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Effects of Nanostructured TiO2

Photocatalysis on

Disinfection By-product Formation

by

Aleksandra Sokolowski P.Eng.

A thesis submitted in conformity with the requirements

for the Masters in Applied Science degree

Graduate Department of Civil Engineering

University of Toronto

© Copyright by Aleksandra Sokolowski (2014)

ii

Effects of Nanostructured TiO2 Photocatalysis on

Disinfection By-product Formation

Aleksandra Sokolowski

Masters in Applied Science, 2014

Department of Civil Engineering

University of Toronto

ABSTRACT

The current research used simulated solar light and demonstrated that Aeroxide® P25

and innovative TiO2 photocatalytic nanomaterials decreased the trihalomethane (THM)

and haloacetic acid (HAA) formation potential (fp) in model and natural river water

sources by degrading natural organic matter (precursors) before disinfection with

chlorine. A low and high UV dose (28 and 827 mJ/cm2, respectively) were applied and,

overall, synthetic water THM fp reduced by up to 41 % and HAA fp reduced by up to 36

% while Otonabee River water THM fp reduced by up to 24 % and HAA fp reduced by

up to 13%. P25, P25 mixed with 1% of a silver-based product, anatase, and nitrogen

doped anatase performed relatively similarly. Advancement in treatment efficiencies

emerging from innovations in material science and reactor design, and understanding of

water quality impacts and degradation mechanisms, increase the feasibility of

incorporating TiO2 photocatalysis in drinking water treatment systems.

iii

ACKNOWLEDGEMENTS

I would like to thank my supervisor Susan Andrews for her encouragement, expertise,

and advice during my time as a graduate research student. It has been an honour to have

her as a female engineering academic mentor; her patience, intellect, and guidance kept

me on a steep learning curve. I am grateful for the insight provided by Ron Hofmann,

Natural Sciences and Engineering Research Council (NSERC) Industrial Research Chair

holder and second reader of my thesis and Robert Andrews, also an NSERC Industrial

Research Chair holder in the Drinking Water Research Group (DWRG).

This research was financially supported by an NSERC Strategic Project Grant. Project

meetings with project partners Trojan UV, Peterborough Utilities Commission (PUC),

Regional Municipalities of York and Peel, University of Waterloo Centre for Advanced

Materials Joining (CAMJ) and Department of Biology, and the University of Toronto

DWRG included thought provoking dialogue that shaped my thesis. Additional thanks to

the Southern Ontario Water Consortium (Solar Simulator) and PUC (raw water supply).

A special thanks to Stephanie Gora for all the invaluable collaboration on this project,

and Tassia Brito-Andrade and Adrielle Costa Souza who greatly assisted with meticulous

lab work. I would like to thank Robert Liang and Melissa Hatat-Fraile who provided the

innovative TiO2 nanomaterials studied in my experimental research. I am especially

grateful to Jim Wang, who taught me the analytical methods for disinfection by-products.

The assistance of the brilliant team in the DWRG is so appreciated; including Heather

Wray, Ding Wong, Jacque-Ann Grant, Stephanie Loeb, Jamal Azzeh, Emma Shen,

Russell D’Souza, Isabelle Netto, and Jules Carlson.

I would like to thank Wayne Lee, an engineering mentor who’s guidance in operating a

sole proprietorship lead to my licensure as a professional engineer. Much thanks to the

many mentors during my engineering career, including Troy Vassos who might find

some interest in my research work enclosed. A big ‘thank you’ to my supportive sister,

extended family, and friends. Most importantly, I thank my parents Elizabeth and Jerzy

Sokolowski and dedicate this thesis to them. Thank you for being by my side throughout

my life and while I have been working towards my master’s degree.

iv

TABLE OF CONTENTS

Abstract …………………………………………………………………………………...ii

Acknowledgements ……………………………………………………………………...iii

Nomenclature …………………………………………………………………………...xiv

1 INTRODUCTION ............................................................................................................. 1

1.1 Background ........................................................................................................... 1

1.2 Research Objectives .............................................................................................. 3

1.3 Outline of Chapters ............................................................................................... 3

2 LITERATURE REVIEW .................................................................................................... 5

2.1 Disinfection By-products ...................................................................................... 5

2.1.1 Human Health Concerns ............................................................................... 11

2.1.2 Regulations ................................................................................................... 12

2.2 DBP Precursors ................................................................................................... 14

2.2.1 Natural Organic Matter ................................................................................. 14

2.2.2 Anthropogenic Matter ................................................................................... 16

2.2.3 Inorganic Halides .......................................................................................... 16

2.3 Current and Emerging DBP Control Strategies .................................................. 16

2.4 DBP Formation Potential Tests .......................................................................... 20

2.5 TiO2 Photocatalysis ............................................................................................ 21

2.5.1 Mechanisms of Action .................................................................................. 22

2.5.2 Reaction Kinetics .......................................................................................... 25

2.5.3 Configurations for TiO2 Photocatalysis ........................................................ 27

2.5.4 Degradation of DBP Precursors .................................................................... 30

2.6 Research Gaps .................................................................................................... 34

3 MATERIALS AND METHODS ........................................................................................ 35

3.1 Materials ............................................................................................................. 36

3.2 Experimental Protocols ....................................................................................... 43

v

3.2.1 TiO2 Photocatalytic Procedures .................................................................... 43

3.2.2 UFC Chlorination Test .................................................................................. 46

3.3 Analytical Methods ............................................................................................. 47

3.3.1 Water Quality Parameters ............................................................................. 47

3.3.2 Chlorine Residual.......................................................................................... 47

3.3.3 Trihalomethane, Haloacetonitrile, Halonitromethane and Haloketone

Analysis......................................................................................................... 47

3.3.4 Haloacetic Acid Analysis .............................................................................. 49

3.3.5 Natural Organic Matter (DOC, UV254, LC-OCD) ........................................ 51

3.3.6 UV Fluence Rate ........................................................................................... 52

3.4 Statistical Analysis of Data ................................................................................. 53

3.4.1 Analysis of Variance ..................................................................................... 53

3.4.2 Coefficient of Determination ........................................................................ 55

3.5 QA/QC Measures ................................................................................................ 55

3.5.1 Analytical QA/QC ........................................................................................ 55

3.5.2 Experimental QA/QC.................................................................................... 56

4 PRELIMINARY TESTS AND TYPICAL DATA SETS.......................................................... 58

4.1 Overview of Experiments ................................................................................... 58

4.2 NOM Reduction .................................................................................................. 59

4.3 DBP fp Reduction ............................................................................................... 66

4.4 TiO2 Configurations ............................................................................................ 70

4.5 Optimum TiO2 Concentration ............................................................................. 71

4.6 Optimum TiO2 Dark Adsorption Time ............................................................... 74

4.7 UV Fluence Rate ................................................................................................. 75

4.8 Summary of Preliminary Results ........................................................................ 78

5 EFFECTS OF TIO2/UV ON DBP FORMATION IN A MODEL RIVER WATER ................... 79

5.1 Overview of Experiments and Results ............................................................... 79

vi

5.2 NOM Reduction .................................................................................................. 83

5.3 DBP fp Reduction ............................................................................................... 85

5.4 Summary of Results ............................................................................................ 93

6 EFFECTS OF TIO2/UV ON DBP FORMATION IN A NATURAL RIVER WATER ................ 94

6.1 Overview of Experiments and Results ............................................................... 94

6.2 Otonabee Water Quality ..................................................................................... 96

6.3 NOM Reduction .................................................................................................. 97

6.4 DBP Reduction ................................................................................................. 101

6.5 Comparison of Results for Synthetic and Natural Waters ................................ 110

6.6 Summary of Results .......................................................................................... 115

7 CONCLUSIONS ........................................................................................................... 116

8 RECOMMENDATIONS ................................................................................................. 119

9 REFERENCES ............................................................................................................. 120

10 APPENDICES .............................................................................................................. 131

10.1 Experimental Data for Chapters 5 and 6 ........................................................... 131

10.1.1 Calibration Data .......................................................................................... 131

10.1.1.1 DOC ................................................................................................... 131

10.1.1.2 THM .................................................................................................. 132

10.1.1.3 HAA ................................................................................................... 133

10.1.1.4 HAN, HNM, HK ............................................................................... 134

10.1.2 QA/QC ........................................................................................................ 135

