comparative degradation of emerging pollutants using
TRANSCRIPT
United Arab Emirates UniversityScholarworks@UAEU
Theses Electronic Theses and Dissertations
11-2017
Comparative Degradation of Emerging PollutantsUsing Chemical and Enzymatic ApproachesKhadega Abobaker Abudullarhman Al-Maqdi
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Recommended CitationAbudullarhman Al-Maqdi, Khadega Abobaker, "Comparative Degradation of Emerging Pollutants Using Chemical and EnzymaticApproaches" (2017). Theses. 729.https://scholarworks.uaeu.ac.ae/all_theses/729
vii
Abstract
Organic pollutants, especially those found in water bodies; pose a direct threat to
various aquatic organisms as well as humans. A variety of different remediation
approaches, including chemical and biological methods have been developed for the
degradation of these organic pollutants. However, comparative mechanistic studies
of pollutant degradation by these different systems are almost non-existent. In this
study, the degradation of two emerging pollutants an antibiotic pollutant,
Sulfamethoxazole (SMX) and a model thiazole pollutant, Thioflavin T (ThT), was
carried out using advanced oxidation process (AOP), using UV+H2O2 or a
peroxidase enzyme system. Optimization conditions for Sulfamethoxazole
degradation by peroxidase enzyme showed an absolute requirement for a redox
mediator (Hydroxybenzotriazole) as well as low pH for pollutant degradation. The
other conditions for the efficient degradation were as follows: the concentration of
hydrogen peroxide, peroxidase enzyme and SMX needed were 56 µM, 78 nM and 5
ppm respectively. The degradation of both pollutants was followed using UV-Vis
spectrophotometry and liquid chromatography-mass spectroscopy (LC-MS), and the
products formed were identified using tandem liquid chromatography-mass
spectrometry-mass spectrometry (LC-MS-MS). The results showed that the two
remediation approaches UV+H2O2 and peroxidases produced different sets of
intermediates suggesting that different degradation schemes were operating in the
two systems (AOP vs. Peroxidase enzyme system). Phytotoxicity studies carried out
using Lactuca sativa (lettuce) showed different levels of detoxification of the
pollutants by the two different remediation approaches. This is the first time that a
detailed comparative study showing in detail the intermediates generated in chemical
and biological remediation methods has been presented. Furthermore, the results
show that different remediation systems have very different degradation schemes and
result in products having different toxicities.
Keywords: Bioremediation, advanced oxidation process, peroxidases, enzymes,
chloroperoxidase, soybean peroxidase.
viii
Title and Abstract (in Arabic)
ببستخدام العبهل الكيويبئي و العبهل الانزيويهقبرنة تحلل الولوثبت العضوية النبشئة
صالولخ
انهىثاث انعضىت لاسا حهك انخىاجذة ف انا نها حأثش يباشش عه انحا انبحشت و عه
الاسا اضا. ي أىاع اسانب انعانجت انسخخذيت ف ححهم انهىثاث انعضىت ه اسخخذاو
حائ. و نك انذساساث انكاكت انخ حماس يا ب ححهم هز انعايم انكائ و انعايم الأ
انهىثاث انعضىت بىاسطت هز انظى انخخهفت حكاد حكى يعذويت. ف هز انذساست، حى إجشاء
ححهم نضاد حى سهفايثىكساصول و يهىد ثاصون وهى ثىفلاف ع طشك عهاث
انظشوف الأيثم نخحهم انسهفايثىكساصول ع و انبشوكسذاص. الأكسذة انخمذيت و ظاو الاضى
طشك اضى انبشوكسذاص أ حخطهب وسظ الأكسذة انهذسوكس بضوحشاصول و وسظ
حض.و اضا، كاج بعط انظشوف انثه الأخشي عه انحى انخان حشكض بشوكسذ
ش حشكض اى يىن 87يكشو يىنش و حشكض اضى انبشوكسذاص 0.0.0انهذسوج
انظشوف الأيثم نخحهم ثىفلاف ع طشك اضى وأجضاء نكم يهى. .انسهفايثىكساصول
واضا، كاج بعط حخطهب وسظ الأكسذة انهذسوكس بضوحشاصول.لا هاانبشوكسذاص أ
يىنش و حشكض هي 1 انظشوف انثه الأخشي عه انحى انخان حشكض بشوكسذ انهذسوج
وبعذ رنك حى ححهم أجضاء نكم يهى. .2اى يىنش حشكض ثىفلاف 10ضى انبشوكسذاص ا
. وأظهشث انخائج أ هج LC-MS-MSوال HPLC انعاث وخائجها بسخخذاو حمت ال
ي انىسائظ. يا شش ان ا هج انعانجت عهى بطشق انجت أخجج يجىعاث يخخهفتعان
حى عم دساساث نهخسى انباح عه انهىثاث، باسخخذاو بزوسانخس و لذ أظهشث يخخهفت. و لذ
هز ه انشة الأون انخ حى فها انخائج يسخىاث يخخهفت ي إصانت انسىو نكم ي انهىثاث.
عشض دساست آنت يماست حب بانخفصم انىسطاء انخىنذة ف طشق انعانجت انكائت
وعلاوة عه رنك، حب انخائج أ ظى انعانجت انخخهفت نذها يخططاث ححهم وانبىنىجت.
يخخهفت جذا وحؤد إن يخجاث نها دسجاث يخخهفت ي انست.
.انعانجت انبىنىجت، عهت الأكسذة انخمذيت، انبشوكسذضاث: هفبهين البحث الرئيسية
ix
Acknowledgements
I would first like to thank my thesis advisor Prof. Syed Salman Ashraf. His
door was always open whenever I ran into a trouble or had a question about my
research. He taught me so much about hand work and work ethics and he always
steered me in the right direction whenever he thought I needed it.
I would also like to acknowledge Dr. Mohammed Meetani and Dr.
Mohammed Abu Haija for being in my examination committee and I am grateful for
their very valuable comments on this thesis.
I would also like to thank my friends that I spent these two years with Aysha
Alnyadai, Ghada, Nora, Lamia, Aysha. Also, I want to thank Afra albaloosi for
always giving me encouraging words.
Finally, I must express my very profound gratitude to my parents and family
for providing me with unfailing support and continuous encouragement throughout
my years of study. This accomplishment would not have been possible without them.
x
Dedication
To my beloved Mom, this is for you.
My adoration for you is immeasurable.
xi
Table of Contents
Title ..................................................................................................................................... i
Declaration of Original Work ........................................................................................... ii
Copyright ......................................................................................................................... iii
Advisory Committee ......................................................................................................... iv
Approval of the Master Thesis ........................................................................................... v
Abstract ........................................................................................................................... vii
Title and Abstract (in Arabic) ........................................................................................ viii
Acknowledgements ........................................................................................................... ix
Dedication .......................................................................................................................... x
Table of Contents .............................................................................................................. xi
List of Tables.................................................................................................................. xiii
List of Figures ................................................................................................................. xiv
List of Abbreviations....................................................................................................... xvi
Chapter 1: Introduction ...................................................................................................... 1
1.1 Overview .......................................................................................................... 1
1.2 Removal of emerging pollutants by various methods ................................. 4
1.2.1 Physical methods ...................................................................................... 5
1.2.2 Chemical methods .................................................................................... 6
1.3 Bioremediation/Biodegradation ....................................................................... 9
1.3.1 Enzymatic assisted removal of organic pollutants ................................... 9
1.3.2 Peroxidases used in emerging pollutants degradation ............................ 10
1.4 Comparative studies between chemical and enzymatic approaches
for emerging pollutants degradation ............................................................. 12
1.5 Examples of emerging pollutants degradation pathways ............................... 12
1.5.1 Emerging pollutants degradation pathways using enzymatic
approach ................................................................................................. 12
1.5.2 Emerging pollutants degradation pathways using AOP
approaches .............................................................................................. 18
1.6 Overall Aim of this study ............................................................................... 21
Chapter 2: Materials and Methods ................................................................................... 26
2.1 Reagents ......................................................................................................... 26
2.2 Sulfamethoxazole and Thioflavin T Degradation .......................................... 26
2.3 LC-MS and MS-MS Analyses ....................................................................... 27
2.4 Phytotoxicity Assay ....................................................................................... 28
xii
Chapter 3: Results and Discussion ................................................................................... 29
3.1 Degradation of SMX ...................................................................................... 29
3.1.1 Degradation of SMX by SBP + H2O2 .................................................... 29
3.1.2 Degradation of SMX by UV + H2O2 ...................................................... 40
3.1.3 Degradation pathway of SMX by SBP+ H2O2 and UV + H2O2............. 46
3.1.4 Toxicity test of SMX degraded by SBP+ H2O2 and UV + H2O2 ........... 53
3.2 Degradation of ThT ........................................................................................ 55
3.2.1 Degradation of ThT by UV + H2O2 ....................................................... 55
3.2.2 Degradation of ThT by CPO+H2O2 ....................................................... 57
3.2.3 Analysis of product formation using LC-MS ......................................... 59
3.2.4 Mechanistic studies ................................................................................ 63
3.2.4 Toxicity studies ...................................................................................... 69
Chapter 4: Conclusion ...................................................................................................... 71
References ........................................................................................................................ 72
List of Publications .......................................................................................................... 80
xiii
List of Tables
Table 1: Summary of some emerging pollutants detected in drinking water
supply. ............................................................................................................ 3
Table 2: Summary of AOPs involving the generation of hydroxyl radicals ................ 7
Table 3: Chemical and physical properties of Sulfamethoxazole (SMX).................. 21
Table 4: Chemical and physical properties of Thioflavin T (ThT) ............................ 23
Table 5: Summary of previous studies using biological or chemical methods for
the degradation of SMX ............................................................................... 47
Table 6: Peak area of intermediates produced during the degradation of ThT by
UV+ H2O2 and CPO + H2O2 processes. ...................................................... 62
xiv
List of Figures
Figure 1: Sources and pathways of emerging pollutants to reach drinking water ....... 4
Figure 2: Different approaches used to remove organic pollutants ............................. 5
Figure 3: Possible pathways of degradation of Bisphenol A by a mixture of
ligninolytic enzymes ............................................................................... 13
Figure 4: Possible pathways of degradation of Bisphenol A by laccase enzymes..... 13
Figure 5: Possible pathways of degradation of MBT by CPO ................................... 14
Figure 6: Possible pathways of degradation of MBT by SBP ................................... 15
Figure 7: Possible pathways of degradation of Diclofenac by CPO .......................... 16
Figure 8: Possible pathways of degradation of naproxen by CPO............................. 17
Figure 9: Possible pathways of degradation of FLM by the Electro-Fenton
process ..................................................................................................... 18
Figure 10: Possible pathways of degradation of diethyl phthalate by UV/H2O2 ....... 19
Figure 11: Possible pathways of degradation of IBP by Photo-Fenton reaction ....... 20
Figure 12: Structures of various Sulfonamide emerging pollutants .......................... 22
Figure 13: Structures of some Thiazole-based emerging pollutants .......................... 24
Figure 14: UV/Vis absorbance spectra of SMX degraded by SBP. ........................... 30
Figure 15: SMX degradation by SBP + H2O2 in the the presence and absence of
HOBT. ..................................................................................................... 31
Figure 16: The pH effect for efficient SMX degradation by SBP. ............................ 32
Figure 17: Optimization of SBP concentration needed for SMX degradation. ......... 33
Figure 18: Optimization of SMX concentration needed for degradation by SBP. .... 34
Figure 19: Optimization of H2O2 concentration needed for SMX degradation. ........ 35
Figure 20: Liquid chromatography‐mass spectroscopy (LC-MS) analyses of
SMX degradation by SBP + H2O2 process. ........................................... 37
Figure 21: LC-MS analyses of SMX degraded by SBP + H2O2 process. .................. 38
Figure 22: Tandem mass spectrometry fragmentation analysis of proposed
intermediates produced after SMX degradation by SBP + H2O2 ............ 39
Figure 23: SMX degradation by a UV + H2O2.. ........................................................ 41
Figure 24: Liquid chromatography‐mass spectroscopy (LC-MS) analyses of
SMX degraded by UV + H2O2 process. .................................................. 43
Figure 25 : LC-MS analyses of SMX degraded by UV + H2O2 process. .................. 44
Figure 26: Tandem mass spectrometry fragmentation analysis of proposed m/z
270 intermediates produced after SMX degradation by UV + H2O2. ..... 45
Figure 27: Degradation of SMX by Chloroperoxidase .............................................. 48
Figure 28: Degradation of SMX by Sulfate-reducing bacteria. ................................. 49
Figure 29: Degradation of SMX by Photo-Fenton ..................................................... 50
Figure 30: egradation of SMX using UV + H2O2 .................................................... 51
Figure 31: Proposed degradation pathways of SMX by both SBP and UV .............. 52
Figure 32: Phytotoxicity of studies SMX .................................................................. 54
Figure 33: Thioflavin T dye degradation by UV + H2O2 ........................................... 56
xv
Figure 34: Thioflavin T dye degradation by Chloroperoxidase (CPO)+ H2O2. ......... 58
Figure 35: LC-MS chromatogram and MS/MS analyses of ThT degraded by
UV + H2O2 and CPO + H2O2 processes. ................................................. 61
Figure 36: Summary of the intermediates produced upon ThT degradation by
UV + H2O2 and CPO + H2O2 processes.. ................................................ 62
Figure 37: Tandem mass spectrometry fragmentation analyses of proposed
intermediates produced after ThT degradation by UV+ H2O2. ............... 64
Figure 38: Proposed structures of some of the intermediates generated during
UV+ H2O2 based degradation of ThT dye. ............................................. 67
Figure 39: Proposed structures of some of the intermediates generated during
CPO+ H2O2 based degradation of ThT dye. ........................................... 68
Figure 40: Phytotoxicity of studies ThT .................................................................... 70
xvi
List of Abbreviations
AOPs Advance Oxidative Processes
CPO Chloroperoxidase
EPs Emerging pollutants
FLM Fluometuron
HOBT 1-Hydroxybenzotriazole
IBP Ibuprofen
LiP Lignin Peroxidase
MBT 2-Mercaptobenzothiazole
MnP Manganese Peroxidase
PAEs Phthalate Esters
PAHs Poly Aromatic Hydrocarbons
RM Redox mediator
SBP Soybean peroxidase
SMX Sulfamethoxazole
ThT Thioflavin T
VP Versatile Peroxidase
WWTPs Waste Water Treatment Plants
1
Chapter 1: Introduction
1.1 Overview
It is now well established that a relatively new class of organic pollutants,
named ―emerging pollutants‖ are increasingly being detected in our water supply
(Petrović et al., 2003). Emerging pollutants (EPs) or contaminants of emerging
concerns are a new set of man-made chemicals that have been studied before and
most of them are still not regulated by existing water-quality regulations (Farré et al.,
2008; Gavrilescu et al., 2015). They consist of a wide group of compounds which
include pesticides, pharmaceuticals (e.g. anti-inflammatory drugs, analgesics and
antibiotics), personal care products, surfactants, hormones and sterols etc. These
pollutants are believed to have potential risks to the environmental ecosystems and
human health (Farré et al., 2008). For example, a 2011 study concluded that the
presence of perfluorinated compounds in the serum were related to increase breast
cancer risk in Greenlandic Inuit women (Bonefeld-Jorgensen et al., 2011).
