate calibration curves were created for varying sample
methanol concentrations.
Chemical oxygen demand (COD) was measured with Chemetrics low range (10-
150 mg/L) COD vials (Midland, VA, USA). The COD removal efficiency as an indication
for mineralization were calculated by Eq. 1:
Mineralization (%) = 1− !"# !!"# !
× 100 (1)
Where [COD]t is the aqueous concentration of COD at a given time (t), and [COD]0 is
the initial aqueous concentration of COD.
Total organic carbon (TOC) was measured with Total Organic Carbon Reagent
Set, Low Range (0.3 – 20.0 mg/L) TOC vials (Loveland, CO, USA). The TOC removal
efficiency as an indication for mineralization was calculated by Eq. 2:
Mineralization (%) = 1− !"# !!"# !
× 100 (2)
Where [TOC]t is the aqueous concentration of TOC at a given time (t), and [TOC]0 is the
initial aqueous concentration of TOC.
5.4. Results and discussion 5.4.1. Characterization of the biochar-TiO2
The SEM images of biochar and the biochar-TiO2 in Figure. 5.1a and Figure 5.1b present the structure of the biochar and biochar-TiO2, respectively. The images
exhibited the biochar attached with TiO2 indicated well-dispersed TiO2 on biochar with
little agglomeration. The XRD patterns of the TiO2 and biochar-TiO2 are presented in
Figure 5.2. The diffraction peaks at 2θ=25.3, 37.8, 48.1 and 54.1 in the biochar-TiO2
were attributed to anatase-TiO2 [41]. The XRD results confirmed that the anatase TiO2
97
in the biochar-TiO2. The EDX analysis of the biochar-TiO2 revealed strong peak for TiO2
(the data not shown).
5.4.2. UV photocatalytic oxidation of SMX Figure 5.3. shows the results of the control experiment for the photocatalytic
oxidation of SMX using the biochar-TiO2 in comparison to direct photolysis and
photocatalytic oxidation using commercial anatase nano-TiO2 (0.1 g and 0.5 g). The
direct photolysis of SMX under UVC led to nearly complete removal of SMX in 3h (~ 100
%) compared with the photocatalytic oxidation (75%) using the biochar-TiO2 and 0.1g
and 0.5 g of the commercial TiO2 (~ 100 %). However, the COD removal of SMX was
the highest when the biochar-TiO2 was used indicating its effective mineralization.
Although SMX removal was higher in the direct photolysis than the photocatalytic
oxidation using the biochar-TiO2, the presence of the biochar-TiO2 leads to oxidize more
persistent products formed during photo oxidation. Such the results were consistent to
those as reported (87% COD removal during photocatalysis and 24% COD removal for
photolysis) by Nasuhoglu et al and Abellán et al [31, 34].
The effect of the catalyst loading on the photocatalytic degradation of SMX is
shown in Figure 5.4. The removal percent of SMX was decreased as the biochar-TiO2
concentration is increased under UVC irradiation. The SMX reduction percent was
decreased because as soon as SMX started to be degraded, both byproducts from
SMX and biochar-TiO2 can compete with SMX for the hydroxyl radical [34]. However,
the COD removal percent enhances as the biochar-TiO2 concentration is increased.
The enhancement of COD removal percent occurs at higher catalyst loading is due to
generation of more OH radical in the presence of the higher amount of biochar-TiO2.
The effect of nitrate concentration on the photocatalytic degradation of SMX is
shown in Figure 5.5. The results indicated that SMX degradation is enhanced in the
presence of higher nitrate ion. The percent of SMX removal was increased from 75% to
90%. Nitrate is abundant in natural water, and it has strong adsorption below 250 nm
and it is known to photochemically produce !OH [42]. Also, it is known nitrate is known
to be an electron scavenger, and it reacts very easily with the generated electron in the
surface of the biochar-TiO2, reducing the undesirable electron-hole recombination [43].
98
Figure 5.6 shows that SMX degradation is improved by the presence of
bicarbonate ion at the solution pH 9. The SMX degradation was increased from 75% to
93%. Other studies reported that the carbonate radical (CO3!-) is a reactive intermediate
that likely plays an important in oxidation reactions [44-46]. Compared to OH radical
(2.72 and 1.89 V vs. SHE at pH 0 and 14), carbonate radical (1.59 V vs. SHE at pH =
126) is a weaker oxidant and reaction rate is slower. The reaction with OH radical with
carbonate (eq (3)) and bicarbonate anion (eq (4)) is considered to be a major source of
carbonate radical in aquatic environment [44].
CO32- + !OH → CO3!
- + OH- (3)
HCO3- + !OH → CO3!- + H2O (4)
Previous studies showed that aniline compounds are susceptible to efficient oxidation
by carbonate radical, and SMX contains an aniline moiety [33, 44]. Therefore, the higher
NaHCO3 concentration led to more effective SMX degradation.
Figure 5.7 represents the effect of reaction time on the removal of SMX and
COD. The data indicates that most of SMX is removed within 0.5 h, and the
intermediates produced (HPLC data is not shown) are degraded as the reaction time
increases. This result can be supported by analyzing the COD removal efficiency which
is increased from 47% to 85% with reaction time of 0.5 h to 6 h, respectively. However,
the SMX removal percent (86-91%) did not significantly increase with reaction time.