10.1.2.1 DOC ................................................................................................... 135

10.1.2.2 THM .................................................................................................. 137

10.1.2.3 HAA ................................................................................................... 138

10.1.2.4 HAN, HNM, HK ............................................................................... 141

10.1.3 Supplementary Data .................................................................................... 143

10.1.3.1 Analysis of Variance ......................................................................... 143

vii

10.1.3.2 NOM Characterization ...................................................................... 144

10.1.3.3 UFC Chlorination Test ...................................................................... 146

10.1.3.4 THM fp .............................................................................................. 147

10.1.3.5 HAA fp .............................................................................................. 149

10.2 Experimental Data for Preliminary Experiments ............................................. 151

10.2.1 Optimal TiO2 Dark Adsorption Time ......................................................... 151

10.2.2 UV Fluence Rate ......................................................................................... 151

10.3 Sample Calculations ......................................................................................... 152

10.3.1 Determining DBP Concentration ................................................................ 152

10.3.2 Determining UV Dose in Published TiO2/UV Studies ............................... 153

viii

LIST OF TABLES

Table 2-1: Chlorination DBPs Studied in the Current Research Study .............................. 8

Table 2-2: DBP Regulations and Guidelines .................................................................... 13

Table 2-3: Comparison of TiO2/UV and other DBP Precursor Reduction Technologies 33

Table 3-1: Apparatus ......................................................................................................... 36

Table 3-2: Reagents .......................................................................................................... 37

Table 3-3: Characteristics of TiO2 Materials .................................................................... 40

Table 3-4: Synthetic and Otonabee River Water Characteristics ..................................... 42

Table 3-5: Preliminary Proof-of-Concept Experiments .................................................... 44

Table 3-6: Preliminary Optimization Experiments ........................................................... 45

Table 3-7: THM and HAN GC-ECD Instrumentation and Operating Conditions ........... 48

Table 3-8: HAA GC-ECD Instrumentation and Operating Conditions ............................ 50

Table 3-9: ANOVA Parameter Description ...................................................................... 53

Table 4-1: Summary % Reduction of THM and HAA fp in Preliminary Experiments

Following 60 min of TiO2/UV Treatment and Chlorination .......................... 71

Table 4-2: Pseudo First Order Reaction Rate Constants .................................................. 73

Table 4-3: UV Fluence Rate Raw Data Calculations ....................................................... 76

Table 4-4: Average UV Fluence Rate Calculations .......................................................... 77

Table 5-1: Two Way ANOVA Results for Synthetic Water ............................................ 82

Table 6-1: Two-way ANOVA for Otonabee Water.......................................................... 96

Table 6-2: THM and HAA Reported by the PUC (PUC, 2013) ....................................... 97

Table 6-3: Summary of DBP % Reduction with TiO2/UV ............................................. 113

Table 6-4: Comparison of THM and HAA fp % Reduction in TiO2/UV Studies .......... 114

Table 10-1: DOC Calibration Data for Synthetic Water Experiments ........................... 131

Table 10-2: DOC Calibration Data for Otonabee Water Experiments ........................... 131

Table 10-3: THM Calibration Data ................................................................................. 132

Table 10-4: THM MDL Results ..................................................................................... 132

ix

Table 10-5: HAA Calibration Data ................................................................................. 133

Table 10-6: HAA MDL Results ...................................................................................... 133

Table 10-7: HAN Calibration Data ................................................................................. 134

Table 10-8: HAN MDL Data .......................................................................................... 134

Table 10-9: QA/QC Data for DOC Analysis .................................................................. 136

Table 10-10: ANOVA Output for UV254 in Synthetic Water ......................................... 143

x

LIST OF FIGURES

Figure 2-1: Example DBPs from DBP Classes Studied in the Current Research .............. 9

Figure 2-2: TiO2 Photo Reactivity, Source: Crittenden et al., 2005 ................................. 23

Figure 3-1: General Schematic of Experiments ................................................................ 35

Figure 3-2: Solar Simulator .............................................................................................. 38

Figure 3-3: Spectral Radiation of Solar Simulator ........................................................... 38

Figure 3-4: As prepared (a) NB, (b) Ag@SiO2@TiO2 (c) 1% Ag@SiO2@TiO2/P25,

(d) Anatase, (e) Anatase-N, and (f) Anatase-B .............................................. 39

Figure 3-5: Residue from Otonabee Water Filtration ....................................................... 41

Figure 3-6: Batch Experimental Set-up ............................................................................ 43

Figure 4-1: DOC and UV254 in Synthetic Water Treated with 0.5 g/L P25 in

Suspension ..................................................................................................... 59

Figure 4-2: UV-Vis Absorbance of Synthetic water Treated with P25 TiO2/UV ............ 60

Figure 4-3: DOC of Otonabee Water Treated with 0.5 g/L P25 in Suspension ............... 61

Figure 4-4: NOM Fractions in Otonabee Water Treated with 0.5 g/L P25 in

Suspension ..................................................................................................... 62

Figure 4-5: DOC and UV254 in Otonabee Water Treated with TiO2/UV at 0.5 and 0.15

g/L in Suspension ........................................................................................... 63

Figure 4-6: DOC and UV254 in Otonabee Water Treated with P25 at 0.15 g/L in

Suspension, P25 Immobilized as a Thin Film, and NB in Suspension .......... 64


Suspension (Duplicate Experiment) ............................................................... 65

Figure 4-8: THM fp in Synthetic Water Following Treatment with P25 at 0.5 g/L in

Suspension and Chlorination ......................................................................... 67

Figure 4-9: THM fp in Otonabee River Following Treatment with TiO2/UV and

Chlorination ................................................................................................... 68

Figure 4-10: HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV at

0.5 g/L in Suspension and Chlorination ......................................................... 69

xi

Figure 4-11: HAA fp in Otonabee Water Following Treatment with TiO2/UV and

Chlorination ................................................................................................... 69

Figure 4-12: Determining Reaction Rate Constant for TiO2/UV at 0.1 g/L ..................... 72

Figure 4-13: DOC in Synthetic Water Following Treatment with P25 TiO2/UV under

Various Dark Adsorption and Irradiation Times ........................................... 75

Figure 5-1: DOC in Synthetic Water Following P25 TiO2/UV Treatment ...................... 83

Figure 5-2: SUVA in Synthetic Water Following P25 TiO2/UV Treatment .................... 84

Figure 5-3: Synthetic Water SUVA % Reduction Following TiO2/UV Treatment .......... 85

Figure 5-4: THM fp in Synthetic Water Following Treatment with P25 TiO2/UV and

Chlorination ................................................................................................... 87

Figure 5-5: HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV and

Chlorination ................................................................................................... 87

Figure 5-6: THMfp/SUVA in Synthetic Water Following Treatment with P25

TiO2/UV and Chlorination ............................................................................. 89

Figure 5-7: HAAfp/SUVA in Synthetic Water Following Treatment with P25

TiO2/UV and Chlorination ............................................................................. 89

Figure 5-8: Sp THM fp in Synthetic Water Following Treatment with P25 TiO2/UV

and Chlorination ............................................................................................. 90

Figure 5-9: Sp HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV

and Chlorination ............................................................................................. 91

Figure 5-10: THM fp % Reduction in Synthetic Water Following Treatment with

Various Nanostructured TiO2/UV and Chlorination ...................................... 92

Figure 5-11: HAA fp % Reduction in Synthetic Water Following Treatment with

Various Nanostructured TiO2/UV and Chlorination ...................................... 93

Figure 6-1: DOC in Otonabee Water Following P25 TiO2/UV Treatment ...................... 98

Figure 6-2: SUVA in Otonabee Water Following P25 TiO2/UV Treatment .................. 100

Figure 6-3: SUVA % Reduction in Otonabee Water Following TiO2/UV Treatment ... 101

xii

Figure 6-4: THMfp in Otonabee water Following Treatment with P25 TiO2/UV and

Chlorination ................................................................................................. 103

Figure 6-5: HAA fp in Otonabee water Following Treatment with P25 TiO2/UV and

Chlorination ................................................................................................. 104

Figure 6-6: THMfp/SUVA in Otonabee Water Following Treatment with P25

TiO2/UV and Chlorination ........................................................................... 106

Figure 6-7: HAAfp/SUVA in Otonabee Water Following Treatment with P25

TiO2/UV and Chlorination ........................................................................... 106

Figure 6-8: THMfp/DOC in Otonabee Water Following Treatment with P25 TiO2/UV

and Chlorination ........................................................................................... 107

Figure 6-9: HAAfp/DOC in Otonabee Water Following Treatment with P25 TiO2/UV

and Chlorination ........................................................................................... 107

Figure 6-10: THM fp % Reduction in Otonabee Water Following Treatment with

Various Nanostructured TiO2/UV and Chlorination .................................... 108

Figure 6-11: HAA fp % Reduction in Otonabee Water Following Treatment with

Various Nanostructured TiO2/UV and Chlorination .................................... 109

Figure 6-12: TiO2/UV Treatment Comparison between Otonabee and Synthetic Water 111

Figure 10-1: THM Calibration Curves ........................................................................... 132

Figure 10-2: THM QA/QC Charts .................................................................................. 137

Figure 10-3: HAA QA/QC Charts .................................................................................. 139

Figure 10-4: HAA QA/QC Charts .................................................................................. 140