Additionally, it has been reported that the presence of pollutants such as
perfluorooctanoate and perfluorooctane sulfonate in blood may be linked to
decreased human reproductive abilities in women (Fei et al., 2009). Although
emerging pollutants may be degraded by numerous wastewater treatment systems it
is noticed that these emerging pollutants have been detected in relatively high
concentrations in different water bodies (Barnes et al., 2008; Rodil et al., 2012;
Sorensen et al., 2015). For example, a recent study have reported the detection of up
to 10 ppm of acetaminophen in the US water streams (Loos et al., 2007). Not
surprisingly, the appearance of these physiologically active chemicals in high
concentration in the water supply is attracting a lot of attention by environment
2
scientists. A literature review conducted to document physiologically active
concentrations of various hormones, antibiotics, and other EPs in the water bodies in
several countries are summarized in Table 1. What makes EPs bigger threat to the
environment is the fact that they do not need to be persistent in the environment to
cause destructive effects as their high transformation and elimination rates can be
offset by their continuous introduction into the environment. The main source of EPs
in drinking and ground water is their incomplete removal from wastewater treatment
plants (WWTPs). That is due to the fact that WWTPs are not designed to remove
these types of compounds and therefore a large portion of emerging compounds and
their metabolites escape removal in WWTPs and enter the aquatic environment
(Petrović et al., 2003). This is summarized in Figure 1 which shows how EPs can
reach drinking water. One of the challenges that come with dealing with EPs is that
risk assessment and ecotoxicological data for these compounds are not available.
Additionally, the absence of analytical methods for their determination at trace levels
make it difficult to predict their fate in the aquatic environment (Petrović et al.,
2003). However, these challenges are being overcome with more and more
toxicological and analytical research being done on these compounds. For example,
Liquid chromatography with mass spectrometry (LC-MS) and Gas chromatography
with mass spectrometry (GC-MS) have become one of the most sensitive techniques
for detecting EPs compounds at trace concentrations in environmental. (Farré et al.,
2008).
3
Table 1: Summary of some emerging pollutants detected in drinking water supply.
Category Chemical Concentration,
µg/L Reference
Pesticides
Fluometuron 317.6 (Papadakis et al., 2015)
Chlorpyrifos 0.42 (Papadakis et al., 2015)
Prometryn 0.48 (Papadakis et al., 2015)
DEET 6.5 (Lapworth et al., 2012)
Pharmaceutical
Atenolol
0.9 (Huerta-Fontela et al., 2011)
Hydrochlorthiazide
1.9 (Huerta-Fontela et al., 2011)
Phenytoin
0.14 (Huerta-Fontela et al., 2011)
Ibuprofen
12 (Lapworth et al., 2012)
Ketoprofen
2.9 (Lapworth et al., 2012)
Lincomycin 0.73 (Kolpin et al., 2002)
Sulfamethoxazole 1.9 (Kolpin et al., 2002)
Tetracycline 0.71 (Kolpin et al., 2002)
Stimulant Caffeine 6 (Kolpin et al., 2002)
PAHs (poly
aromatic
hydrocarbons)
Pyrene 0.84 (Kolpin et al., 2002)
fluoranthene 1.2 (Kolpin et al., 2002)
Phenanthrene 0.53 (Kolpin et al., 2002)
plasticizer
bis(2-ethylhexyl)
adipate 10 (Kolpin et al., 2002)
bis(2-ethylhexyl)
phthalate 20 (Kolpin et al., 2002)
Bisphenol A 12 (Kolpin et al., 2002)
Nonionic
detergent
metabolite
4-nonylphenol 40 (Kolpin et al., 2002)
4
1.2 Removal of emerging pollutants by various methods
A wide range of approaches have been developed for the removal of
emerging pollutants from wastewaters, as well as water bodies. Figure 2
summarized the various physical, chemical, and biological approaches that
can be used to deal with these contaminants.
Hospitals
wastewater
Households
wastewaterAgriculture run off
Pharmaceuticals & personal
care products
Industry
wastewater
Animal farming pharmaceuticals
effluents
WWTP
Water works
Drinking water
Groundwater Rivers and
Oceans
Ineffective
treatment
Figure 1: Sources and pathways of emerging pollutants to reach drinking water
5
1.2.1 Physical methods
There are several physical methods that have been developed for the removal
of organic pollutants. For example, adsorption, filtration osmosis and coagulation
have all been successfully used to remove organic pollutants. Adsorption using
activated carbon, peat, wood chips and silica gel all show potential for remediation of
organic pollutants. The use of adsorption techniques has been increasing lately due to
their effectiveness in the elimination of organic pollutants. One major advantage of
adsorption process is that they are economically reasonable. The most frequently
used adsorbent for organic pollutants removal by adsorption is activated carbon. The
efficiency of this technique depends on what type of carbon and the amount being
used. Wood chips, rice husks and other natural sorbents are considered good
alternatives but they are not as good as other available sorbents such as activated
carbon (Liu et al., 2009; Robinson et al., 2001).
Various groups have published studies using newly developed membrane
filtration systems for the removal of organic pollutants. For membrane filtration
processes there are different types that can be used, which are microfiltration (MF),
Treatment Methods for Organic Pollutants
Chemical Methods Biological MethodsPhysical Methods
Electrolysis
UV Photolysis
OzonationCoagulation
Adsorption
Filtration
Osmosis
Enzymes
Bacteria
Figure 2: Different approaches used to remove organic pollutants
6
and nanofiltration (NF). Microfiltration and nanofiltration are used to eliminate
microparticles and macromolecules. Some drawbacks to membrane filtration are
high capital cost and membrane replacement (Van der Bruggen et al., 2005). For
reverse osmosis membranes it has shown retention rate of 90% for organic aromatic
pollutants and other chemical compounds (Ciardelli and Ranieri, 2001).
Sedimentation is known for its selectivity towards the category of contaminants
existing in wastewater. One disadvantage of this method is the high sludge
production (Robinson et al., 2001).
1.2.2 Chemical methods
Some of the most commonly used chemical methods for the degradation of
organic pollutants include ozonation, NaOCl treatment, and advanced oxidation
processes (AOPs). The most commonly used chemical method for water treatment is
advanced oxidation processes. The most common feature in all AOPs is the in situ
generation of hydroxyl radicals (HO•) which rapidly and efficiently react with
organic pollutants to degrade them (Bokare and Choi, 2014). Advanced oxidation
processes have been effectively used to degrade various types of organic
contaminants in simulated as well as real wastewater samples. There are numerous
review papers that discuss how advanced oxidation processes have a positive impact
in water cleanup (Comninellis et al., 2008; Poyatos et al., 2010; Robinson et al.,
2001). Table 2 shows summary of the various different AOPs all involving the
generation of hydroxyl radicals (Bokare and Choi, 2014). In spite of their usefulness,
the use of AOPs has some disadvantage like the fact that it can be very expensive
due to the continuous addition of costly chemicals and energy (Gupta and Suhas,
2009).
7
Table 2: Summary of AOPs involving the generation of hydroxyl radicals
Type of
AOP
Brief description of overall
process and operating
conditions
Typical examples and key reaction
pathways
UV
Photolysis
(direct &
indirect)
UV photons (150–400 nm)
irradiate aqueous solutions with
high fluences (500–
4000 mJ/cm2). Homogeneous or
heterogeneous chemical species
(or materials) present in water
then absorb the incident UV
light and undergo photo-
transformation reactions to
generate free hydroxyl radicals.
UV/Water photolysis
H2O + hν (<190 nm) → H• + HO
•
UV/ozone
O3 + H2O + hν(254 nm) → 2HO• + O2
UV/H2O2
H2O2 + hν (254 nm) → 2HO•
UV/Ozone/H2O2
2O3 + H2O2 + hν(254 nm) → 2HO• + 3O
2
UV/Fenton (Photo-Fenton)
Fe3+
+ H2O + hν(>300 nm) → Fe2+
+ H
O• + H
+
H2O2 + hν → 2HO•
TiO2 photocatalysis
TiO2 + hν (>300 nm) → e−
cb + h+
vb
2e−
cb + O2 + 2H+ → H2O2
e−
cb + H2O2 → HO• + OH
−
h+
vb + OH−
ads → HO•
Ozonation Using high voltage alternating
current (6–20 kV) across a
dielectric discharge gap
containing pure oxygen or dry
air, O2 molecules are dissociated
into oxygen atoms to form
ozone (O3). This ozone gas
stream is bubbled into water for
decomposition into free
hydroxyl radicals.
Ozone/alkali
O3 + HO2− → O3
•− + HO2
•
O3•−
+ H+ → HO
• + O2
Ozone/H2O2
2O3 + H2O2 → 2HO• + 3O2
Ozone/Iron(II)
Fe2+
+ O3 → FeO2+
+ O2
FeO2+
+ H2O → Fe3+
+ HO• + HO
−
Ozone/TiO2Photocatalyst
TiO2 + hν (> 300 nm) → e−
cb + h+
vb
e−
cb + O3 → O3− → O2 + O
−
O− + H2O → HO
• + OH
−
h+
vb + OH− → HO
•
8
Type of
AOP
Brief description of overall
process and operating
conditions
Typical examples and key reaction
pathways
Electro-
chemical
treatment
Using indirect and/or direct
anodic reactions, electrical
energy vector drives the in
situ formation of oxidants
(HO• or H2O2 or both) and their
reaction with organic species in
water.
Anodic Oxidation
H2O → H+ + HO
•
Electro-Fenton
O2(g) + 2H+ + 2e
− → H2O2(cathode)
Fe2+
+ H2O2 → Fe3+
+ HO• + OH
−
Ultrasound
(US)
irradiation
High frequency sound waves
(20 kHz–1 MHz) generate
transient or stable acoustic
cavitations in water solution.
The cavitation involves
formation, growth and collapse
of pressure bubbles with
inducing thermal dissociation of
water into hydroxyl radicals.
US/Ozone
H2O
+))) → H• + HO
• O3+))) → O2(g) + O(
3P)
O(3P)(g) + H2O → 2HO
•
US/Hydrogen Peroxide
H2O2 +))) → 2HO•
Non-
thermal
plasma
An electric discharge in water
initiates collisions of charged
particles and excitation–
ionization of neutral molecules.
This generates a non-thermal
plasma containing energetic
electrons (e−*) and radicals.