Oxygen uptake rate (OUR) for UV or UV+Biochar-TiO2 byproducts is analyzed in
order to estimate the biodegradability after UV treatment (Figure 5.8). Compare to
endogenous OUR, the OUR with UV+biochar-TiO2 showed the fast and highest oxygen
removal in the result which might indirectly indicate it can be biodegradable. On the
other hand, the result of OUR with UV treatment alone shows slight decrease in OUR. It
can be explained that UV treatment alone lead to generate more toxic or persistent
byproducts [31].
99
5.5. Acknowledgement This work was supported by University of Hawaii at Manoa (Project number:
2300016).
5.6. References [1] D. Avisar, Y. Lester, D. Ronen, Sulfamethoxazole contamination of a deep
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105
Figure 5.1. SEM images of (a) biochar and (b) biochar-TiO2.
(a)
(b)
106
Figure 5.2. XRD patterns of (a) TiO2 (b) biochar and (c) biochar-TiO2.
(a)
(b)
0 20 40 60 80 100
Inte
nsity
(a.u
.)
2θ (degree)
0 20 40 60 80 100
Inte
nsity
(a.u
.)
2θ (degree)
107
Figure 5.3. Removal of SMX and COD using photolysis and photocatalysis. Conditions: [SMX]0, 10 mg/L; volume of SMX solution, 100 mL; catalyst (biochar-TiO2) loading, 0.5 g; initial pH,4; reaction time, 3 h; temperature, 20 ± 2 ℃.
0
10
20
30
40
50
60
70
80
90
100
SMX + UV SMX + UV + TiO2 (0.1g)
SMX + UV+ TiO2 (0.5g)
SMX + UV + BT (0.5g)
SMX/
CO
D re
mov
al e
ffici
ency
(%)
SMX removal
COD removal
108
Figure 5.4. Effect of the catalyst loading on removal of SMX and COD. Conditions: [SMX]0, 10 mg/L; volume of SMX solution, 100 mL; catalyst (biochar-TiO2) loading, 0.25 g – 1 g; reaction time, 3 h; temperature, 20 ± 2 ℃.
0
10
20
30
40
50
60
70
80
90
100
0 0.2 0.4 0.6 0.8 1 1.2
SMX/
CO
D re
mov
al e
ffici
ency
(%)
Catalyst loading (g)
SMX removal
COD removal
109
Figure 5.5. Effect of sodium nitrate on removal of SMX. Conditions: [SMX], 10 mg/L; volume of SMX solution, 100 mL; catalyst (biochar-TiO2) loading, 0.5 g; [NaNO3]0, 0.1 – 1 M; reaction time, 3 h; temperature, 20 ± 2 ℃.
0
10
20
30
40
50
60
70
80
90
100
0 0.5 1 1.5 2 2.5 3 3.5 4 4.5
SMX
rem
oval
effi
cien
cy (%
)
NaNO3 concentration (M)
SMX removal
110
Figure 5.6. Effect of sodium bicarbonate on removal of SMX. Conditions: [SMX], 10 mg/L; volume of SMX solution, 100 mL; catalyst (biochar-TiO2) loading, 0.5 g; [NaHCO3]0, 0.02 – 0.1 M; reaction time, 3 h; temperature, 20 ± 2 ℃.
0
10
20
30
40
50
60
70
80
90
100
0 1 2 3 4 5
SMX
rem
oval
effi
cien
cy (%
)
NaHCO3 concentration (M)
SMX removal
111
Figure 5.7. Effect of reaction time on removal of SMX. Conditions: [SMX], 10 mg/L; volume of SMX solution, 100 mL; catalyst (biochar-TiO2) loading, 0.5 g; reaction time, 1 - 6 h; temperature, 20 ± 2 ℃.
0
10
20
30
40
50
60
70
80
90
100
0 1 2 3 4 5 6 7
SMX
rem
oval
effi
cien
cy (%
)
Reaction time (h)
SMX removal
COD removal
112
Figure 5.8. The result of oxygen uptake rate of byproduct formed from SMX+UV and SMX+ UV+Biochar-TiO2
0
10
20
30
40
50
60
70
80
90
100
0 2 4 6 8 10 12
Dis
solv
ed o
xyge
n (%
)
Time (min)
Endogenous
SMX + UV
SMX + UV + Biochar-TiO2
113
Chapter 6. Conclusion
In chapter 2, the present study presents that the acid-treated iron-amended
activated carbon (Fe-GAC) is an effective adsorbent and catalyst for the adsorption and
oxidation of methylene blue from aqueous solution. The modification of carbon surface
by acid treatment has proven to play an important role for more uniform deposition of
iron in the GAC. The equilibrium data were fitted to linear models of Freundlich and
Langmuir model, and the equilibrium data of Fe-GAC were best described by the
Langmuir isotherm model with the maximum monolayer adsorption capacity of
238.1m ± 0.78 mg/g at 20 ℃. Adsorption kinetic studies of MB demonstrated that the
adsorption rate followed the pseudo-second-order kinetic model. For the heterogeneous
Fenton oxidation of MB in the Fe-GAC, the increase of H2O2 concentration from 7 to
140 mmol H2O2/ mmol MB led to an increase in MB removal efficiency from 62.6 % to
100%. Besides, the multiple application of H2O2 to the MB-saturated Fe-GAC increased
the oxidation efficiency of MB in the GAC up to 84.1% which was more effective than
that by the single application of H2O2.