Figure 10-5: HAN QA/QC Charts .................................................................................. 141

Figure 10-6: HNM and HK QA/QC Charts .................................................................... 142

Figure 10-7: DOC and UV254 in TiO2/UV Treated Synthetic Water .............................. 144

Figure 10-8: DOC and UV254 in TiO2/UV Treated Otonabee Water .............................. 145

Figure 10-9: UFC Chlorination Test Data for AgSiO2/P25 and Anatase TiO2/UV

Treated Otonabee Water .............................................................................. 146

xiii

Figure 10-10: THM fp in Synthetic Water Following Treatment with TiO2/UV and

Chlorination ................................................................................................. 147

Figure 10-11: THM fp in Otonabee Water Following Treatment with TiO2/UV and

Chlorination ................................................................................................. 148

Figure 10-12: HAA fp in Synthetic Water Following TiO2/UV and Chlorination ......... 149


Chlorination ................................................................................................. 150

Figure 10-14: UV254 in Synthetic Water Following Treatment with P25 TiO2/UV

under Various Dark Adsorption and Irradiation Times ............................... 151

Figure 10-15: UV-Vis Absorbance of a 0.1 g/L TiO2 suspension in Milli-Q®.............. 151

Figure 10-16: P25 TiO2/UV Methylene Blue Degradation with and without a Vortex . 152

xiv

NOMENCLATURE

Ag silver

Ag@SiO2@TiO2 triplex core-shell photocatalyst

Ag@SiO2@TiO2/P25 triplex core-shell photocatalyst mixed with P25 (1:99 ratio)

ANOVA analysis of variance

AOC assimilable organic carbon

AOP advanced oxidation process

B boron

BCAA bromochloroacetic acid

BCAN bromochloroacetonitrile

BDCAA bromodichloroacetic acid

BDCM bromodichloromethane

Br- bromide ion

BrO3- bromate

C carbon

CB conduction band

Cl chlorine

CL control limit

Cl- chloride ion

Cl2 chlorine gas

ClO2 chlorine dioxide

ClO2- chlorite

ClO3- chlorate

cm centimeter

cm2 centimeter squared

CO2 carbon dioxide

CP chloropicrin (trichloronitromethane)

CSTR continuously stirred tank reactor

d delta (the change in)

DBAA dibromoacetic acid

DBAN dibromoacetonitrile

DBP disinfection by-product

DBCAA dibromochloroacetic acid

DBCM dibromochloromethane

DCAA dichloroacetic acid

DCAN dichloroacetonitrile

DCP 1,1-dichloro-2-propanone

DF divergence factor

DI deionized

xv

DOC dissolved organic carbon

DON dissolved organic nitrogen

DWRG Drinking Water Research Group

E band gap energy, Einstein (energy of one mole of photons)

e- electron

e.g. for example

Eq. equation

Eqs. equations

eV electron volt

fp formation potential

FP formation potential (test)

g gram

GC-ECD gas chromatography – electron capture detection

H hydrogen

h+ hole (valence band hole from loss of electron)

H+ hydrogen ion

photon energy

H2O water

H2O2 hydrogen peroxide

HA humic acids

HAA haloacetic acid

HAN haloacetonitrile

HCCl3 trichloromethane

HK haloketone

HNM halonitromethane

HO*

hydroxyl radical

HO-

hydroxyl ion

HOBr hypobromous acid

HOCl hypochlorous acid

HOI hypoiodous acid

hr hour

i.e. in essence

IO3- iodate

k reaction rate constant

KAds adsorption constant

KJ kilojoule

kWh kilowatt hour

L liter

L-H Langmuir-Hinshelwood

LC-OCD liquid chromatography – organic carbon detection

xvi

LCL lower control limit

LMW low molecular weight

ln natural logarithm

LP low pressure

LWL lower warning limit

M mole per liter

m3 cubic meters

MBAA monobromoacetic acid

MCAA monochloroacetic acid

MDL method detection limits

mg milligram

Milli-Q® ultrapure laboratory grade water

min minutes

mJ millijoule

mL milliliter

MP medium pressure

MTBE methy-tert-butyl-ether

mW milliwatt

MX 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone

N nitrogen

N-DBP nitrogenous DBP

N2 nitrogen gas

NaOCl sodium hypochlorite

NB nanobelts

NCl3 trichloramine

NDMA N-nitrosodimethylamine

ng nanogram

NH2Cl monochloramine

NH3 ammonia

NHCl2 dichloramine

nm nanometer

NOM natural organic matter

NSERC Natural Science and Engineering Research Council

O oxygen

O2 oxygen molecule

O2*-

super oxide

O3 ozone

OBr- hypobromite ion

OCl-

hypochlorite ion

OI- hypoiodite ion

xvii

P25 Aeroxide® P25 industry standard nanostructured TiO2

PEC photoelectrocatalysis

PES polyethersulfone

pH decimal logarithm of reciprocal of hydrogen ion activity

PPCP pharmaceuticals and personal care products

PUC Peterborough Utilities Commission

QA/QC quality assurance/quality control

r reaction rate

R organic molecule

R- organic molecule, negatively charged

R* organic radical

R2

coefficient of determination

RC=CR’ organic molecule with double bond

RC(OH)C(Cl)R’ chlorinated hydrocarbon

RCHO carbonyl

RCOCH3 ketone

RCOOH carboxylic acid

RF reflection factor

ROS reactive oxygen species

RPM rotations per minute

Ru ruthenium

s seconds

SBR sequencing batch reactor

SDS simulated distribution systems

SEM scanning electron microscope

SiO2 silicon dioxide (silica)

SOC synthetic organic compound

sp specific (normalized to DOC)

SUVA specific ultraviolet absorbance

t time

TBAA tribromoacetic acid

TBM tribromomethane (bromoform)

TCAA trichloroacetic acid

TCAN trichloroacetonitrile

TCM trichloromethane (chloroform)

TCP 1,1,1-trichloropropanone

TFBA 2,3,4,5-tetrafluorobenzoic acid

THM trihalomethane

THNM trihalonitromethane

Ti titanium

xviii

TiO2 titanium dioxide

TiO2/UV titanium dioxide photocatalysis

tsp teaspoon

UCL upper control limit

UFC uniform formation conditions

US EPA United States Environmental Protection Agency

UV ultraviolet light

UV-Vis ultraviolet and visible light

UV254 ultraviolet absorbance at 254 nanometer wavelength

UVA ultraviolet light, 315 – 400 nm

UVB ultraviolet light, 280 – 315 nm

UVC ultraviolet light, 200 – 280 nm

UWL upper warning limit

V volt

VB valence band

WF water factor

WL warning limit

WTP water treatment plant

wavelength

µ micro

µg microgram

µL microliter

µm micrometer

% percent

® registered trademark

> greater than

< less than

@ at

+ plus

- minus, carbon to carbon single bond in molecule

chemical reaction direction

chemical reaction (equilibrium) oC degrees Celsius

= equal to, carbon to carbon double bond in molecule

/ divided by, or

[] concentration

fraction

A. Sokolowski Effects of TiO2/UV on DBP fp

1

1 INTRODUCTION

1.1 Background

Since the early 1970’s, research has shown that the process of disinfecting drinking water

to kill and/or inactivate pathogens may have the unintended consequence of forming by-

products (Rook, 1976). These by-products are the result of the reaction of the disinfectant

with organic or inorganic matter present in the water. Over 600 disinfection by-products

(DBPs) have been identified from disinfection by chlorination yet more than 50% of the

total organic halogen formed remain unidentified (Pressman et al., 2010, Richardson,

2007). DBP formation depletes the amount of disinfectant available for microbiological

control, and some DBPs have adverse human health effects, including carcinogenicity

and genotoxicity (Pressman et al., 2010; Richardson et al., 2007). Trihalomethanes

(THMs) and haloacetic acids (HAAs) are typical products of the reaction of free chlorine

with natural organic matter (NOM) and account for the major two classes of DBPs

formed by chlorination (Villanueva, 2012). Some of them have been considered to

have/be associated with potential human health concerns (Jeong et al., 2012), although

this is being re-examined (Hrudey, 2009), and they are regulated world-wide (Villanueva,

2012). They were some of the first DBPs identified and are relatively easy to detect and

quantify (Weinberg, 1999). Haloacetonitriles (HANs), halonitromethanes (HNMs),

haloketones (HKs) are re-emerging classes of DBPs of concern based on their occurrence

and potential health effects and are associated with the reaction of free chlorine with

NOM (Richardson, 2007; Plewa et al., 2008; Krasner et al., 2006). The Canadian

Drinking Water Quality Guideline for THMs is 100 µg/L (maximum allowable

concentration) and for HAAs is 80 µg/L (as low as reasonably achievable); there are no

guidelines for HANs, HNMs or HKs (Health Canada, 2012).