Electrical discharge → e−*
H2O + e−* → HO
• + H
• + e
−
Table 2: Summary of AOPs involving the generation of hydroxyl radicals (continued)
9
1.3 Bioremediation/Biodegradation
Although physical and chemical methods enjoy wide-scale applicability and
are currently used in various large-scale processes, they still face some significant
limitations and challenges. The biggest downside of physico-chemical methods are
the high costs involved as well as the sludge produced by the processes. Greener
alternatives, such as the use of micro-organisms or phytoremediation, are considered
to be more environmentally friendly and have shown promising results for the
removal of low concentrations of organic pollutants. Compared with traditional
physico-chemical methods, bioremediation may be a safer, less disruptive, and more
cost-effective treatment strategy. However, a fundamental shortcoming of
bioremediation is that the organisms used for this purpose may be unable to thrive in
adverse and unfavorable environmental conditions, as well as the potential
inefficiency of the process (Boopathy, 2000; Rauf and Salman Ashraf, 2012).
1.3.1 Enzymatic assisted removal of organic pollutants
Enzymatic bioremediation is an emerging method of supplementing bio-
treatment techniques. The class of enzyme that is most commonly used for
bioremediation purposes is the ―oxidoreductase‖ class of enzymes, which includes
oxygenases, monooxygenases, dioxygenases, laccases, and peroxidases. These
enzymes carry out redox reactions on a relatively wide range of substrates, including
polycyclic aromatic hydrocarbons (PAHs), polynitrated aromatic compounds,
pesticides, bleach-plant effluents, synthetic dyes, polymers, and various emerging
pollutants (Ali et al., 2013; Bibi et al., 2011; Franciscon et al., 2010; Kalsoom et al.,
2013). Within the oxidoreductase family of enzymes, peroxidases such as
ligninperoxidase (LiP), manganese-dependent peroxidase (MnP), versatile
10
peroxidase (VP), soybean peroxidase (SBP), and chloroperoxidase (CPO) have been
extensively studied due to their high potential for the degradation of various organic
compounds (CHENG et al., 2007; Kalsoom et al., 2013; Ryan et al., 2006). It has
been postulated that these enzymes could oxidize organic compounds through the
generation of reactive free radicals, leading to the formation of lower molecular
weight compounds and eventually mineralizing the pollutant. It is worth mentioning
that some of the potential shortcomings of using enzymes for large-scale
bioremediation purposes are their relatively high costs as well as the lack of the
reusability of the enzymes and their stabilities during the remediation processes.
However, most of these issues can be ameliorated by employing molecular biology
approaches for the microbial recombinant expression of wild-type as well as mutated
forms of these enzymes, which can then be immobilized on various supports, such as
beads or membranes, to increase their stabilities as well as reusability (Adrio and
Demain, 2014).
1.3.2 Peroxidases used in emerging pollutants degradation
1.3.2.1 Soybean peroxidase (SBP)
The seed coat of the soybean plant is the source of the peroxidase called
soybean peroxidase (SBP) (Gijzen et al., 1993). This enzyme works by oxidizing
different substances by using hydrogen peroxide as a co-substrate. It is well known
that heme peroxidases such as SBP can react with H2O2 to generate an enzymatic
iron oxyradical called Compound I, which can react with organic substrates to be
converted to the Compound II form of the enzyme and an organic radical. The
Compound II form of the enzyme can react with another organic substrate to create
11
another organic radical molecule and regenerate the resting form of the enzyme
(Rauf and Salman Ashraf, 2012), as shown below:
Peroxidase + H2O2 → Compound I + H2O (1)
Compound I + SH → Compound II + S• (2)
Compound II + SH → Peroxidase + S• + H2O (3)
where SH indicates a generic substrate.
There are many studies that show the ability of this enzyme to degraded
different classes of molecules and its capability to clean water. One study shows that
SBP was able to remove 96% of phenol from water in only 20 min (Flock et al.,
1999). Another study shows how SBP is able to efficiently degrade an emerging
pollutant 2-mercaptobenzothiazole (MBT) that is found in water bodies (Alneyadi
and Ashraf, 2016).
1.3.2.2 Chloroperoxidase (CPO)
The fungus Caldariomyces fumago secretes heme containing peroxidase
called Chloroperoxidase (CPO), also known as chloride: hydrogen- peroxide
oxidoreductase. In addition to catalyzing halogenation reactions, CPO also exhibits
peroxidase, catalase, and cytochrome P450-like activities. Although CPO shares
common features with other heme-containing peroxidases as shown in equations 1 to
3, its structure is unique, being composed of a tertiary assembly consisting primarily
of eight helical segments. This enzyme have been used for the efficient and rapid
degradation of various organic pollutants (Alneyadi et al., 2017; Alneyadi and
Ashraf, 2016; Hofrichter et al., 2010; Liu et al., 2014; Sundaramoorthy et al., 1995)
12
1.4 Comparative studies between chemical and enzymatic approaches for
emerging pollutants degradation
Although a tremendous amount of research has been published on the use of
AOPs (Abdelraheem et al., 2016; Cesaro and Belgiorno, 2016; Oturan and Aaron,
2014) and enzymatic (Alneyadi et al., 2017; Alneyadi and Ashraf, 2016; Dhillon and
Kaur, 2016; Rodríguez-Delgado and Ornelas-Soto, 2017) approaches to the
degradation of organic pollutants, including some ―combination approaches‖ (Calza
et al., 2014; García-Montaño et al., 2008a, 2008b; Nogueira et al., 2015; Sánchez
Peréz et al., 2014), to date there have been few detailed studies published that
compare the degradation of a specific organic compound by these two different
methods. Many questions remain about the pollutant degradation pathways, the types
of intermediates generated, and the residual toxicity of the pollutants when using
these two very different methods. The ultraviolet (UV) + H2O2 AOP approach relies
on the pollutant molecules reacting with the hydroxyl radicals generated upon the
photolysis of H2O2, whereas the SPB and CPO enzyme method relies on the enzyme
heme-iron oxyradical. Therefore, it is expected that there may be significant
differences in the intermediates produced during the two processes.
1.5 Examples of emerging pollutants degradation pathways
1.5.1 Emerging pollutants degradation pathways using enzymatic approach
Bisphenol A is commonly used in industrial applications such as the synthesis
of polymers including polycarbonates, polyesters and polyacrylates. Bisphenol A has
been documented as an Endocrine Disrupting Chemicals, consequently, it is essential
to measure its biodegradability or fate in the environment (Chhaya and Gupte, 2013).
A detailed study by Gassara and colleagues shows the complete degradation of
13
Bisphenol A by a mixture of ligninolytic enzymes (MnP, LiP and laccase) as shown
in Figure 3. Another study of degradation of Bisphenol A using laccase enzymes is
shown in Figure 4.
Figure 3: Possible pathways of degradation of Bisphenol A by a mixture of
ligninolytic enzymes (MnP, LiP and laccase)
(Gassara et al., 2013)
(Mohapatra et al., 2010)
2-Mercaptobenzothiazole (MBT) a bicyclic heteroatomic molecule is
extensively used in the production of tires, rubber shoes and other rubber items as a
Figure 4: Possible pathways of degradation of Bisphenol A by laccase enzymes
14
rubber vulcanizer. Consequently; MBT has been detected in the effluents of
industrial WWTP and also in different aquatic environment (Li et al., 2008). A
recent study by Alneyadi and Ashraf shows the degradation pathways of MBT using
two different enzymes CPO and SBP as shown in Figure 5 and 6 respectively.
(Alneyadi and Ashraf, 2016)
Figure 5: Possible pathways of degradation of MBT by CPO
15
(Alneyadi and Ashraf, 2016)
Diclofenac and Naproxen are anti-inflammatory drugs, which are commonly
used for the treatment of arthritis, ankylosing spondylitis, and acute muscle pain. It
has been found that these drugs are tough to degrade by general waste treatment
approaches and have caused serious environmental concerns (Li et al., 2017). A 2017
study has shown the possible pathways for the degradation of Diclofenac and
Naproxen using CPO enzyme as shown in Figures 7 and 8.
Figure 6: Possible pathways of degradation of MBT by SBP
16
(Li et al., 2017)
Figure 7: Possible pathways of degradation of Diclofenac by CPO
17
(Li et al., 2017)
Figure 8: Possible pathways of degradation of naproxen by CPO
18
1.5.2 Emerging pollutants degradation pathways using AOP approaches
Fluometuron (FLM) is a herbicide that is extensively used in agriculture. It is
a phenylurea derivative, and therfore highly soluble in water. They have been
reported to present risks for human health due to their toxicity and unfortunately they
have been found in various groundwater sources (Diaw et al., 2017). A recent study
by Diaw and colleagues reported that Electro-Fenton approach could be used to
completely mineralize this pollutant. Figure 8 shows proposed mineralization
mechanism of FLM in aqueous solution by the Electro-Fenton process.
(Diaw et al., 2017)
Figure 9: Possible pathways of degradation of FLM in aqueous solution by
the Electro-Fenton process
19
Phthalate esters (PAEs) are mainly used as plasticizers for plastics and
therefore, it is produced by many countries around the world in high quantities. It
was found that these compounds can bioaccumulate in animal fat. Furthermore, it has
been established that PAEs compounds can act as endocrine disruptors and lead to
negative effects in both human and animals even in low concentrations. For humans,
they lead to diseases like disorders of male reproductive tract, breast and testicular
cancers. In recent years, PAEs have been widely detected in aquatic environment in
concentrations range between ng/L and μg/L (Xu et al., 2007). One example of PAE
compounds is diethyl phthalate. A 2007 study showed that diethyl phthalate can be
completely degraded by UV/H2O2. Figure 10 shows the mechanism of degradation of
diethyl phthalate by UV/H2O2.
(Xu et al., 2007)
Figure 10: Possible pathways of degradation of diethyl phthalate by UV/H2O2
20
Ibuprofen (IBP) is a widely used anti-inflammatory drug which has been
found in effluents of wastewater treatment plants (WWTPs). Studies show that its
concentration in the environment can range between 10 ng/L and 169 μg/L (Méndez-
Arriaga et al., 2010). A detailed study by Méndez-Arriaga and colleagues has shown
the possible pathways for the degradation of IBP by Photo-Fenton reaction as shown
in Figure 11.
(Méndez-Arriaga et al., 2010)
Figure 11: Possible pathways of degradation of IBP by Photo-Fenton reaction
21
1.6 Overall Aim of this study
As mentioned earlier that detailed comparative studies for the degradation of
a given pollutant by two different remediation methods are few. Therefore, the
present work was carried out to compare the degradation of two different pollutants
Sulfamethoxazole and Thioflavin T by two different approaches. Enzymatic
approach using peroxidases enzymes and UV/H2O2 approach.
Sulfamethoxazole is a sulfonamide compound and it shares the sulfonamide
core with different emerging pollutants, like Furosemide (drug used to treat fluid
build-up due to heart failure), Celebrex (non-steroidal anti-inflammatory drug), and
Glyburide (diabetes drug), as shown in Figure 13. Both Sulfamethoxazole and
Furosemide have been detected in different aquatic environments in concentrations
1.9 µg/L and 22 ng/L respectively (Huerta-Fontela et al., 2011; Kolpin et al., 2002).
Table 3: Chemical and physical properties of Sulfamethoxazole (SMX)
Compound Class Molecular
formula
Formula
weight λ max Structure
Sulfameth-
oxazole Sulfonamide C10H11N3O3S 253.27 265nm
22
Figure 12: Structures of various Sulfonamide emerging pollutants
23
The other compound chosen for the present study was the thiazole compound
Thioflavin T. Thioflavin T is a useful model for pollution remediation studies, as its
thiazole core is a common feature in some of the important emerging pollutants, such
as 2-mercaptobenzothiazole (a plasticizer/vulcanizing agent), Thiabendazole and
Tricyclazole (fungicidal drugs), and Meloxicam, (a non-steroidal anti-inflammatory
drug) as shown in Figure 13. Alarmingly, all of these and other thiazole derivatives
have been detected in various aquatic environments (Clarke and Smith, 2011;
Collado et al., 2014).
Table 4: Chemical and physical properties of Thioflavin T (ThT)
Compound Class Molecular
formula
Formula
weight λ max Structure
Thioflavin T Thiazole C17H19ClN2S 318.86 412nm
24
Figure 13: Structures of some Thiazole-based emerging pollutants
25
The main objectives of the current work are summarized below:
1. Optimize the conditions for enzymatic degradation of Sulfamethoxazole and
Thioflavin T such as requirement of redox mediator, pH, concentration of the
enzyme (SBP) and H2O2.
2. Identify the intermediates produced in the degradation of both pollutants,
Sulfamethoxazole by SBP and UV/H2O2, Thioflavin T by CPO and
UV/H2O2.
3. Propose a degradation pathway for the intermediates produced in the
enzymatic and UV/H2O2 degradation of both pollutants, Sulfamethoxazole
and Thioflavin T.
