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

Follow this and additional works at: https://scholarworks.uaeu.ac.ae/all_theses

Part of the Chemistry Commons

This Thesis is brought to you for free and open access by the Electronic Theses and Dissertations at Scholarworks@UAEU. It has been accepted forinclusion in Theses by an authorized administrator of Scholarworks@UAEU. For more information, please contact fadl.musa@uaeu.ac.ae.

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

nte

nsit

y

10 12 14 16 18 20

0

100000

200000

300000

Time (min)

Rela

tive I

nte

nsit

y

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