In chapter 3, the CdS/MWCNT-TiO2 composite catalyst was synthesized using
the sol-gel method for the photo-Fenton oxidation of methylene blue under visible light.
The characterization of the catalyst confirmed well-dispersed mixture of CdS (6-10 nm),
TiO2 (6.0-8.5 nm) and MWCNT with little agglomeration which attached on the MWCNT.
It also indicated the hexagonal structure of CdS and the anatase TiO2 in CdS/MWCNT-
TiO2 composite. The experimental results presented that the solution pH, the initial H2O2
and Fe3+ concentration had strong influences on the decolorization and mineralization of
methylene blue. The photo-Fenton oxidation led to high decolorization and
mineralization (98% and 83% in 30 min) at the optimum conditions (30 mg catalyst, 50
µM methylene blue, 180 µM Fe3+, 600 µM H2O2, initial pH of 3.5 and temperature of
20±2oC). The decolorization and mineralization by the photo-Fenton oxidation was
much higher than the dark Fenton oxidation and the photocatalysis alone. The molar
ratio of [H2O2]/[methylene blue] in the photo-Fenton oxidation also indicated cost-
effectiveness of the photo-Fenton oxidation compared with other Fenton and photo-
Fenton oxidation of methylene blue. The reusability tests confirmed the stability of the
catalyst over the multiple oxidation of methylene blue. Therefore, the experimental
114
results clearly indicated that the photo-Fenton oxidation process was effective for the
degradation of methylene blue at low H2O2 and Fe3+ concentration under visible light
irradiation.
In chapter 4, the effects of temperature on adsorption and Fenton oxidation of
BPA on the Fe-GAC were investigated. First, the adsorption of BPA on the raw GAC
and the Fe-GAC followed the same isotherm model (Freundlich isotherm) and exhibited
similar adsorption capacity. The adsorption rates of BPA onto the Fe-GAC at various
temperatures followed the pseudo second order kinetic and were enhanced with
increasing temperature mainly due to increase in diffusion of BPA. The calculation of
thermodynamic parameters associated with adsorption of BPA suggested that the
adsorption of BPA onto the Fe-GAC was a spontaneous, exothermic and physical
adsorption process. The molar ratio [H2O2]:[BPA] of 36 – 108 in the Fenton oxidation of
the BPA-spent Fe-GAC led to 91 – 99 % removal of BPA with negligible aromatic
products and some soluble organic acids. However, the oxidation rates of BPA and
H2O2 during the Fenton oxidation of the BPA-spent Fe-GAC was drastically enhanced
by the factor of 2.5 and 5 when the reaction temperature increased from 293 – 333 K.
The temperature-dependent relative increase in Fenton oxidation rate of BPA in the Fe-
GAC was found to be similar to that in diffusion of BPA. It suggested that the BPA
oxidation in the Fe-GAC is mainly controlled by diffusive transport of BPA from the Fe-
GAC. Besides, the Thiele-modulus analysis clearly supported more significant pore
diffusion limitation of H2O2 with increasing temperature and the Fe-GAC particle size.
In chapter 5, the novel biochar-TiO2 hybrid catalyst was synthesized for the
photocatalytic oxidation of sulfamethoxazole under UVC light. The characterization of
the catalyst confirmed well-dispersed TiO2 on biochar with little agglomeration. It also
indicated the anatase TiO2 in the biochar-TiO2 hybrid. The experimental results
presented that the catalyst loading, the concentration of nitrate and bicarbonate had
strong influences on the degradation and mineralization of sulfamethoxazole. The
photocatalytic oxidation led to high degradation and mineralization (75% and 86%) at
the optimum conditions (0.1 g catalyst, 10 mg/L, initial pH of 4 and temperature of 20 ±
2 ℃). The mineralization by the photocatalytic was much higher than the photolytic
oxidation. The longer UV reaction time with the biochar-TiO2 leads to higher removal
115
efficiency of COD, but did not affect on the SMX removal. The result of OUR shows the
byproducts formed after UV+biochar-TiO2 treatment can be biodegradable. However,
UV treatment alone leads to generation of more toxic or persistent product. Therefore,
the experimental results clearly indicated that the photocatalytic oxidation process using
the novel biochar-TiO2 hybrid was effective for the degradation of SMX under UVC
irradiation, and this system could be combined with biodegradation process.
In overall, combination of adsorption and (photo) catalytic oxidation or
biodegradation demonstrated high potential for treating wastewater containing emerging
contaminant including endocrine disrupting compound and antibiotics.
116
APPENDIX. Proof of publications
117
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