4. Carry out a comparative degradation analysis using two different approaches
enzymatic approach using SBP or CPO and UV/H2O2 approach to degrade
two different pollutants Sulfamethoxazole and Thioflavin T.
5. Carry out phytotoxicity studies using Lactuca sativa seeds to compare the
toxicity of enzymatically + UV/H2O2 degraded Sulfamethoxazole and
Thioflavin T.
26
Chapter 2: Materials and Methods
2.1 Reagents
Sulfamethoxazole and Thioflavin T were purchased from AnaSpec (Fremont,
CA, USA). Hydrogen peroxide (30% w/v) and LC-MS grade solvents, such as
formic acid and acetonitrile, were purchased from Sigma-Aldrich (St. Louis, MO,
USA). Experiments were carried out using universal buffers with 0.1 M citrate acid
and 0.2 M K2HPO4. CPO with a specific activity of 1296 IU/mg (17 mg/mL, 405
µM) and SBP with a specific activity of 2700 IU/mg (1 mg/mL, 26 µM) were
purchased from Bio-Research Products (North Liberty, IA, USA).
2.2 Sulfamethoxazole and Thioflavin T Degradation
For the photolytic treatment of SMX and ThT using UV + H2O2, 0.056 mM
H2O2 was added to SMX samples in phosphate citrate universal buffer, (pH 4), 1 mM
H2O2 was added to ThT samples in citrate buffer (pH 2), these samples were then
irradiated with a UV lamp (UVGL-58, J-129, Upland, NJ, USA) from a distance of
1.5 cm. The instrument had a UV power output of 6 W and was selectively used in
the 254 nm output mode for these studies. Under these conditions, no significant
warming up of the irradiated solution was observed.
For the enzymatic treatment of SMX using SBP + H2O2 experiments were
carried out to find the optimum conditions for the reaction, which include the
requirement of redox mediator (HOBT), pH optimization and the optimum
concentration of SMX, H2O2 and SBP. The degradation experiments were carried out
as follow SMX and H2O2 (0.056 mM) in phosphate citrate universal buffer (pH 4)
was exposed to SBP (78 nM), and the changes in the SMX spectrum were monitored.
27
For the enzymatic treatment of ThT using CPO + H2O2, experiments were carried out
as follow ThT and H2O2 (1 mM) in citrate buffer (pH 2) was exposed to CPO (10
nM), and the changes in the full ThT spectrum were monitored.
In both the enzymatic and the photolytic studies, spectra were collected in the
range of 200–800nm using a Carry 60 Spectrophotometer (Agilent Technologies,
Santa Clara, CA, USA), with a path length of 1 cm (4 mL quartz cuvette) and 265
nm and 412 nm was used as λmax for Sulfamethoxazole and Thioflavin T
respectively.
2.3 LC-MS and MS-MS Analyses
The materials produced after the enzymatic and the photolytic treatments
were analyzed using LC-MS. SMX and ThT samples were filtered using a 0.45 µm
CA syringe filter prior to injection. The LC-MS was fitted with a ZORBAX Eclipse
Plus C18 column (Agilent Technologies, Santa Clara, CA, USA) with a particle size
of 1.8 µm, an inner diameter of 2.1 mm, and a length of 50 mm. The column was
maintained at 35 ◦C, and a constant flow rate of 0.2 mL/min was maintained. The
column was coupled to a 6420 Triple Quad LC-MS System detector (Agilent
Technologies). Two mobile phases were used: A is water containing 0.1% formic
acid, and B is 100% acetonitrile. The LC method was set as follows: 5 min of 100%
A, followed by a 0–100% gradient of B from 5–20 min, then 5 min of 100% B after
the gradient, and finally 5 min of 100% A. For the MRM method it was set as
follows: 1 min of 100% A followed by a 0–100% gradient of B from 1–4 min, and
finally 3 min of 100% A. The electrospray ionization source in the LC-MS system
was in positive polarity mode, the capillary voltage was set at 4000 V, the nebulizer
pressure was maintained at 45 psi, the drying gas (N2) flow was 11 L/min, and the
28
drying temperature was set at 325◦C. The mass range monitored for all of the runs
was between 50 and 1000 Da. In the tandem MS experiments using the product ion
mode, nitrogen gas was used for fragmentation and different collision energies were
used.
2.4 Phytotoxicity Assay
The toxicity of SMX and ThT before and after UV+H2O2 and the enzymatic
treatments was measured using the lettuce seed growth inhibition assay, similar to a
previously described method but with slight modifications (Alneyadi et al., 2017).
Briefly, twenty L. sativa seeds were placed on sterilized Whatman filter paper, No. 3,
in a Petri dish and saturated with 4 mL of the samples. The petri dishes were
incubated for 5 days in a humidified chamber at 25±2 °C. Distilled water was used as
a negative control, SMX and ThT was used as a positive control, with each sample
being tested in duplicate. The effects of the original SMX and ThT, the SMX and
ThT samples degraded by UV + H2O2 and the enzymes (SBP or CPO) were
examined by measuring the lengths of the roots of the germinated seeds. Statistical
analyses were conducted for each group of treated seeds.
29
Chapter 3: Results and Discussion
3.1 Degradation of SMX
3.1.1 Degradation of SMX by SBP + H2O2
3.1.1.1 Degradation optimization of SMX by SBP + H2O2
3.1.1.1.1 The requirement of redox mediator (RM)
It is well known that some peroxidase enzymes work better in degrading
organic compounds in the present of redox mediators (Husain and Husain, 2007).
Initial experiments using SBP + H2O2 alone showed almost no degradation of SMX
Figure 14. However, addition of the redox mediator 1-hydroxybenzotriazole (HOBT)
to the reaction mixture resulted in a new peak around λmax 260 nm as shown if Figure
14. This new peak was not due to HOBT (data not shown) and suggested new
compounds being generated. To prove that the increase in the absorbance at 260 nm
was the result of the degradation of SMX and the formation of new compounds LC-
MS-MS analyses were carried out. We used the highly sensitive and specific MRM
mode in LC-MS-MS to analyze SMX alone, SMX + SBP + H2O2 and SMX + SBP +
H2O2 + HOBT. As shown in Figure 15 A the addition of HOBT resulted in dramatic
and complete degradation of SMX. Figure 15 B shows a time course for this and
within 10 minutes 80% of SMX was degraded in the presences of SBP + H2O2 +
HOBT.
30
*
Figure 14: UV/Vis absorbance spectra of SMX degraded by SBP in the
presence and absence of HOBT; asterisk (*) indicate the shift in λmax =
265 nm and the formation of new products. [SMX] = 5 ppm, pH = 4,
[H2O2] = 0.056 mM.
31
Figure 15: SMX degradation by SBP + H2O2 in the presence and absence of HOBT.
(A) MRM scans of SMX degradation by SBP + H2O2 in the presence and absence of
HOBT. (B) Time course of SMX degradation by SBP + H2O2 + HOBT. [SMX] = 5
ppm, [H2O2] = 0.056 mM, [SBP] = 78 nM, pH 4.
A
B
MRM (254 -> 156) SMX alone
MRM (254 -> 156) SMX +SBP+ H2O2
MRM (254 -> 156) SMX +SBP+ H2O2 + HOBT
32
3.1.1.1.2 pH optimization
It is well known that enzymatic reactions are pH dependent; therefore,
experiments were done to determine the optimum pH for the degradation of SMX by
SBP. We carried out degradation experiments at different pH values ranging from
pH 2 to pH 7, while maintaining all other parameters constant. Figure 16 shows that
the degradation of SMX by SBP was pH dependent and that SBP enzyme worked
best for degrading SMX at low pH values. The result showed that from pH 2 to pH 6
the degradation of SMX was almost 100%, while at pH 7 the degradation of SMX
decreased to below 40%. This is consistent with previous studies that showed SBP to
be most active at low pH values (Kalsoom et al., 2013; Marchis et al., 2011) . Based
on these results, pH 4 was chosen as the pH for all subsequent optimization studies.
Figure 16: The pH effect for efficient SMX degradation by SBP. [SMX] = 5 ppm,
[H2O2] = 0.056 mM, [SBP] = 78 nM.
33
3.1.1.1.3 SBP concentration optimization
Another important factor that can affect the degradation of SMX is enzyme
(SBP) concentration. Due to the high cost of enzymes it is very vital to know the
lowest enzyme concentration that can lead to the best degradation of SMX. The
concentrations of SBP tested in the present work ranged from 0.62 nM to 78 nM. The
results are summarized in Figure 17, which showed that having low concentrations
(0.62 nM to 15.6 nM) of SBP could not efficiently degrade SMX. However, 31.2 nM
SBP could degrade about 50% of SMX and almost 90% of SMX was degraded when
SBP concentration was 78 nM. Based on this data, 78 nM SBP was chosen for all
subsequent studies.
Figure 17: Optimization of SBP concentration needed for efficient SMX degradation.
[SMX] = 5 ppm, pH = 4, [H2O2] = 0.056 mM.
34
3.1.1.1.4 SMX concentration optimization
The concentration of SMX present in the reaction mixture can also effect
SBP mediated degradation of this pollutant. Therefore, we tested the SBP + H2O2 +
HOBT system at concentrations of SMX (5, 25, and 50 ppm), while keeping all the
other parameters constant. As shown in Figure 18 the best SMX degradation was
observed at the lowest SMX concentration (5 ppm). Additional increase in SMX
concentration to 25 and 50 ppm resulted in slower degradation to about 45% and
30% respectively. Based on this result, 5 ppm SMX was chosen for all other
reactions. In a separate study we found that 500 ppm SMX could also be degraded
(78%) under the same conditions (SBP concentration 78 nM, H2O2 concentration
0.056 mM, HOBT concentration 0.05 mM) if the reaction was allowed to proceed for
3 hours (data not shown).
Figure 18: Optimization of SMX concentration needed for degradation by SBP. pH =
4, [H2O2] = 0.056 mM, [SBP] = 78 nM.
35
3.1.1.1.5 H2O2 concentration optimization
Peroxidase enzymes use H2O2 as a co-reactant, therefore the initial
concentration of H2O2 in a peroxidase mediated reaction is very important. If the
H2O2 concentration is too low the enzyme activity becomes low, on the other hand,
very high H2O2 concentration can oxidize the enzyme and cause its permanent
inactivation. Therefore, studies were carried out at different H2O2 concentration
while keeping all other parameters constant. As seen from Figure 19 the optimum
H2O2 concentration for the degradation of SMX was 0.056 mM. It was also seen that
SBP enzyme was stable at high H2O2 concentration of up to 18.36 mM.
Figure 19: Optimization of H2O2 concentration needed for efficient SMX
degradation by SBP. [SMX] = 5 ppm, pH = 4, [SBP] = 78nM.
36
3.1.1.2 HPLC and LCMS analysis for SMX degraded by SBP + H2O2
As discussed previously, the UV/Vis spectroscopic data for degradation of
SMX by SBP + H2O2 (Figure 14) showed an increase in a peak with λmax = 260 nm,
therefore suggesting the formation of new products. This was confirmed using LC-
MS and MS-MS analyses. Figure 20 shows the total ion chromatogram (TIC) LC-
MS of nest SMX and SMX treated with SBP + H2O2. As can be seen from the
Figure, the LC-MS analysis of SMX alone showed a single major peak with the
retention time of 15.61 min, and m/z = 254. The LC-MS (TIC) of the treated sample
(5 ppm SMX) showed a dramatic decrease in the SMX peak (m/z 254) and the
appearance of a few new peaks at retention time of 13.73 min and 15.23 min. Since
the new peaks appeared to be very small, the degradation experiment was repeated
with a 20 times higher concentration of SMX (100 ppm) which better showed the
newly generated intermediates peaks (Figure 21 A). Figure 21 B shows the LC-MS
spectrum for the three new peaks which are m/z 99, m/z 120 and m/z 375 with
retention time 12.59, 13.73 and 15.23 respectively. We further carried out tandem
mass spectrometry-mass spectrometry (MS-MS) analyses to come up with proposed
structures for m/z 99 and m/z 120 two intermediates generated in treatment of SMX
by SBP + H2O2 as in Figure 22.
37
Figure 20: Liquid chromatography‐mass spectroscopy (LC-MS) analyses of SMX
degradation by SBP + H2O2 process; asterisk (*) indicate the formation of new
products. The conditions for SBP + H2O2 degradation were: [SMX] = 5 ppm, pH = 4,
[H2O2] = 0.056 mM.
10 12 14 16 18 20
500000
1000000
1500000
10 12 14 16 18 20
500000
1000000
1500000
Time (min)
SMX alone
SMX + SBP + H2O2
10 12 14 16 18 20
500000
1000000
1500000
Time (min)
Rela
tive I
nte
nsit
y
* *
38
10 12 14 16 18 200
1000000
2000000
3000000
4000000
Time (min)
Rela
tive I
nte
nsit
y
SMX + SBP + H2O2
* **
Figure 21: LC-MS analyses of SMX degraded by SBP + H2O2 process. (A) Total ion
chromatogram of 100 ppm SMX degraded by SBP + H2O2. (B) Extracted ion
chromatograms for intermediates with m/z 99, 120 and 375. The conditions for SBP
+ H2O2 degradation were: [SMX] = 100 ppm, pH = 4, [H2O2] = 0.056 mM.