The use of technologies to reduce the concentration or reactivity of NOM before

disinfection and/or alternative disinfection practices are effective DBP management

strategies and are actively being researched. TiO2 photocatalysis with ultraviolet-based

photoactivation (TiO2/UV) can achieve both these functions while requiring no chemical

addition other than the initial TiO2 catalyst. It can be categorized as an advanced


2

oxidation process (AOP) because research shows it relies significantly on the hydroxyl

radical (HO*) that is produced. Nanostructured TiO2/UV has been shown to reduce

THMs and HAAs during subsequent chlorination by altering or removing precursors (Liu

et al., 2008a; Liu et al., 2008b). However, it has also been shown to increase the

formation potential (fp) of these DBPs during subsequent disinfection by altering NOM

into more reactive compounds (Liu et al., 2008b). Nonetheless, research shows that with

sufficient treatment the concentration of DBP precursors is significantly reduced and the

remaining recalcitrant compounds are much less reactive (Richardson et al., 1996; Liu et

al., 2010). This technology compares favorably to other DBP control management

strategies that aim to reduce precursors and is also a promising alternative disinfection

strategy that appears to produce innocuous by-products (Richardson et al., 1996). The

treatment efficiency of TiO2/UV is dependent on such parameters as UV dose (mJ/cm2),

TiO2 type and concentration, and reactor configuration.

To further understand how THM, HAA, HAN, HNM, and HK fp may be managed by

TiO2/UV treatment prior to chlorination, the current research studied the formation

potential of these DBPs in model and real waters with newly developed and industry

standard TiO2 nanomaterials. The degradation of DBP precursors was investigated

through the measurement of dissolved organic carbon (DOC), 254 nm ultraviolet light

(UV) absorbance, and concentrations of NOM fractions via liquid chromatography-

organic carbon detection (LC-OCD). The uniform formation condition chlorination test,

which employs a chlorine residual of 1 mg/L after 24 hr at a pH of 8 and temperature of

20oC was followed to produce the DBPs. An SS150AAA Solar Simulator was the light

source for the experiments. It matched the natural solar electromagnetic radiation

spectrum at approximately 108 mW/cm2 “one sun” light intensity (300 to 1100 nm),

including approximately 13 mW/cm2 as UV-Vis light (300-424 nm) within the

photoactive range of the TiO2 nanomaterials studied.


3

1.2 Research Objectives

The current research objective was to study the changes in formation potential of

common (THM and HAA) and re-emerging (HAN, HNM, and HK) drinking water

chlorination DBPs with TiO2 photocatalytic treatment prior to chlorination. The overall

hypotheses were that TiO2/UV would either increase or decrease DBP fp, and that

different types of TiO2 would have varying degrees of effect. This research also

considered both short-term exposures (up to a few minutes) representative of flow-

through treatment systems, and longer term exposures of up to 30 min or more that may

be more representative of batch reactor systems. The specific objectives were broken

down as follows:

1. To examine % reduction of THM, HAA, HAN, HNM and HK precursors with

innovative and industry standard nanostructured TiO2/UV. The new materials have

been fabricated by project partners to sensitize TiO2 to visible light and increase

quantum efficiency.

2. To compare the efficiency of selected TiO2/UV systems for treating synthetic river

water and a natural river water source. Lab-prepared ‘synthetic water’ was used to

limit the variations in water quality from different real sources.

1.3 Outline of Chapters

Chapter 2 reviews the literature pertaining to disinfection by-products and provides a

brief summary of disinfection and the DBPs studied (THM, HAA, HAN, HNM and

HK). The health concerns of these DBPs and the guidelines and standards used to

regulate them is provided. Best practices to control DBP formation in drinking water

treatment, DBP precursors, and formation pathways are discussed. A review of TiO2

photocatalysis is also presented, including mechanisms of action, reaction kinetics,

configurations, and degradation pathways.

Chapter 3 provides an overview of the materials and methods for the experiments

conducted. TiO2/UV and UFC chlorination protocols are described and the analytical

methods employed to quantify THM, HAA, HAN, HNM and HK, UV254 absorbance,


4

DOC, and LC-OCD are summarized. The method to determine the UV dose based on

UV fluence rate and irradiation time is provided. Statistical methods used to analysis

experimental data are explained and QA/QC measures taken are summarized.

Preliminary results are given in Chapter 4including six proof-of-concept TiO2/UV

experiments preformed as an initial reconnaissance in the lab. DBP and DBP precursor

reduction are discussed with respect to TiO2 concentration and configuration, and

source water. Preliminary experiments to investigate optimal TiO2 concentration and

dark adsorption time, and chose photocatalytic reaction times used in subsequent

experiments are also summarized. Calculations to determine UV fluence rate through

the sample during photocatalysis are also provided.

Chapter 5 summarizes the results of experiments using lab-controllable synthetic

water that compare DBP formation and precursor reduction with industry standard

Aeroxide® P25 and innovative TiO2 nanomaterials. The new materials included P25

mixed with a silver based product, nanobelts, anatase, nitrogen doped anatase, and

boron doped anatase.

Chapter 6discusses the results of experiments using Otonabee River water, a natural

river water source. These experiments used the same TiO2 nanomaterials and

treatment conditions as those with synthetic water. The results are compared with the

synthetic river water experiments and similar TiO2 experiments from literature.

Concluding remarks are provided in Chapter 7while Chapter 8provides

recommendations for future research.

Chapter 9lists the references used throughout this thesis.

Raw environmental, calibration, and quality control and quality assurance (QA/QC)

data, and sample calculations are appended as Chapter 10.0.


5

2 LITERATURE REVIEW

2.1 Disinfection By-products

Research has shown that the process of disinfecting drinking water to kill and/or

inactivate pathogens may have the unintended consequence of forming by-products.

Their management and control in the drinking water treatment process has become a

major focus in drinking water treatment since the 1970’s when they were first discovered

(Rook, 1976). Although most of the focus of DBP research has been in drinking water

treatment; sanitary, industrial, and recreational water treatment may produce higher levels

of DBPs during disinfection if they have higher concentrations of precursors (Chen et al.,

2010, Richardson, 2003).

Drinking water disinfection is generally classified as either primary or secondary.

Primary disinfection kills or inactivates microorganisms including pathogenic viruses,

bacteria, and protozoa. Secondary disinfection is the application of a long lasting

disinfectant that remains active after water leaves a water treatment plant (WTP) and

enters a distribution system, protecting the distribution system and ensuring that water

quality objectives are met at the point of use. Some drinking water treatment plants that

draw on surface water also practice seasonal prechlorination at the source water intake to

control zebra mussel control.

In the early 1900’s the spread of waterborne illness through potable water supplies was

significantly restrained with the application of chlorine disinfection to surface and

groundwater sources. Disinfection by chlorine is accomplished by the inactivation of

microorganisms through oxidation of cell wall constituents causing lysis and death.

Chlorine is a strong oxidant second only to fluorine and has fast kinetics. It is economical

to manufacture and safe to handle and store. It continues to be employed in drinking

water treatment worldwide.

Chlorine is added to water as sodium hypochlorite (NaOCl) or chlorine gas (Cl2), forming

hypochlorous acid (HOCl), and hypochlorite ion (OCl-) upon equilibrium with water

according to Eqs. (2.1) and (2.2):


6

(2.1)

(2.2)

The exact mechanisms by which DBPs are formed are mostly unknown and involve

many chain reactions with complex organic and inorganic constituents in water

(Villanueva et al., 2012; Norwood et al., 1987). Some DBPs will form quickly during

primary disinfection in the clearwell and others will form gradually with the secondary

disinfectant in the distribution system. Increasing disinfectant and precursor

concentrations and reaction time will increase DBP formation (Crittenden et al., 2005).

During chlorination several reactions can take place between chlorine and contaminants

to create DPBs including the oxidation of organic and inorganic constituents, ammonia

substitution, chlorine substitution of hydrogen, decomposition, and chlorine addition to

double bonds.

Eqs 2.3 and 2.4 provide examples of the oxidation of a carbonyl to carboxylic acid and

subsequent decarboxylation, respectively. Complete oxidation of organic compounds to

CO2 does not typically occur at the Cl2 concentrations used in drinking water disinfection

and remaining compounds can include organics such as carboxylic acids (Richardson,

2003).

(2.3)

(2.4)

Halo organic compounds typically associated with chlorination DBPs can form from

substitution and addition reactions between chlorine and organic molecules. In a

substitution reaction chlorine replaces hydrogen in a hydrocarbon molecule. In Eq. 2.5

for example, free chlorine attacks the ketone at the carbonyl (carbon-carbon bond),

forming a carboxylic acid and trichloromethane. In an addition reaction, a carbon-carbon

double bond is broken and chlorine binds to a carbon, as shown in Eq. 2.6.