10 12 14 16 18 20
0
20000
40000
60000
Time (min)
Rela
tive I
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nsit
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10 12 14 16 18 20
0
100000
200000
300000
Time (min)
Rela
tive I
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10 12 14 16 18 20
0
20000
40000
60000
Time (min)
Rela
tive I
nte
nsit
y
(EIC) m/z 99 (EIC) m/z 120
(EIC) m/z 375
A
B
39
Figure 22: Tandem mass spectrometry fragmentation analysis of proposed
intermediates produced after SMX degradation by SBP + H2O2 (m/z of 99, panels A)
and (m/z of 120, panels B).
60 80 100 120
5000
10000
15000
20000
m/z
Rela
tive I
nte
nsit
y
40 60 80 100
200
400
600
800
1000
m/z
Rela
tive I
nte
nsit
y
99
44 72
H2N
ON
CH3
44
72
120
91
91
S
O
OH
65
65
A
B
H2N
ON
CH3
44
72
91
S
O
OH
65
40
3.1.2 Degradation of SMX by UV + H2O2
In the present study, we were also interested in the H2O2 -assisted
photochemical oxidation of SMX. Figure 23 A shows the UV/Vis spectra of SMX
before and after UV + H2O2 treatment. As can be seen from the Figure, the exposure
of the SMX to UV + H2O2 caused a gradual decrease in the intensity of λmax 265 nm,
indicating the degradation of SMX. Figure 23 B shows the decrease of λmax 265 nm
as a function of time. The degradation of SMX upon exposure to UV + H2O2 was
most likely caused by the hydroxyl radicals produced from H2O2 when exposed to
UV radiation as indicated below:
H2O2 + → 2 OH• (4)
SMX + OH• → degraded products (5)
41
SMX
SMX + UV + H2O2
Figure 23: SMX degradation by a UV + H2O2. (A) UV/Vis absorbance spectra for
SMX degradation by UV + H2O2. (B) Changes in absorbance at 265 nm of SMX
after treatment with UV + H2O2 as a function of time. [SMX] = 5 ppm, pH = 4,
[H2O2] = 0.056 mM. UV-Vis scans were taken every 5 min.
A
B
42
3.1.2.1 HPLC and LCMS analysis for SMX degraded by UV + H2O2
Based on the UV/Vis spectroscopic data for degradation of SMX by UV +
H2O2 (Figure 23) which showed a decrease in SMX λmax = 265 nm, we wanted to
confirm the degradation of SMX and the formation of new products using liquid
chromatography-mass spectroscopy (LC-MS). Figure 24 shows the LC-MS
chromatograms of SMX, as well as SMX (5 ppm) treated with UV + H2O2. As can
be seen from the Figure, exposure of UV + H2O2 to SMX caused a complete
degradation of the SMX peak m/z 254 at retention time of 15.61 min. To be able to
observe possible intermediates the same study was done but with higher SMX
concentration (100 ppm) as seen in Figure 25 A; which showed two new product
peak (indicted with asterisk (*)). Figure 25 B shows the extracted ion chromatograms
LC-MS spectrum for the two new peaks which are m/z 245 and m/z 270 with
retention time 13.15 and 15.20 respectively. Using tandem mass spectrometry-mass
spectrometry (MS-MS) we were able to propose structure for m/z 270 intermediate as
can be seen in Figure 26.
43
Figure 24: Liquid chromatography‐mass spectroscopy (LC-MS) analyses of SMX
degraded by UV + H2O2 process. The conditions for UV + H2O2 degradation were:
[SMX] = 5 ppm, pH = 4, [H2O2] = 0.056 mM.
10 12 14 16 18 20
500000
1000000
1500000
Time (min)
10 12 14 16 18 20
500000
1000000
1500000
SMX alone
SMX + UV + H2O2
10 12 14 16 18 20
500000
1000000
1500000
Time (min)
Rela
tive I
nte
nsit
y
44
Figure 25 : LC-MS analyses of SMX degraded by UV + H2O2 process. (A) Total ion
chromatogram of 100 ppm SMX degraded by UV + H2O2. (B) Extracted ion
chromatogram for intermediates m/z 254 and 270. The conditions for UV + H2O2
degradation were: [SMX] = 100 ppm, pH = 4, [H2O2] = 0.056 mM.
B
45
Figure 26: Tandem mass spectrometry fragmentation analysis of proposed m/z 270
intermediates produced after SMX degradation by UV + H2O2
0 100 200 300
200
400
600
800
1000
m/z
Rela
tive I
nte
nsit
y
270
71
8657
H2N
S
O O
NH
ON
CH3HO
71
57
86H2N
S
O O
NH
ON
CH3HO
71
57
86
46
3.1.3 Degradation pathway of SMX by SBP+ H2O2 and UV + H2O2
One of the aims of our present work was also to compare the degradation
pathways of SMX using the two different remediation processes SBP + H2O2 and
UV + H2O2. We were able to identify two of three of the intermediates generated
during the enzymatic degradation of SMX, namely m/z 99, and m/z 120. Similarly,
we obtained chemical structure for one of the two new intermediates generated
during the UV + H2O2 degradation of SMX, namely m/z 270 and m/z 254. For m/z
254 we were not able to confirm the structure using tandem mass spectrometry
because this compound has the same mass as the original SMX but elutes at a
different retention time. However, we were still able to propose a structure for this
new compound based on rearrangement of SMX. Figure 27 summarize these
findings and shows the structures of the intermediates generated during the two
different remediation approaches. The fact that we don’t have any common
intermediates between the two different remediation processes; suggests that
different degradation pathways are employed in the enzymatic and AOP degradation
of SMX.
Our results presented above and summarized in Figure 31 are consistent with
previously published degradation studies of SMX. Table 5 summarizes these studies
and shows that biological (peroxidases-mediated or microbial-assisted) or chemical
(AOP-based approaches) remediation of SMX produced different sets of compounds,
with only two common intermediates (m/z 99 and m/z 288). Additionally, three of the
intermediates detected in our studies (m/z 99, m/z 254 and m/z 270) have also been
detected by others (Table 5). Furthermore, proposed degradation schemes from these
studies are shown in Figures 27, 28, 29 and 30. It is interesting to note that the SMX
47
intermediates observed in our present study are slightly different than those
published earlier, which subsequently led us to propose a slightly different
degradation scheme by SBP+H2O2 or UV+H2O2 (Figure 31).
Table 5: Summary of previous studies using biological or chemical methods for the
degradation of SMX
Biological methods Reference
Versatile
Peroxidase m/z 99, 158, 245 and 503 (Eibes et al., 2011)
Chloroperoxidase m/z 145 and 287 (Zhang et al., 2016)
Sulfate-reducing
bacteria m/z 173, 216, 254 and 288 (Jia et al., 2017)
Chemical (AOP) methods Reference
UV + H2O
2 m/z 99, 115, 133 and 190 (Lekkerkerker-
Teunissen et al., 2012)
Photo-Fenton m/z 92, 108, 156 and 276 (Trovó et al., 2009)
TiO2
Photocatalysis m/z 99, 270, 288 (Hu et al., 2007)
48
Figure 27: Degradation of SMX by Chloroperoxidase (Zhang et al., 2016).
49
Figure 28: Degradation of SMX by Sulfate-reducing bacteria (Jia et al., 2017).
50
Figure 29: Degradation of SMX by Photo-Fenton (Trovó et al., 2009).
51
Figure 30: Degradation of SMX using UV + H2O2 (Lekkerkerker-Teunissen et al.,
2012).
52
Figure 31: Proposed degradation pathways of SMX by SBP + H2O2 and UV + H2O2
SBP + H2O2Sulfamethoxazole
[M+H]+, m/z 254RT = 15.61
UV + H2O2
H2N
ON
CH3
[M+H]+, m/z 99RT = 15.59
[M+H]+, m/z 120RT = 13.73
[M+H]+, m/z 375RT = 15.23
[M+H]+, m/z 254RT = 13.15
[M+H]+, m/z 270RT = 15.20
OrNH2
S
O O
NH N
OCH3
SNH
ON
CH3
O
OH2N
NH2
S
O O
NH
ON
CH3
S
O
OH
H2N
S
O O
NH
ON
CH3HO
53
3.1.4 Toxicity test of SMX degraded by SBP+ H2O2 and UV + H2O2
The fact that the two remediation methods generated different sets of
intermediates from SMX prompted us to compare the residual phytotoxicities of the
two remediated solutions. The phytotoxicity studies were carried out using Lactuca
sativa seeds by exposing them to SMX before and after treatment with the two
different remediation processes. Figure 32 shows the results of the toxicity studies, in
which the root lengths of L. sativa seeds exposed to distilled water (control), neat
SMX, SMX treated with SBP + H2O2, or SMX treated with UV + H2O2 were
measured. As can be seen from the figure, neat SMX had a significant phytotoxic
effect on the seeds, causing a dramatic and significant decrease in the mean root
length. Significant inhibition of root lengths was also observed in the SMX + SBP +
H2O2 and showed no significant difference as compared to SMX alone (p < 0.05).
However, a significant improvement in mean root length was observed when SMX
was treated with UV + H2O2. Statistical analyses using t-test showed that the SMX +
UV + H2O2 sample was significantly (p < 0.01) less toxic than neat SMX sample.
This result suggests that the UV+ H2O2 treatment of SMX almost completely
eliminated the toxicity of SMX.
54
Control SMX
SMX + UV+ H2O2SMX + SBP + H2O2
Figure 32: Phytotoxicity of SMX, SMX treated by SBP + H2O2 and SMX treated by
UV + H2O2 on L. sativa seeds, as measured by the mean root lengths (cm). Statistical
analyses were performed using unpaired t-test (n = 40), asterisk (*) denotes a
significant difference (p < 0.01). [SMX] = 500 ppm, pH = 4, [H2O2] = 0.056 mM.
55
3.2 Degradation of ThT
3.2.1 Degradation of ThT by UV + H2O2
In the present work, we initially investigated the H2O2-assisted
photochemical oxidation of ThT. Figure 33 shows the chemical structure as well as
the absorption spectrum of ThT. Also shown is the major peak in the yellow visible
region of the ThT absorption spectrum (λmax = 412 nm). The exposure of the ThT
solution to UV + H2O2 caused an immediate and gradual decrease in the intensity of
λmax, indicating that new compounds had formed. Figure 29 B shows the degradation
of λmax as a function of time, and it can be seen that about 70% of the compound had
degraded after 60 min of this AOP treatment. No ThT degradation was observed in
the presence of UV light or H2O2 alone. The degradation of ThT was attributed to the
hydroxyl radicals (OH•) produced from H2O2 when exposed to UV radiation.
Hydroxyl radicals are known to be strong oxidizing agents that can react with ThT
molecules to produce intermediates that are responsible for decoloring/degradation of
the original solution (Georgiou et al., 2002; Konstantinou and Albanis, 2004). A
simplified reaction scheme for this process is outlined below:
H2O2 + → 2 OH• (4)
ThT + OH• → degraded products (5)
The results obtained here are consistent with numerous studies by our group
and others, which show that UV + H2O2 can readily degrade an array of organic
compounds (Rauf et al., 2016; Zhang et al., 2013).
56
Thioflavin T(ThT) Dye
l = 412nm
250 300 350 400 450 5000.0
0.5
1.0
1.5
2.0
2.5
Wavelength (nm)
Ab
so
rban
ce
Figure 33: Thioflavin T (ThT) dye degradation by UV + H2O2 advanced oxidation
process. (A) UV/Vis absorbance spectra for ThT degradation by UV + H2O2. [ThT
dye] = 25 ppm, pH = 2, [H2O2]= 1mM. The UV-Vis scans were taken every 5 min.
(B) Percentage of ThT remaining (decrease in Absorbance at 412 nm) after treatment
with UV + H2O2. [ThT dye] = 25 ppm, pH = 2, [H2O2] = 1mM.
A
B
57
3.2.2 Degradation of ThT by CPO+H2O2
In order to compare the UV/ H2O2-induced degradation of ThT with an
enzymatic approach, we used the well-known peroxidase CPO to degrade ThT.
Figure 34 shows the absorbance spectra of ThT when exposed to CPO +
H2O2 as a function of time. As was seen for the UV + H2O2 degradation of ThT in
Figure 29, the ThT started to degrade immediately upon exposure to CPO + H2O2, as
evidenced by the decrease in the absorbance at 412 nm. Interestingly, as the peak at
412 nm decreased in intensity a new peak with λmax = 350 nm appeared, with the
intensity of this peak increasing over time. This is shown more clearly in Fig. 30 B.