(2.5)

( ) ( ) (2.6)

The reaction kinetics of addition and substitution reactions between chlorine with NOM

are in the order of hours and days (Crittenden et al., 2005) and thus concentrations of the


7

products of these reactions will be highest at locations along a distribution system where

the hydraulic retention time is the greatest.

The oxidation of inorganic compounds, particularly bromide and iodide by chlorine are

also of concern for inorganic and organic DBP formation. Chlorine will oxidize elements

that are less electronegative than it is, so these reduced halogen species are prime targets.

The reaction of free chlorine with bromide ions is described in Eq. 2.7, and a similar

reaction occurs between iodide and chlorine. The oxidized products of these reactions

(e.g. hypobromous acid) can in turn oxidize other contaminants to create bromo- and

iodo- DBPs.

(2.7)

When ammonia is present, chlorine successively replaces hydrogen atoms in the

ammonia molecule (NH3) to create chloramines and finally N2 gas. The reaction kinetics

between chlorine and ammonia and chloramines is very fast, consuming the initial

chlorine added during disinfection, and must be accounted for to obtain required chlorine

residual. Other readily oxidizable species such as iron and manganese contribute to this

“instantaneous” chlorine demand (Crittenden et al., 2009). Chlorine will also decompose

to yield chlorate, which is considered an inorganic DBP with associated health concerns

(Crittenden et al., 2009).

Known classes of halogenated DBPs from chlorination include halomethanes, haloacids,

halonitriles, haloketones, haloaldehydes, halonitromethanes, haloamides, halofuranones,

haloacetamides, haloacetonitriles, and oxyhalides. The types and concentrations of DBPs

formed will depend on factors including type and concentration of disinfectant used,

available NOM precursors, presence of bromide or iodide, reaction time, and overall

water quality (pH, alkalinity, temperature). An example distribution of the DBPs formed

at a chlorination demonstration plant is: 2.8 % bromochloroacetic acid, 10 % haloacetic

acids (HAA5), 1.5 % chloral hydrate, 2 % haloacetonitriles, 20.1 % trihalomethanes

(THMs), 1 % cyanogen chloride, and 62.4 % unidentified organic halides (Richardson,

2003). The HAA5 referred to the sum of the 5 HAAs regulated by the US EPA:

monochloro-, dichloro-, trichloro-, monobromo-, and dibromoacetic acid.


8

Trihalomethanes (THMs), haloacetic acids (HAAs) and to a lesser degree

haloacetonitriles (HANs) are typical products of the reaction of free chlorine disinfectant

with natural organic matter (NOM) in water.

Table 2-1 provides a list of the DBPs studied in the current research project. The list

includes the chloro- and bromo- THMs and HAAs produced in relatively high

concentrations during chlorination as well as other DBPs that are not as common but still

have been identified from chlorination (Yang et al., 2007; Richardson et al., 1996;

Weinberg et al., 2002).

Table 2-1: Chlorination DBPs Studied in the Current Research Study

By-product Molecular Formula Acronym

Trihalomethanes THMs

Chloroform CHCl3 TCM

Bromodichloromethane CHBrCl2 BDCM

Dibromochloromethane CHBr2Cl DBCM

Bromoform CHBr3 TBM

Haloacetic acids HAAs

Monochloroacetic acid CH2ClCOOH MCAA

Dichloroacetic acid CHCl2COOH DCAA

Trichloroacetic acid CCl3COOH TCAA

Bromochloroacetic acid CHBrClCOOH BCAA

Bromodichloroacetic acid CBrCl2COOH BDCAA

Dibromochloroacetic acid CBr2ClCOOH DBCAA

Monobromoacetic acid CH2BrCOOH MBAA

Dibromoacetic acid CHBr2COOH DBAA

Tribromoacetic acid CBr3COOH TBAA

Haloacetonitriles HANs

Trichloroacetonitrile CCl3CN TCAN

Dichloroacetonitrile CHCl2CN DCAN

Bromochloroacetonitrile CHBrClCN BCAN

Dibromoacetonitrile CHBr2CN DBAN

Trihalonitromethanes THNMs

Trichloronitromethane

(chloropicrin) CCl3NO2 CP

Haloketones HKs

1,1-Dichloro-2-propanone C3H4Cl2O DCP

1,1,1-Trichloropropanone C3H4Cl2O TCP


9

These and other DBPs are typically present in drinking water at µg/L or ng/L levels, so

their detection and measurement requires rigorous analytical methods. The DBPs studied

during this thesis are typically analyzed by gas chromatography with electron capture

detection (GC-ECD) and are relatively easier to study compared to other DBPs because

they are thermally and chemically stable, volatile/semi-volatile, and/or neutral (Weinberg

et al., 2002; Mori et al., 2013; Philippe et al., 2010; Li et al, 1996). Example compound

chemical structures from each DBP class studied are provided in Figure 2-1.

Figure 2-1: Example DBPs from DBP Classes Studied in the Current Research

The four trihalomethanes bromoform (also known as tribromomethane, TBM),

chloroform (trichloromethane, TCM), chlorodibromomethane (CDBM), and

bromodichloromethane (BDCM) are herein referred to as THMs. There are also other

iodo-THMs which are considered re-emerging DBPs but these were not studied in the

current research. THMs occur in low to mid µg/L levels in chlorinated drinking water

with TCM typically highest (Richardson, 2007). The bromo- and iodo- THMs will only

be present if the source water contains bromide or iodide.


10

There are many haloacetic acids (HAAs) and the term can be used to include a range of

compounds. Typically, they are reported as a subset of 5, 6 or 9 chloro- bromo- acetic

acids labelled HAA5, HAA6 and HAA9, respectively. For example, the Canadian

Drinking Water Quality Guidelines, HAA as low as reasonably achievable guideline is

for the HAA5 subset monochloro-, dichloro-, trichloro-, monobromo-, dibromoacetic

acids, which is the same HAA5 subset regulated by the US EPA and State of California

(Health Canada, 2012; US EPA, 2009; State of California, 2013). The subset of HAA6

typically includes bromochloroacetic acid. The current research investigated the nine

haloacetic acids listed in Table 2-1 and herein refers to them collectively as HAAs.

Similarly with THMs, the bromo- and iodo- HAAs will only be present if the source

water contains bromide or iodide. Mono-, di- and trichloroacetic acids have one, two

and three chlorine molecules, respectively and likewise with the brominated HAAs.

Monochloroacetic acid typically quickly converts to the di- and trichloroacetic acids in

the presence of free chlorine.

Haloacetonitriles (HANs) and halonitromethanes (HNMs) are nitrogenous classes of

DBPs and are associated with the reaction of free chlorine with nitrogeneous NOM

(Richardson, 2007; Plewa et al., 2008; Krasner et al., 2006). Chloramines may also react

to produce nitrogenous DBPs including HANs and HNMs (Plewa et al., 2008).

Nitrogeneous DBP concentrations are typically low (Villanueva, 2012). One study found

up to 14 µg/L of HANs (approximately 10% of THM) and up to 10 µg/L

halonitromethanes (HNMs) during chlorination of water with high precursor loading

(Krasner et al., 2006). The current study includes the analysis of four chloro- bromo-

acetonitriles and one trihalonitromethanes, trichloronitromethane, also commonly known

as chloropicrin (CP).

Haloketones are also considered re-emerging DBPs of concern and the aforementioned

study found up to 9 µg/L (Krasner et al., 2006). Two haloketones, 1,1-dichloro-2-

propanone (DCP) and 1,1,1-trichloropropanone (TCP), were included in the current

research. TCP, as well as HAN, THM, and HAA, have been identified following

treatment with TiO2/UV and chlorination, albeit generally in lower concentrations

compared to chlorination alone (Richardson et al., 1996).


11

2.1.1 Human Health Concerns

Studies have shown that a lifetime exposure of chlorinated drinking water increases the

risk for cancer, and this has in part been attributed to DPBs (Pressman et al., 2010;

Richardson et al., 2007). DBPs have also been linked to having adverse reproductive and

developmental effects (Richardson, 2003). Of the 600 known DBPs, few have been

studied in depth for toxicological information, and all of the unknown DBPs pose a

potential health risk as well (Richardson et al., 2007). Also, most toxicity/exposure

studies have focused on ingestion of DBPs but other exposure routes such as dermal

absorption and inhalation through activities such as showering must also be further

investigated, especially for volatile DBPs like THMs (Richardson, 2007).

The Guidelines for Canadian Drinking Water Quality (Health Canada, 2012) list THMs

as having confirmed liver effects (fatty cysts), and causing kidney and colorectal cancers.

Chloroform is classified as a possible carcinogen. DCAA is confirmed as causing liver

cancer and is classified as a probable carcinogen; DCAA, DBAA and TCAA are

confirmed as causing organ cancers; and MCAA as effecting body, kidney and testes

weights. Studies have also found that brominated DBPs are more genotoxic and

carcinogenic than their chlorinated counterparts, and similarly for iodinated DBPs

(Richardson 2007).