The increase in the absorbance this peak suggests that a new compound was being
generated upon the treatment of ThT with CPO + H2O2, something that was not seen
in the case of UV + H2O2 treatment of ThT. This observation suggests that the UV +
H2O2 and CPO + H2O2 processes likely have different effects on ThT.
58
l = 350nm
l = 412nm
Figure 34: Thioflavin T (ThT) dye degradation by Chloroperoxidase (CPO)+ H2O2.
(A) UV/Vis absorbance spectra for ThT dye degradation by CPO + H2O2. [ThT dye]
= 25 ppm, pH = 2, [H2O2]= 1 mM, [CPO] = 10 nM. The UV-Vis scans were taken
every 5 min. (B) Changes in absorbances at 412 nm and 350 nm of ThT dye after
treatment with CPO + H2O2. [ThT dye] = 25 ppm, pH = 2, [H2O2] = 1 mM, [CPO] =
10 nM.
A
B
59
3.2.3 Analysis of product formation using LC-MS
Since the UV-Vis spectroscopic data for degradation of ThT by UV+ H2O2
and CPO + H2O2 suggested that different degradative pathways might be operating in
the two remediation methods, we wanted to confirm this using LC-MS. Figure 35
shows the LC-MS chromatograms of ThT, ThT exposed to UV + H2O2 , and ThT
treated with CPO + H2O2 . Additionally, the insets show the mass spectra of the
major peaks in the three samples. LC-MS analysis of the neat ThT dye shows a
single major peak at a retention time of 16.11 min, with m/z = 283 (inset). This
molecular weight is in agreement with the loss of the chloride counter ion from the
molecular weight of the ThT dye (MW = 318.86 Da), which would be expected as
the LC-MS system was operating in the positive mode. LC-MS analysis of the ThT +
UV + H2O2 sample shows a much smaller ThT peak (retention time = 16.11 min),
which is consistent with the degradation of ThT. However, another major peak can
also be seen at a retention time of 15.64 min. The mass spectrum of this new product
peak (inset) shows it to have an m/z value of 269. In addition to this major product
peak, several other smaller peaks can also be seen at 14–16 min range. LC-MS
analysis of ThT dye treated with CPO + H2O2 shows a very different LC profile, with
the major peak having a retention time of 16.19 min. The mass spectrum of this peak
shows it to have an m/z value of 317, indicating the presence of a compound that is
very different from both the original ThT (m/z = 283) and the major intermediate
produced during the UV + H2O2 AOP degradation of ThT (m/z = 269). In addition to
this m/z = 317 peak, additional minor peaks can also be seen in the ThT + CPO +
H2O2 sample.
60
Detailed analyses of all the peaks detected after the UV + H2O2 and CPO +
H2O2 treatments of ThT are shown in Table 6. As can be seen from this table, AOP-
mediated degradation of ThT (ThT + UV + H2O2) generated four intermediates with
m/z values of 297, 288, 269, and 255, with the m/z = 269 peak being the most
prominent intermediate (based on peak area). Interestingly, the intermediates
produced during the enzymatic treatment of ThT (ThT + CPO + H2O2) were very
different, with m/z values of 397, 361, 351, 321, 317, 303, 289, and 269. It is worth
noticing that except for the m/z = 269 species, the two different treatments produced
completely different intermediates. This interesting and novel finding is also
graphically represented in Figure 36.
61
Figure 35: LC-MS chromatogram and MS/MS analyses of ThT degraded by UV +
H2O2 and CPO + H2O2 processes. The conditions for UV + H2O2 degradation were:
[ThT] = 25 ppm, pH = 2, [H2O2] = 1mM, and for CPO + H2O2 degradation for: [ThT
dye] = 25 ppm, pH = 2, [H2O2] = 1 mM, [CPO] = 10 nM.
10 12 14 16 18 20
0
1000000
2000000
3000000
4000000
5000000
Rela
tive I
nte
nsit
y
10 12 14 16 18 20100
10500000
101000000
101500000
Rela
tive I
nte
nsit
y
10 12 14 16 18 20
0
100000
200000
300000
400000
500000
10 12 14 16 18 20
0
1000000
2000000
3000000
4000000
5000000
Rela
tive I
nte
nsit
y
ThT
ThT + CPO + H2O2
ThT + UV + H2O2
0 100 200 300 400
50000
100000
150000
200000
m/z
Rela
tive I
nte
nsit
y
0 100 200 300 400
50000
100000
150000
200000
m/z
Rela
tive I
nte
nsit
y
0 100 200 300 4000
100000
200000
300000
m/z
Rela
tive I
nte
nsit
y
0 100 200 300 400
50000
100000
150000
200000
250000
m/z
Rela
tive I
nte
nsit
y
283
269 283
317
62
Figure 36: Summary of the intermediates produced upon ThT degradation by UV +
H2O2 and CPO + H2O2 processes. The asterisk (*) indicate intermediates whose
structures are shown in Figures 34 and 35.
Table 6: Peak area of intermediates produced during the degradation of ThT by UV+
H2O2 and CPO + H2O2 processes.
269*
CPO + H2O2
255*
288
297
UV + H2O2
289* 303*
317* 321*
351* 361
397
63
3.2.4 Mechanistic studies
An attempt was made to propose plausible structures for the various different
intermediates generated in the two different treatments by using tandem MS-MS
data. Indeed, as can be seen in Figure 37, we were able to propose structures for
seven of the intermediates produced, with the relevant peaks being indicated by
asterisks (*) in Figure 36. Based on the structure of the intermediates, we were able
to develop plausible schemes for ThT breakdown during the UV + H2O2 (Figure 38)
and CPO + H2O2 (Figure 39) remediation processes. In the UV + H2O2 process, the
main mechanism involves the formation of OH• radicals by the UV-mediated
homolysis of hydrogen peroxide. These reactive OH• radicals attack ThT, leading to
stepwise demethylation of the tertiary amine and resulting in the formation of
intermediates with m/z values of 269. Subsequent demethylation produces the m/z =
255 compound. This demethylation mechanism has been previously reported by our
group as well as others (He et al., 2011; Hisaindee et al., 2013; Meetani et al., 2010).
The decrease in the retention times of these intermediates (15.64 min and 15.8 min)
on the reverse-phase column when compared to ThT (16.11 min) reflects their
increased polarities. Other, more polar intermediates with m/z values of 297 and 288
and retention times of 15.11 and 14.35 min, respectively, were also produced during
the UV + H2O2 treatment of ThT.
64
Figure 37: Tandem mass spectrometry fragmentation analyses of proposed
intermediates produced after ThT degradation by UV+ H2O2 (m/z of 269, panel A) or
CPO + H2O2 as shown in panels A, B (m/z of 289), C (m/z of 321) and D (m/z of
351).
A B
C D
150 200 250 300
5000
10000
15000
m/z
Rela
tive I
nte
nsit
y
226269
253
S
NH
N
253226
253
200 220 240 260 280 300
1000
2000
3000
m/z
Rela
tive I
nte
nsit
y
289254
274
S
NH2
N
Cl254274
200 250 300 350
2000
4000
6000
8000
10000
m/z
Rela
tive I
nte
nsit
y
267 277290
321
305
S
NH
N
ClO305
267
290
277
200 250 300 350 400
1000
2000
3000
m/z
Rela
tive I
nte
nsit
y
351
315307
300272
335S
N
N
Cl
Cl
307272
300
315
65
Unlike the UV-induced photolysis of H2O2, which generates reactive
hydroxyl radicals, the CPO + H2O2 system entails a different mechanism. It is well
known that heme peroxidases such as CPO can react with H2O2 to generate an
enzymatic iron oxyradical called Compound I, which can react with organic
substrates to be converted to the Compound II form of the enzyme and an organic
radical. The Compound II form of the enzyme can react with another organic
substrate to create another organic radical molecule and regenerate the resting form
of the enzyme (Rauf and Salman Ashraf, 2012), as shown below:
Peroxidase + H2O2 → Compound I + H2O (1)
Compound I + SH → Compound II + S• (2)
Compound II + SH → Peroxidase + S• + H2O (3)
where SH indicates a generic substrate.
Although the above peroxidase reaction cycle does not explicitly show the
generation of hydroxyl radicals, it is possible that OH radicals may also be produced
in this case as CPO is known to cleave the peroxide O-O bond through a glutamic
acid residue present in its active site (Siegbahn and Blomberg, 2001).
Besides the well-known H2O2-peroxidase cycle described above, CPO is also
known to catalyze chlorination of organic compounds (Alneyadi and Ashraf, 2016;
Zhang et al., 2016), as shown below:
R-H + Cl− + H2O2 + H
+ → R-Cl + 2 H2O (6)
66
In fact, based on the MS-MS data, our proposed degradation scheme suggests
that the action of CPO on ThT in the presence of H2O2 can occur through two
overlapping pathways, with the first being the chlorination of ThT, and the second
being the production of OH radicals that can react with ThT and its chlorinated
products (Figure 39). A comparison of the abundance of the proposed species (Table
6) as determined by LC-MS-MS suggests that the chlorination pathway (317 m/z
species) was at least twenty times more active than the demethylation pathway (269
m/z species). A second chlorination also occurs, yielding a compound with an m/z
value of 351 (Figure 37 D). The mono-chlorinated ThT could also undergo stepwise
demethylation, yielding structures with m/z values of 303 and 289. An intermediate
with m/z = 321 was also observed, which corresponds to the addition of an OH
(oxidation) to the demethylated form of chlorinated-ThT, as seen in Figure 39.
The generation of chlorinated products of ThT observed during the CPO +
H2O2 treatment of ThT was not completely unexpected, as ThT contains a chloride
counter ion and it is well established that CPO, in the presence of halide ions can
lead to halogenation of various organic substrates (Hager et al., 1966). In fact, we
have previously published that in contrast to ThT, a different but related thiazole
compound (2-mercaptobenzothiazole) which did not contain a chloride counterion,
did not produce any chlorinated products (Alneyadi and Ashraf, 2016). However, it
is expected that most industrial waste-streams would have relatively high
concentrations of various ions, including halides (Correia et al., 1994) and hence, the
results presented here could be valuable and relevant in real life remediation
situations.
67
Figure 38: Proposed structures of some of the intermediates generated during UV+
H2O2 based degradation of ThT dye.
[M+H]+ , m/z 269
RT =15.64
[M+H]+ , m/z 255
RT =15.86
Demethylation
Demethylation
[M+H]+ , m/z 297
RT =15.11
[M+H]+ , m/z 288
RT =14.35
[M+H]+ , m/z 283
RT =16.11
Thioflavin T
S
N
N
S
NH
N
S
NH2
N
68
Figure 39: Proposed structures of some of the intermediates generated during CPO+
H2O2 based degradation of ThT dye.
[M+H]+ , m/z 283
RT =16.11
[M+H]+ , m/z 303
RT =15.97
[M+H]+ , m/z 317
RT =16. 19
[M+H]+ , m/z 351
RT =17. 11
[M+H]+ , m/z 269
RT =15.64
[M+H]+ , m/z 289
RT =15.39
[M+H]+ , m/z 321
RT =16.21
Chlorination
Chlorination
Demethylation
Demethylation
Chlorination
Demethylation+ Oxidation
Thioflavin T[M+H]+ , m/z 397
RT =17.27
[M+H]+ , m/z 361
RT =16.38
S
N
N
S
NH
N
S
N
N
Cl
S
N
N
Cl
Cl
S
NH
N
Cl
S
NH2
N
Cl
S
NH
N
ClO
69
3.2.4 Toxicity studies
The toxicity of degradation products should be analyzed where possible, as
they can often be more toxic than the parent compound (Silva et al., 2012).
Therefore, we carried out phytotoxicity studies using Lactuca sativa seeds by
exposing them to AOP-remediated ThT and enzymatically treated ThT solutions.
Germination of Lactuca sativa L. var. Buttercrunch seeds is a standard protocol for
assessing toxicity in water and soil matrices, and is recommended for bioassays by
the U.S. Environmental Protection Agency, the Food and Drug Administration, and
the Organization for Economic Cooperation and Development. Figure 40 shows the
results of the toxicity studies, in which the root lengths of L. sativa seeds exposed to
distilled water (control), neat ThT, ThT treated with UV + H2O2, or ThT treated with
CPO + H2O2 were measured. As can be seen from the figure, ThT had a significant
phytotoxic effect on the seeds, causing a dramatic and significant decrease in the
mean root length. Significant inhibition of root lengths were also observed in the
ThT + UV + H2O2 and ThT + CPO + H2O2 samples (Figure 40). However, t-test
analysis showed that the ThT + CPO + H2O2 sample was significantly (p < 0.05) less
toxic than the ThT + UV + H2O2 sample. This result was quite unexpected, and
suggests that one or more of the intermediates generated during the UV + H2O2
treatment of ThT could be toxic. Alternatively, the fact that the ThT + CPO + H2O2
sample still exhibited significant phytotoxicity could be explained by the fact that
this sample still contained large amounts of the toxic undegraded ThT dye. Further
studies will need to be carried out to allow the nature of the phytotoxicity of the ThT
+ UV + H2O2 sample to be understood; however, the present data shows that the UV
70
+ H2O2 and CPO + H2O2 treatments of ThT produce different intermediates that
could have differing toxicities – an observation that has not been published earlier.