Although THMs and HAAs have been linked to carcinogenicity and other adverse health

effects in many studies (Richardson et al, 2007), ongoing research in the field is

suggesting that their actual risk is low compared to other unregulated DBPs (Hrudey,

2009). For example, 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (referred to

as Mutagen X or MX) was found to produce 20 – 50 % of the toxicity of drinking water

(Kronberg et al., 1988) even though it is typically present at concentrations 2-3 orders of

magnitude lower than THMs and HAAs (Richardson et al., 2007).

Haloacetonitriles (HAN), haloketones (HK), and halonitromethanes (HNM) have also

been labeled as high priority compounds due to their health concerns and occurrence

(Weinberg et al., 2002; Richardson, 2003). Studies suggest that nitrogeneous DBPs may


12

be more harmful than their carbonaceous counterparts. HANs and HNMs are

nitrogeneous DBPs that appear to be genotoxic and cytotoxic (Plewa et al., 2004a).

2.1.2 Regulations

Regulators around the world have set limits on the levels of specific DBPs permitted in

drinking water. These limits must be met regardless of the process(s) used for water

treatment. THMs and HAAs are the most commonly regulated organic DBPs and a

summary table of permitted or recommended maximum levels of THMs and HAAs in a

selection of jurisdictions are provided in Table 2-2 (Health Canada, 2012; European

Commission, 1998; Government of Singapore, 2008; Ontario Ministry of the

Environment, 2008; US EPA, 2009; World Health Organization, 2011; Ministry of

Environmental Protection of the Government of the People’s Republic of China, 2006;

State of California, 2013). Table 2-2 also includes data for two HANs studied in this

report however none of the selected jurisdictions had any regulation of the other HANs,

HNM or HKs studied herewith.

DBPs are increasingly regulated, and other regulated DBPs in the jurisdictions reviewed

include nitrosamines, organic such as formaldehyde, and inorganics such as bromate,

chlorate and chlorite.

The human health effects of ingesting THMs and HAAs in drinking water have been an

issue of debate. However, it is unlikely that the regulations will be removed anytime in

the near future. The regulation of THM and HAA results in their monitoring which can

be used to generally determine overall DBP levels. Analytical procedures for their

detection and quantification are well-known and relatively easy to perform. However,

care must be taken because they may not correlate with all DBP level (Villanueva, 2012).

Not all water treatment processes effect DBP precursors and DBPs similarly, for

example, THMs are volatile and may be removed by aeration while other DBPs would

not be.


13

Table 2-2: DBP Regulations and Guidelines

Disinfection By-product Ontario

(µg/L)

Canada

(µg/L)

California

(µg/L)

US EPA

(µg/L)

WHO

(µg/L)

China

(µg/L)

EU

(µg/L)

Singapore

(µg/L)

Total THMs (TCM,

TBM, BDCM, DBCM) 100 100 80 80

(e) (e) 100

(e)

Bromodichloromethane - - - 60 60 - 60

Bromoform - - - 100 100 - 100

Chloroform - - - 300 60 - 300

Dibromochloromethane - - - 100 100 - 100

HAA5 (MCAA,DCAA,

TCAA, MBAA, DBAA) (a)

80b 60 60 - - - -

Dichloroacetate

(dichloroacetic acid) - - - 50

c 50 - 50

Monochloroacetate

(monochloroacetic acid) - - - 20 - - 20

Trichloroacetate

(trichloroacetic acid) - - - 200 100 - 200

Dibromoacetonitrile - - - 70 - - 70

Dichloroacetonitrile - - - 20d - - 20

a A guideline value of 60 µg/L is currently under review.

b As low as reasonably achievable.

c Provisional guideline because disinfection is likely to create higher values.

d Provisional guideline because health database has uncertainties.

e The sum of the ratio of the concentration of each to its respective guideline value should not exceed 1.

“-“ Not applicable

Sources: ((Health Canada, 2012; European Commission, 1998; Government of Singapore, 2008; Ontario Ministry of the Environment, 2008; US

EPA, 2009; World Health Organization, 2011; Ministry of Environmental Protection of the Government of the People’s Republic of China, 2006;

State of California, 2013


14

2.2 DBP Precursors

DBP precursors include natural and anthropogenic organic matter, and inorganic halides

(bromide, iodide). Many of these precursors are innocuous until they react with a

disinfectant. Natural organic matter (NOM) is typically the main precursor of halo

organic disinfection by-products in drinking water (Valencia, 2014) and is found in all

surface waters. Groundwater typically contains significantly less NOM compared to

surface water, but may have more inorganic halides.

2.2.1 Natural Organic Matter

Natural organic matter is ubiquitous in surface water (Norwood and Christman, 1987)

forming a heterogeneous mix of humic substances, carboxylic acids, proteins, amino

acids, hydrocarbons, and polysaccharides (Liu et al, 2008b) formed from the degradation

and activity of plants, algae, microbes, etc (Stevenson, 1994). Its composition often

varies seasonally and between different source waters (Crittenden et al., 2005).

NOM can be classified in a variety of ways according to many analytical instruments

with different procedures for detection and quantification. In the current research project,

NOM was measured as dissolved organic carbon (DOC), UV254 absorbance, and with

liquid chromatography- organic carbon detection (LC-OCD).

Aromatic molecules, organic compounds with benzene rings, absorb UV254 light and can

be measured in surface water typically without significant interference. UV254 absorbance

is a measure of hydrophobic compounds with high apparent molecular weights and has

been correlated with the formation of DBP (Valencia et al., 2014) including THMs and

HAAs (Kim and Yu, 2005). SUVA is the specific UV254 absorbance and is calculated by

normalizing UV254 absorbance to DOC concentration.

Double bonds generally are not considered as reactive as the benzene ring but are more

reactive than saturated compounds because saturated compounds must undergo hydrogen

abstraction (Eqs 2.5 and 2.6) whereas double bonds need only undergo addition


15

(Crittenden et al., 2005). Double bonds in aliphatic compounds absorb light at a

wavelength of 280 nm (Valencia et al., 2014).

LC-OCD further classifies NOM into biopolymers (polysaccharides, proteins and amino

sugars), humic substances (humic and fulvic acids), building blocks (breakdown products

of humics), low molecular weight acids (monoprotic organic acids with mass less than

350 Dalton), and low molecular weight neutral (mono-oligosaccharides, alcohols,

aldehydes, and ketones).

Humic substances (including humic and fulvic acids) are the major fraction in NOM

found in most surface waters, have high molecular weights and carbon content, are

hydrophobic with both aliphatic and aromatic structures, and are rich in oxygen-

containing functional groups such as carboxyl, phenol, alcohol, and quinoid (Cho and

Choi, 2002). Fulvic acids are lower in molecular weight and higher in oxygen content

compared to humics acids. Humic substances are the major fraction of NOM responsible

for DBP formation and have been correlated to THM and HAA formation potential (Mori

et al., 2013; Liu et al., 2008a; Liu et al., 2008b; Cho and Choi, 2002; Zhang et al., 2008;

Kim and Yu, 2005). Aromatic, unsaturated aliphatic structures and electron donating

(activating) functional groups make humics very attractive to electronegative compounds

like chlorine (Kim and Yu, 2005; Liu et al., 2008a). However, studies show that other

fractions of NOM can also form THMs and HAAs, such as hydrophilic compounds (Liu

et al., 2008a; Liu et al., 2008b).

In the past decade new attention has been directed to nitrogenous DBPs (N-DBPs) such

as haloacetonitriles (HANs) and trihalonitromethanes (THNMs) formed from NOM

precursors containing nitrogen. Dissolved organic nitrogen (DON) compounds include

amino acids, proteins, and amino sugars and can be found in higher levels in surface

water sources influenced by wastewater effluent (Krasner et al., 2008; Pehlivanoglu-

Mentas and Sedlak, 2006). Various techniques are available to measure DON in water,

including the LC-OCD analysis. DON is mostly found in hydrophilic neutral and base

compounds which are poorly removed in traditional NOM precursor reduction


16

technologies (e.g. coagulation) making it likely to pass through a water treatment plant to

the disinfection steps (Mitch et al., 2009; Bond et al., 2011).

2.2.2 Anthropogenic Matter

Anthropogenic matter is typically a small fraction of the organic matter present in surface

water. It can originate from such sources as municipal sanitary and industrial wastewater

effluents, and runoff; and includes pesticides, pharmaceuticals and personal care products

(PPCPs), and textile dyes (Richardson, 2007; Krasner et al., 2008). Many of them have

activated aromatic rings that can readily react with chlorine (Richardson, 2007; Bond et

al., 2011). Synthetic organic matter, synthesized by human made processes, can be

recalcitrant in typical water treatment plants not designed to degrade or remove them.

2.2.3 Inorganic Halides

Naturally occuring inorganic halides; particularly bromide and iodide, can be oxidized to

form bromate (BrO3-), iodate (IO3

-), and hypobromous (HOBr/OBr

-) and hypoiodous acid

(HOI/OI-). Hypobromous and hypoiodous acid can then react with NOM to produce

bromo-, iodo- organic compounds (Crittenden, 2005; Liu et al., 2013; Selcuk et al., 2006)

by chlorine, ozone and AOPs (Crittenden et al., 2005). Bromide is naturally present in

many source waters (Plewa et al., 2004b) and iodide can occur naturally in coastal areas

experiencing salt water intrusion (Cancho et al., 2000; Richardson, 2003). Brominated

DPBs have been detected in finished drinking water for some time and the first iodo-

acids were reported during a United States of America nationwide DBP occurrence study

(Weinberg et al., 2002).

2.3 Current and Emerging DBP Control Strategies

DBPs from chlorination form when water contaminants are oxidized or otherwise

transformed. Technologies to reduce the concentration of NOM before disinfection and

alternative disinfection practices are effective DBP management strategies practiced in

the drinking water industry. Typical alternative disinfectants include chloramines,

chlorine dioxide (ClO2), ozone (O3), and low or medium pressure ultraviolet light (UV)


17

for primary disinfection; and chloramination and ClO2 for secondary disinfection. These

disinfectants have also shown to create disinfection by-products through oxidation or

transformation of water contaminants and can have adverse health effects (Richardson,

2003; Krasner et al., 1989; Bond et al., 2009; Buchanan et al., 2006; Lyon et al., 2012).

The class of chloramines includes monochloramine (NH2Cl), dichloramine (NHCl2), and

trichloramine (NCl3) however monochloramine is typically used for disinfection since the

other two can impart an unpleasant taste and odor to the water (Shorney-Darby and

Harms, 2010). Chloramines are formed by mixing chlorine with ammonia and the

higher the Cl:N ratio the more di- and tri- chloramines are formed (Crittenden et al.,

2005). Chloramines have been shown to create lower concentrations of DBPs associated

with chlorination, but produce other DBPs that have human health and regulatory

concerns (Krasner, 2006). For example, chloramines produce less THMs and HAAs, but

produce more N-nitrosodimethylamine (NDMA) (Crittenden et al., 2005). Also,

chloramines themselves have been linked to negative health effects in dialysis patients

and on aquatic organisms (Shorney-Darby and Harms, 2010).

Chlorine dioxide is a common disinfectant used in Europe (Hoehn et al., 2010). It has

been shown to create lower concentrations of DBPs associated with chlorination because

its chlorine atom does not react in electrophilic substitution reactions to form chlorinated

organic compounds like free chlorine (Aieta and Berg, 1986). It does produce other DBPs

that have human health and regulatory concerns, such as chlorate (ClO3-) and chlorite

(ClO2-) which are widely regulated (Krasner, 2006; Crittenden et al., 2005).

Ozone is a very strong oxidant that either oxidizes microbial and other organic and

inorganic constituents in water or decays very quickly. NOM that is not completely

oxidized by ozone to CO2 and minerals has been observed to be readily biodegradable,

affecting biological growth in distribution systems if not managed (US EPA, 1999).

Ozone also plays a significant role in bromate (BrO3-) and hypobromous acid

(HOBr/OBr-) formation the latter of which can further react to create brominated organic

compounds (Crittenden et al., 2005; Richardson, 2003). Similarly, hypoiodous acid

(HOI), iodate, and iodinated organic compounds may form from ozonation.


18

Ultraviolet disinfection has become popular for both municipal and small scale systems

worldwide. The germicidal range of the UV spectrum (100 – 400nm) is considered to be

200 to 300 nm (UVB and UVC), where nucleic acids and to a lesser extent proteins

absorb light and undergo chemical change. Microorganisms are inactivated when nucleic

acids undergo chemical change and prevent cell functioning and replication (Bolton,

2008). The two common UV lamps, low (LP) and medium (MP) pressure mercury vapor

lamps, emit light at 254 nm (82% of total emission) and 200 – 580 nm wavelengths,

respectively. The medium pressure mercury vapor lamp has a few peaks including one at

254 nm. A UV dose of approximately 40 mJ/cm2 is usually adequate for disinfection

although some microorganisms require up to 186 mJ/cm2 (Bolton, 2008). The UV dose is

the product of the UV fluence rate (intensity of light) and irradiation time.

Advanced oxidation processes (AOPs) utilizing the hydroxyl radical (HO*) including

UV/O3, UV/ hydrogen peroxide (UV/H2O2), and H2O2/O3 are typically incorporated into

a water treatment process train to degrade recalcitrant compounds such synthetic organic

compounds (SOCs) (Philippe et al., 2010). They use the same UV lamps as UV

disinfection however require a UV fluence (i.e. UV dose) of as much as 50 times that

required for UV disinfection and are not typically employed as alternative primary

disinfectants although they do provide concurrent disinfection upon application (Bolton,

2008). They are expected to produce DBPs similar to ozone. Other oxidants used in water

treatment, for example for taste, odor, color, iron or manganese reduction, may also

produce what are typically considered DBPs (Li et al., 2008).

TiO2 photocatalysis is actively being researched as a potential alternative disinfectant as

well (Gerrity et al., 2008) and some research into DBPs formed has been completed

(Richardson et al., 1996). However, because of the photocatalytic nature of the product,

its scale up into water treatment plants would place it as a primary disinfectant and a

secondary disinfectant would still be required.

Since chlorination, alternative disinfection processes, and other oxidation processes may

create by-products with adverse health effects, reducing their precursors and preventing

precursors from forming is another key DBP control strategy. Reducing DBP precursors


19

has the additional benefit of decreasing membrane fouling; biofilm grown in distribution

systems; and taste, color and odor issues.

The current best practice for DBP precursor reduction is enhanced coagulation, where

coagulation parameters (e.g. pH, coagulant dose) are adjusted to optimize NOM

reduction (Crittenden et al., 2005). Aluminum or iron salts are added to water and induce

dissolved NOM compounds to agglomerate together into flocs that can be removed

through clarification (e.g. sedimentation or flotation) or size exclusion (media or

membrane filtration). Enhanced coagulation is particularly effective at removing large

hydrophobic acids and less effective at removing hydrophilic NOM (Marhaba and

Pipada, 2000; Sohn et al., 2007) and dissolved organic nitrogen (DON) (Bond et al.,

2011).

Activated carbon, which adsorbs NOM, ion exchange, and ozone followed by bio

filtration are also sometimes used for NOM control (Crittenden et al., 2005). Biological

filters can remove assimilable organic carbon (AOC) DBP precursors (Richardson, 2003)

and the ozonation can degrade NOM to be more assimilable (Crittenden et al., 2005).

Emerging advanced oxidation processes (AOPs) incorporating UV/H2O2/O3 to produce

hydroxyl radicals have been shown to significantly decrease DBP formation by NOM

degradation; however these processes are energy and chemical intensive. TiO2/UV

catalyzes the production of hydroxyl radicals with the longest wavelength/lowest energy

light in the UV light spectrum (UVA), and is also effective at degrading DBP precursors

(Hadnadjev, 2010). It was the focus of the current research.

Membrane filtration, particularly nanofiltration and reverse osmosis, are capable or

removing organic and inorganic DBP precursors through pressure driven size exclusion

technology. Ultrafiltration removes high molecular weight dissolved organic matter such

as humic acids and biopolymers which have been identified as DBP precursors. These

filters typically become reversibly and irreversibly fouled with biopolymers, humics, and

other water constituents and have high operating/maintenance costs due to

backwashing/scouring and high pressure requirements (Crittenden et al., 2005).


20

Degradation of NOM under UV disinfection has been studied (photolysis), but results of

DBPs and DBP formation upon subsequent chlorination among different studies have

varied (Bond et al., 2009; Zhao et al., 2008; Buchanan et al., 2006; Lyon et al., 2012).

2.4 DBP Formation Potential Tests

A DBP formation potential test measures the formation of DBPs from a simulated

disinfection process and may be executed to determine DBP formation potential (fp) of

source water, the effects of pre-treatment processes, compliance with regulations, and the

effects of alternative disinfectants. Factors that affect DBP fp during disinfection include

type of disinfectant, time of reaction, pH, temperature, and disinfectant concentration.

These are controlled in a DBP formation test so that pre-treatments, source waters, etc.

can be compared.

The Uniform Formation Condition (UFC) chlorination test simulates typical secondary

disinfection in a distribution system (Summer et al., 1996). It employs a ph of 8.0 +/- 0.2,

reaction time of 24 +/- 1 hr, temperature of 20.0 +/- 1.0 oC, and 1.0 +/- 0.4 mg/L free

chlorine residual after 24 hr. Typically samples and free chlorine solution are buffered to

remain within the test parameters.