Figure 40: Phytotoxicity of ThT, ThT treated by CPO + H2O2 and ThT treated by UV
+ H2O2 on L. sativa seeds, as measured by the mean root lengths (cm). Statistical
analyses were performed using unpaired t-test (n = 40), asterisk (*) denotes a
significant difference (p < 0.05). [ThT] = 25 ppm, pH = 2, [H2O2] = 1 mM.
Contr
ol
ThT 2O2
ThT + U
V +
H
2O2
ThT + C
PO +
H
4
5
6
7
8
p < 0.05
*
Mean
ro
ot
len
gth
(c
m)
71
Chapter 4: Conclusion
In summary, this is one of the few studies to present a systematic comparison
of the use of two different remediation methods (chemical and enzymatic) to treat
two toxic organic pollutants belonging to different classes of emerging
pollutantsThioflavin T (ThT) and Sulfamethoxazole.
(SMX) with UV+ H2O2 and Peroxidase + H2O2 produced very different
degradation products. This suggested that different organic pollutant degradation
schemes were involved in these two remediation approaches. Additionally, we
showed that the AOP (UV+ H2O2) and enzymatic (Peroxidase + H2O2) treated
pollutant solutions (ThT & SMX) had significantly different toxicities for L. sativa
seeds.
Interestingly, peroxidase (Chloroperoxidase + H2O2) treatment caused a small
but significant decrease in the phytotoxicity of ThT, which was not observed in the
case of AOP treatment. This was very different than the case with SMX, where SBP
+ H2O2 + HOBT (peroxidase) treatment did not significantly reduce any of the
toxicity of SMX, but the AOP treatment caused a dramatic elimination of all of the
phytotoxicity of SMX.
The unexpected and intriguing data presented here also highlights the need
for better understandings of different remediation approaches and for additional
research into such comparative remediation studies.
72
References
Abdelraheem, W.H.M., He, X., Komy, Z.R., Ismail, N.M., Dionysiou, D.D., 2016.
Revealing the mechanism, pathways and kinetics of UV254nm/H2O2-based
degradation of model active sunscreen ingredient PBSA. Chem. Eng. J. 288,
824–833. doi:10.1016/j.cej.2015.12.046
Adrio, J.L., Demain, A.L., 2014. Microbial Enzymes: Tools for Biotechnological
Processes. Biomolecules 4, 117–139. doi:10.3390/biom4010117
Ali, L., Algaithi, R., Habib, H.M., Souka, U., Rauf, M.A., Ashraf, S.S., 2013.
Soybean peroxidase-mediated degradation of an azo dye– a detailed
mechanistic study. BMC Biochem. 14, 35. doi:10.1186/1471-2091-14-35
Al-Maqdi, K.A., Hisaindee, S.M., Rauf, M.A., Ashraf, S.S., 2017. Comparative
Degradation of a Thiazole Pollutant by an Advanced Oxidation Process and
an Enzymatic Approach. Biomolecules 7. doi:10.3390/biom7030064
Alneyadi, A.H., Ashraf, S.S., 2016. Differential enzymatic degradation of thiazole
pollutants by two different peroxidases – A comparative study. Chem. Eng. J.
303, 529–538. doi:10.1016/j.cej.2016.06.017
Alneyadi, A.H., Shah, I., AbuQamar, S.F., Ashraf, S.S., 2017. Differential
Degradation and Detoxification of an Aromatic Pollutant by Two Different
Peroxidases. Biomolecules 7, 31. doi:10.3390/biom7010031
Barnes, K.K., Kolpin, D.W., Furlong, E.T., Zaugg, S.D., Meyer, M.T., Barber, L.B.,
2008. A national reconnaissance of pharmaceuticals and other organic
wastewater contaminants in the United States — I) Groundwater. Sci. Total
Environ. 402, 192–200. doi:10.1016/j.scitotenv.2008.04.028
Bibi, I., Bhatti, H.N., Asgher, M., 2011. Comparative study of natural and synthetic
phenolic compounds as efficient laccase mediators for the transformation of
cationic dye. Biochem. Eng. J. 56, 225–231. doi:10.1016/j.bej.2011.07.002
Bokare, A.D., Choi, W., 2014. Review of iron-free Fenton-like systems for activating
H2O2 in advanced oxidation processes. J. Hazard. Mater. 275, 121–135.
doi:10.1016/j.jhazmat.2014.04.054
Bonefeld-Jorgensen, E.C., Long, M., Bossi, R., Ayotte, P., Asmund, G., Krüger, T.,
Ghisari, M., Mulvad, G., Kern, P., Nzulumiki, P., Dewailly, E., 2011.
Perfluorinated compounds are related to breast cancer risk in greenlandic
inuit: A case control study. Environ. Health 10, 88. doi:10.1186/1476-069X-
10-88
73
Boopathy, R., 2000. Factors limiting bioremediation technologies. Bioresour.
Technol. 74, 63–67. doi:10.1016/S0960-8524(99)00144-3
Calza, P., Avetta, P., Rubulotta, G., Sangermano, M., Laurenti, E., 2014. TiO2-
soybean peroxidase composite materials as a new photocatalytic system.
Chem. Eng. J. 239, 87–92. doi:10.1016/j.cej.2013.10.098
Cesaro, A., Belgiorno, V., 2016. Removal of Endocrine Disruptors from Urban
Wastewater by Advanced Oxidation Processes (AOPs): A Review. Open
Biotechnol. J. 10. doi:10.2174/1874070701610010151
CHENG, X.-B., JIA, R., LI, P.-S., ZHU, Q., TU, S.-Q., TANG, W.-Z., 2007. Studies
on the Properties and Co-immobilization of Manganese Peroxidase. Chin. J.
Biotechnol. 23, 90–96. doi:10.1016/S1872-2075(07)60006-5
Chhaya, U., Gupte, A., 2013. Possible role of laccase from Fusarium incarnatum UC-
14 in bioremediation of Bisphenol A using reverse micelles system. J.
Hazard. Mater. 254, 149–156. doi:10.1016/j.jhazmat.2013.03.054
Ciardelli, G., Ranieri, N., 2001. The treatment and reuse of wastewater in the textile
industry by means of ozonation and electroflocculation. Water Res. 35, 567–
572. doi:10.1016/S0043-1354(00)00286-4
Clarke, B.O., Smith, S.R., 2011. Review of ―emerging‖ organic contaminants in
biosolids and assessment of international research priorities for the
agricultural use of biosolids. Environ. Int. 37, 226–247.
doi:10.1016/j.envint.2010.06.004
Collado, N., Rodriguez-Mozaz, S., Gros, M., Rubirola, A., 2014. Pharmaceuticals
occurrence in a WWTP with significant industrial contribution and its input
into the river system. Environ. Pollut. 185, 202–212.
doi:10.1016/j.envpol.2013.10.040
Comninellis, C., Kapalka, A., Malato, S., Parsons, S.A., Poulios, I., Mantzavinos, D.,
2008. Advanced oxidation processes for water treatment: advances and trends
for R&D. J. Chem. Technol. Biotechnol. 83, 769–776. doi:10.1002/jctb.1873
Correia, V.M., Stephenson, T., Judd, S.J., 1994. Characterisation of textile
wastewaters ‐ a review. Environ. Technol. 15, 917–929.
doi:10.1080/09593339409385500
Dhillon, G.S., Kaur, S., 2016. Agro-Industrial Wastes as Feedstock for Enzyme
Production: Apply and Exploit the Emerging and Valuable Use Options of
Waste Biomass. Academic Press.
74
Diaw, P.A., Oturan, N., Seye, M.D.G., Coly, A., Tine, A., Aaron, J.-J., Oturan, M.A.,
2017. Oxidative degradation and mineralization of the phenylurea herbicide
fluometuron in aqueous media by the electro-Fenton process. Sep. Purif.
Technol. 186, 197–206. doi:10.1016/j.seppur.2017.06.005
Eibes, G., Debernardi, G., Feijoo, G., Moreira, M.T., Lema, J.M., 2011. Oxidation of
pharmaceutically active compounds by a ligninolytic fungal peroxidase.
Biodegradation 22, 539–550. doi:10.1007/s10532-010-9426-0
Farré, M. la, Pérez, S., Kantiani, L., Barceló, D., 2008. Fate and toxicity of emerging
pollutants, their metabolites and transformation products in the aquatic
environment. TrAC Trends Anal. Chem., Advanced MS Analysis of
Metabolites and Degradation Products - II 27, 991–1007.
doi:10.1016/j.trac.2008.09.010
Fei, C., McLaughlin, J.K., Lipworth, L., Olsen, J., 2009. Maternal levels of
perfluorinated chemicals and subfecundity. Hum. Reprod. 24, 1200–1205.
doi:10.1093/humrep/den490
Flock, C., Bassi, A., Gijzen, M., 1999. Removal of aqueous phenol and 2-
chlorophenol with purified soybean peroxidase and raw soybean hulls. J.
Chem. Technol. Biotechnol. 74, 303–309. doi:10.1002/(SICI)1097-
4660(199904)74:4<303::AID-JCTB38>3.0.CO;2-B
Franciscon, E., Piubeli, F., Fantinatti-Garboggini, F., Ragagnin de Menezes, C.,
Serrano Silva, I., Cavaco-Paulo, A., Grossman, M.J., Durrant, L.R., 2010.
Polymerization study of the aromatic amines generated by the biodegradation
of azo dyes using the laccase enzyme. Enzyme Microb. Technol. 46, 360–
365. doi:10.1016/j.enzmictec.2009.12.014
García-Montaño, J., Domènech, X., García-Hortal, J.A., Torrades, F., Peral, J.,
2008a. The testing of several biological and chemical coupled treatments for
Cibacron Red FN-R azo dye removal. J. Hazard. Mater. 154, 484–490.
doi:10.1016/j.jhazmat.2007.10.050
García-Montaño, J., Pérez-Estrada, L., Oller, I., Maldonado, M.I., Torrades, F., Peral,
J., 2008b. Pilot plant scale reactive dyes degradation by solar photo-Fenton
and biological processes. J. Photochem. Photobiol. Chem. 195, 205–214.
doi:10.1016/j.jphotochem.2007.10.004
Gassara, F., Brar, S.K., Verma, M., Tyagi, R.D., 2013. Bisphenol A degradation in
water by ligninolytic enzymes. Chemosphere 92, 1356–1360.
doi:10.1016/j.chemosphere.2013.02.071
75
Gavrilescu, M., Demnerová, K., Aamand, J., Agathos, S., Fava, F., 2015. Emerging
pollutants in the environment: present and future challenges in biomonitoring,
ecological risks and bioremediation. New Biotechnol. 32, 147–156.
doi:10.1016/j.nbt.2014.01.001
Georgiou, D., Melidis, P., Aivasidis, A., Gimouhopoulos, K., 2002. Degradation of
azo-reactive dyes by ultraviolet radiation in the presence of hydrogen
peroxide. Dyes Pigments 52, 69–78. doi:10.1016/S0143-7208(01)00078-X
Gijzen, M., Huystee, R. van, Buzzell, R.I., 1993. Soybean Seed Coat Peroxidase (A
Comparison of High-Activity and Low-Activity Genotypes). Plant Physiol.
103, 1061–1066. doi:10.1104/pp.103.4.1061
Gupta, V.K., Suhas, 2009. Application of low-cost adsorbents for dye removal – A
review. J. Environ. Manage. 90, 2313–2342.
doi:10.1016/j.jenvman.2008.11.017
Hager, L.P., Morris, D.R., Brown, F.S., Eberwein, H., 1966. Chloroperoxidase II.
UTILIZATION OF HALOGEN ANIONS. J. Biol. Chem. 241, 1769–1777.
He, Y., Grieser, F., Ashokkumar, M., 2011. The mechanism of sonophotocatalytic
degradation of methyl orange and its products in aqueous solutions. Ultrason.
Sonochem. 18, 974–980. doi:10.1016/j.ultsonch.2011.03.017
Hisaindee, S., Meetani, M.A., Rauf, M.A., 2013. Application of LC-MS to the
analysis of advanced oxidation process (AOP) degradation of dye products
and reaction mechanisms. TrAC Trends Anal. Chem. 49, 31–44.
doi:10.1016/j.trac.2013.03.011
Hofrichter, M., Ullrich, R., Pecyna, M.J., Liers, C., Lundell, T., 2010. New and
classic families of secreted fungal heme peroxidases. Appl. Microbiol.
Biotechnol. Heidelb. 87, 871–97.
doi:http://dx.doi.org.ezproxy.uaeu.ac.ae/10.1007/s00253-010-2633-0
Hu, L., Flanders, P.M., Miller, P.L., Strathmann, T.J., 2007. Oxidation of
sulfamethoxazole and related antimicrobial agents by TiO2 photocatalysis.
Water Res. 41, 2612–2626. doi:10.1016/j.watres.2007.02.026
Huerta-Fontela, M., Galceran, M.T., Ventura, F., 2011. Occurrence and removal of
pharmaceuticals and hormones through drinking water treatment. Water Res.
45, 1432–1442. doi:10.1016/j.watres.2010.10.036
Husain, M., Husain, Q., 2007. Applications of Redox Mediators in the Treatment of
Organic Pollutants by Using Oxidoreductive Enzymes: A Review. Crit. Rev.
Environ. Sci. Technol. 38, 1–42. doi:10.1080/10643380701501213
76
Jia, Y., Khanal, S.K., Zhang, H., Chen, G.-H., Lu, H., 2017. Sulfamethoxazole
degradation in anaerobic sulfate-reducing bacteria sludge system. Water Res.
119, 12–20. doi:10.1016/j.watres.2017.04.040
Kalsoom, U., Ashraf, S.S., Meetani, M.A., Rauf, M.A., Bhatti, H.N., 2013.
Mechanistic study of a diazo dye degradation by Soybean Peroxidase. Chem.
Cent. J. 7, 93. doi:10.1186/1752-153X-7-93
Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber,
L.B., Buxton, H.T., 2002. Pharmaceuticals, Hormones, and Other Organic
Wastewater Contaminants in U.S. Streams, 1999−2000: A National
Reconnaissance. Environ. Sci. Technol. 36, 1202–1211.
doi:10.1021/es011055j
Konstantinou, I.K., Albanis, T.A., 2004. TiO2-assisted photocatalytic degradation of
azo dyes in aqueous solution: kinetic and mechanistic investigations. Appl.
Catal. B Environ. 49, 1–14. doi:10.1016/j.apcatb.2003.11.010
Lapworth, D.J., Baran, N., Stuart, M.E., Ward, R.S., 2012. Emerging organic
contaminants in groundwater: A review of sources, fate and occurrence.
Environ. Pollut. 163, 287–303. doi:10.1016/j.envpol.2011.12.034
Lekkerkerker-Teunissen, K., Benotti, M.J., Snyder, S.A., van Dijk, H.C., 2012.
Transformation of atrazine, carbamazepine, diclofenac and sulfamethoxazole
by low and medium pressure UV and UV/H2O2 treatment. Sep. Purif.
Technol. 96, 33–43. doi:10.1016/j.seppur.2012.04.018
Li, F., Liu, C., Liang, C., Li, X., Zhang, L., 2008. The oxidative degradation of 2-
mercaptobenzothiazole at the interface of β-MnO2 and water. J. Hazard.
Mater. 154, 1098–1105. doi:10.1016/j.jhazmat.2007.11.015
Li, X., He, Q., Li, H., Gao, X., Hu, M., Li, S., Zhai, Q., Jiang, Y., Wang, X., 2017.
Bioconversion of non-steroidal anti-inflammatory drugs diclofenac and
naproxen by chloroperoxidase. Biochem. Eng. J. 120, 7–16.
doi:10.1016/j.bej.2016.12.018
Liu, L., Zhang, J., Tan, Y., Jiang, Y., Hu, M., Li, S., Zhai, Q., 2014. Rapid
decolorization of anthraquinone and triphenylmethane dye using
chloroperoxidase: Catalytic mechanism, analysis of products and degradation
route. Chem. Eng. J. 244, 9–18. doi:10.1016/j.cej.2014.01.063
Liu, Z., Kanjo, Y., Mizutani, S., 2009. Removal mechanisms for endocrine
disrupting compounds (EDCs) in wastewater treatment — physical means,
biodegradation, and chemical advanced oxidation: A review. Sci. Total
Environ. 407, 731–748. doi:10.1016/j.scitotenv.2008.08.039
77
Loos, R., Wollgast, J., Huber, T., Hanke, G., 2007. Polar herbicides, pharmaceutical
products, perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA), and
nonylphenol and its carboxylates and ethoxylates in surface and tap waters
around Lake Maggiore in Northern Italy. Anal. Bioanal. Chem. 387, 1469–
1478. doi:10.1007/s00216-006-1036-7
Marchis, T., Avetta, P., Bianco-Prevot, A., Fabbri, D., Viscardi, G., Laurenti, E.,
2011. Oxidative degradation of Remazol Turquoise Blue G 133 by soybean
peroxidase. J. Inorg. Biochem. 105, 321–327.
doi:10.1016/j.jinorgbio.2010.11.009
Meetani, M.A., Hisaindee, S.M., Abdullah, F., Ashraf, S.S., Rauf, M.A., 2010.
Liquid chromatography tandem mass spectrometry analysis of
photodegradation of a diazo compound: A mechanistic study. Chemosphere
80, 422–427. doi:10.1016/j.chemosphere.2010.04.065
Méndez-Arriaga, F., Esplugas, S., Giménez, J., 2010. Degradation of the emerging
contaminant ibuprofen in water by photo-Fenton. Water Res., Emerging
Contaminants in water: Occurrence, fate, removal and assessment in the
water cycle (from wastewater to drinking water) 44, 589–595.
doi:10.1016/j.watres.2009.07.009
Mohapatra, D.P., Brar, S.K., Tyagi, R.D., Surampalli, R.Y., 2010. Degradation of
endocrine disrupting bisphenol A during pre-treatment and biotransformation
of wastewater sludge. Chem. Eng. J. 163, 273–283.
doi:10.1016/j.cej.2010.07.062
Nogueira, V., Lopes, I., Freitas, A.C., Rocha-Santos, T.A.P., Gonçalves, F., Duarte,
A.C., Pereira, R., 2015. Biological treatment with fungi of olive mill
wastewater pre-treated by photocatalytic oxidation with nanomaterials.
Ecotoxicol. Environ. Saf. 115, 234–242. doi:10.1016/j.ecoenv.2015.02.028
Oturan, M.A., Aaron, J.-J., 2014. Advanced Oxidation Processes in
Water/Wastewater Treatment: Principles and Applications. A Review. Crit.
Rev. Environ. Sci. Technol. 44, 2577–2641.
doi:10.1080/10643389.2013.829765
Papadakis, E.-N., Tsaboula, A., Kotopoulou, A., Kintzikoglou, K., Vryzas, Z.,
Papadopoulou-Mourkidou, E., 2015. Pesticides in the surface waters of Lake
Vistonis Basin, Greece: Occurrence and environmental risk assessment. Sci.
Total Environ. 536, 793–802. doi:10.1016/j.scitotenv.2015.07.099
Petrović, M., Gonzalez, S., Barceló, D., 2003. Analysis and removal of emerging
contaminants in wastewater and drinking water. TrAC Trends Anal. Chem.
22, 685–696. doi:10.1016/S0165-9936(03)01105-1
78
Poyatos, J.M., Muñio, M.M., Almecija, M.C., Torres, J.C., Hontoria, E., Osorio, F.,
2010. Advanced Oxidation Processes for Wastewater Treatment: State of the
Art. Water. Air. Soil Pollut. 205, 187. doi:10.1007/s11270-009-0065-1
Rauf, M.A., Ali, L., Sadig, M.S.A.Y., Ashraf, S.S., Hisaindee, S., 2016. Comparative
degradation studies of Malachite Green and Thiazole Yellow G and their
binary mixture using UV/H2O2. Desalination Water Treat. 57, 8336–8342.
doi:10.1080/19443994.2015.1017745
Rauf, M.A., Salman Ashraf, S., 2012. Survey of recent trends in biochemically
assisted degradation of dyes. Chem. Eng. J. 209, 520–530.
doi:10.1016/j.cej.2012.08.015
Robinson, T., McMullan, G., Marchant, R., Nigam, P., 2001. Remediation of dyes in
textile effluent: a critical review on current treatment technologies with a
proposed alternative. Bioresour. Technol. 77, 247–255. doi:10.1016/S0960-
8524(00)00080-8
Rodil, R., Quintana, J.B., Concha-Graña, E., López-Mahía, P., Muniategui-Lorenzo,
S., Prada-Rodríguez, D., 2012. Emerging pollutants in sewage, surface and
drinking water in Galicia (NW Spain). Chemosphere 86, 1040–1049.
doi:10.1016/j.chemosphere.2011.11.053
Rodríguez-Delgado, M., Ornelas-Soto, N., 2017. Laccases: A Blue Enzyme for
Greener Alternative Technologies in the Detection and Treatment of
Emerging Pollutants, in: Green Technologies and Environmental
Sustainability. Springer, Cham, pp. 45–65. doi:10.1007/978-3-319-50654-8_2
Ryan, B.J., Carolan, N., Ó’Fágáin, C., 2006. Horseradish and soybean peroxidases:
comparable tools for alternative niches? Trends Biotechnol. 24, 355–363.
doi:10.1016/j.tibtech.2006.06.007
Sánchez Peréz, J.A., Carra, I., Sirtori, C., Agüera, A., Esteban, B., 2014. Fate of
thiabendazole through the treatment of a simulated agro-food industrial
effluent by combined MBR/Fenton processes at μg/L scale. Water Res. 51,
55–63. doi:10.1016/j.watres.2013.07.039
Siegbahn, P.E.M., Blomberg, M.R.A., 2001. Mechanisms for enzymatic reactions
involving formation or cleavage of O-O bonds. Theor. Comput. Chem.,
Theoretical Biochemistry 9, 95–143. doi:10.1016/S1380-7323(01)80004-2
Silva, M.C., Corrêa, A.D., Amorim, M.T.S.P., Parpot, P., Torres, J.A., Chagas,
P.M.B., 2012. Decolorization of the phthalocyanine dye reactive blue 21 by
turnip peroxidase and assessment of its oxidation products. J. Mol. Catal. B
Enzym. 77, 9–14. doi:10.1016/j.molcatb.2011.12.006
79
Sorensen, J.P.R., Lapworth, D.J., Nkhuwa, D.C.W., Stuart, M.E., Gooddy, D.C.,
Bell, R.A., Chirwa, M., Kabika, J., Liemisa, M., Chibesa, M., Pedley, S.,
2015. Emerging contaminants in urban groundwater sources in Africa. Water
Res., Occurrence, fate, removal and assessment of emerging contaminants in
water in the water cycle (from wastewater to drinking water) 72, 51–63.
doi:10.1016/j.watres.2014.08.002
Sundaramoorthy, M., Terner, J., Poulos, T.L., 1995. The crystal structure of
chloroperoxidase: a heme peroxidase–cytochrome P450 functional hybrid.
Structure 3, 1367–1378. doi:10.1016/S0969-2126(01)00274-X
Trovó, A.G., Nogueira, R.F.P., Agüera, A., Fernandez-Alba, A.R., Sirtori, C.,
Malato, S., 2009. Degradation of sulfamethoxazole in water by solar photo-
Fenton. Chemical and toxicological evaluation. Water Res., AOPs for
Effluent Treatment 43, 3922–3931. doi:10.1016/j.watres.2009.04.006
Van der Bruggen, B., Cornelis, G., Vandecasteele, C., Devreese, I., 2005. Fouling of
nanofiltration and ultrafiltration membranes applied for wastewater
regeneration in the textile industry. Desalination, Conference on Fouling and
Critical Flux: Theory and Applications 175, 111–119.
doi:10.1016/j.desal.2004.09.025
Xu, B., Gao, N.-Y., Sun, X.-F., Xia, S.-J., Rui, M., Simonnot, M.-O., Causserand, C.,
Zhao, J.-F., 2007. Photochemical degradation of diethyl phthalate with
UV/H2O2. J. Hazard. Mater. 139, 132–139.
doi:10.1016/j.jhazmat.2006.06.026
Zhang, Q., Li, C., Li, T., 2013. UV/H2O2 Process Under High Intensity UV
Irradiation: A Rapid and Effective Method for Methylene Blue
Decolorization. CLEAN – Soil Air Water 41, 1201–1207.
doi:10.1002/clen.201100582
Zhang, X., Li, X., Jiang, Y., Hu, M., Li, S., Zhai, Q., 2016. Combination of
enzymatic degradation by chloroperoxidase with activated sludge treatment
to remove sulfamethoxazole: performance, and eco-toxicity assessment. J.
Chem. Technol. Biotechnol. 91, 2802–2809. doi:10.1002/jctb.4888
80
List of Publications
Al-Maqdi, K.A., Hisaindee, S.M., Rauf, M.A., Ashraf, S.S., 2017. Comparative
Degradation of a Thiazole Pollutant by an Advanced Oxidation Process and an
Enzymatic Approach. Biomolecules 7, 64. doi:10.3390/biom7030064