Other typical chlorination tests include the simulated distribution system (SDS) test and

the formation potential (FP) test. The SDS test simulates the conditions of a site-specific

distribution system that is being investigated. The FP test chlorinates water at high doses

(typically 3 – 5 mg/L at the end of the incubation period), for long incubation times

(typically 7 days), at typically 25 oC and 7 pH. This may lead to elevated DBP

concentrations and may be especially useful for detection DBPs that typically form in

low concentrations.

The UFC test as well may produce higher than expected DBP results because of the

chlorination parameters. Also, the chlorinated water DBP fp may not be representative of

a water supply if it uses raw source water that has not gone through the water treatment

plant. However the UFC chlorination test is useful to compare results with literature and


21

using raw source water may be an effective way to compare the fp of source water

between water treatment plants that might have different unit processes.

2.5 TiO2 Photocatalysis

TiO2 is one of a few semiconductor powders that has been studied for its ability to photo

catalytically degrade pollutants. TiO2 nanoparticle powders are common today in many

commercial products such as paints and coatings for self-cleaning walls, and other air and

water purifiers (Hadnadjev et al., 2010; Hurum et al., 2003). They are currently in the

experimental stage of research for incorporation into the drinking water treatment process

with potential to be a viable and economical process element for disinfection, and

reduction of recalcitrant or other contaminants and DBP precursors. The term

nanotechnology applies to a wide variety of products that have nano-meter sized

dimension(s) and have already entered the water treatment field with, for example,

nanofiltration for contaminant reduction by size exclusion (Crittenden, 2005; Narayan,

2010).

TiO2 photocatalysis (TiO2/UV) can decrease disinfection by-product formation potential

(DBP fp) in water by degrading precursor compounds before a disinfection process, by

providing an alternative disinfection process, or by degrading DBPs and precursors after

or during disinfection. TiO2 photocatalysis may also contribute to DBPs by forming DBP

during disinfection or by forming DBP precursors during a pre-disinfection stage.

Although the potential exists, similar to other alternative disinfectants that have been

used as a compliment/substitute to chlorination for disinfection, TiO2/UV may create its

own DBPs while decreasing the DBPs associated with chlorination, research to date has

shown that this is not a concern, and no DBPs have yet been definitively linked to

TiO2/UV (Richardson et al., 1996). TiO2 may itself be considered a DBP if TiO2

nanoparticles escape from a water treatment plant and the health effects of these

nanoparticles are not yet fully understood (Love, 2012). The current research focused on

the effects of TiO2 photocatalytic on NOM precursors of THMs, HAAs, HANs, HNMs

and HKs.


22

2.5.1 Mechanisms of Action

TiO2 nanomaterials degrade NOM through oxidation and reduction reactions catalyzed

by ultraviolet (UV) light (Valencia et al., 2014). They can either directly oxidize or

reduce NOM or produce reactive oxygen species (ROS) that can degrade NOM

(Hashimoto et al., 2005). TiO2 photocatalysis (TiO2/UV) produces hydroxyl radicals

(OH*) and can be classified as an innovative AOP that has the potential to be a low

energy intensive water treatment process requiring no chemical addition (Valencia et al.,

2014; Malato et al., 2009). TiO2 absorbs light and dissipates it through the excitation of

an electron from its valence band (VB) to its conduction band (CB), creating what is

termed electron/hole (e-/h

+) pairs (Cho and Choi, 2002; Nosaka and Nosaka, 2013).

TiO2 can be present in two phases, anatase and rutile. The anatase phase band gap of 3.22

eV corresponds to a UVA wavelength of 385 nm and the rutile phase band gap of 3.02

eV corresponds to a wavelength of 410 nm (Hurum, 2013) as per Eq. 2.8. The TiO2 will

absorb energy equal to or greater than the band gap energy (Ghaly et al., 2011). The

anatase phase has been shown to be more photoreactive than the rutile phase however a

mixture of the two phases may increase quantum efficiency (Fujishima and Zhang, 2006;

Hurum et al., 2003).

Where E = Band gap energy (eV)

= wavelengh ( )

(2.8)

UVA (400 – 315 nm) light makes up 3% of solar spectrum and is lower in energy

compared to the UVB (280 – 315 nm) and UVC (200 -280 nm) wavelengths required in

other AOPs and UV disinfection (Malato et al., 2009; Bolton and Cotton, 2008). The sun

has an irradiance of approximately 100 mW/cm2 (400 – 1100 nm) with 3 mW/cm

2

(standard global irradiance of UV light under clear skies in sunny countries) available for

TiO2/UV as a low-technology application and more possible with solar concentrators

such as the parabolic-trough reactor and compound parabolic concentrator (Malato et al.,

2009). Medium pressure lamps used in AOPs and UV disinfection have an output peak at


23

the 365 nm wavelength and both low and medium pressure lamps have an output peak at

254 nm. Both can be used for TiO2 photocatalysis because TiO2 will absorb wavelengths

with energy equal to or greater than its band gap (Bolton and Linden, 2003; Ghaly et al.,

2011). Other strictly UVA lamps can also be used.

The electrons and holes can directly reduce and oxidize NOM, respectively, create

reactive oxygen species (ROS) which can then degrade NOM, or recombine (Nosaka and

Nosaka, 2013). Eqs 2.9 to 2.15 and Figure 2-2 illustrate the formation and subsequent

reactions from these e-/h

+ pairs. ROS include the hydroxyl radical (HO

*), one of the most

powerful oxidants used today (Herrmann, 2010), and the super oxide (O2*-

) molecule.

(2.9)

(2.10)

(2.11)

(2.12)

(2.13)

(2.14)

(2.15)

Figure 2-2: TiO2 Photo Reactivity, Source: Crittenden et al., 2005


24

For anatase at a pH of 7, the hole in the e-/h

+ pair has a reduction potential of

approximately 2.9 V while the electron has a reduction potential of approximately -0.3 V

(Crittenden et al., 2005). Often, the e-/h

+ pairs recombine and produce heat and or light

causing TiO2/UV to have low quantum efficiency (Crittenden et al., 2005). Some e-/h

+

pairs migrate to the TiO2 surface where molecules with high electro-potentials such as

oxygen can attract the electron as shown in Eqs 2.10, 2.11, and 2.12 (Nosaka and

Nosaka, 2013; Hu et al., 2011). The holes can be filled through the oxidation of organic

matter, ions such as the hydroxyl radical, or water as shown in Eqs 2.13, 2.14, and 2.15

(Crittenden et al., 2005; Nosaka and Nosaka, 2013). The standard electrode potential for

the formation of HO* at a pH of 7 is approximately 2.177 V (Crittenden et al., 2005). It

has been determined that the hydroxyl radical quantum efficiency (hydroxyl radical

production/photons absorbed) is 4 % during TiO2/UV (Bolton, 2001). Many studies

suggest that the main reactive species in AOPs and TiO2 is the hydroxyl radical. The

second-order reaction rate constant for reactions with HO* radicals in aqueous solution

are typically in the order of 108 to 10

10 M

-1s

-1 and are 3 to 4 orders of magnitude higher

than other oxidants (Malato et al., 2009; Crittenden et al., 2005). The HO* radical can

oxidize NOM by electron transfer, hydrogen abstraction or HO* addition to double bonds

(Philippe et al., 2010; Bolton, 2001) as per Eqs. (2.16 – 2.18).

(2.16)

(2.17)

( ) (2.18)

The reactivity of e-/h

+ pairs and ROS make it such that degradation of constituents such

as NOM is enhanced when they are adsorbed to the TiO2 or is in close proximity to these

generated species. TiO2 has a negative surface charge above 7 (Ghaly et al., 2011), while

HAs are typically negatively charged above pH of 4. Therefore, there is a stronger

affinity between TiO2 and humic acid (HA) at pH between 4 and 7 (Liu et al., 2008).

Some studies look at using TiO2 as an absorbent because of the excellent adsorption

found under dark conditions, and then using photocatalysis to regenerate the absorbing

capacity (Liu et al., 2014). At an alkaline pH there is repulsion between HA and TiO2 as

both become negatively charged, however, increasing pH increases hydroxyl ion

concentration. This may increase the formation of HO* which may counterbalance at


25

least partially the lower affinity between TiO2 and HA (Liu et al., 2008a). It was also

speculated that at higher pH, large HA molecules uncoil due to deprotonation and the loss

of hydrogen bonding and an increase in repelling negative charges, leaving them more

susceptible to HO* attack. Although the overall charge on HA may be negative, HA

contains many functional groups which may interact with the TiO2 surface. pH can be

adjusted to optimize TiO2/UV (Valencia et al., 2014) and will depend on the chemical

characteristics of the contaminants of interest.

High alkalinity and ionic strength in water may decrease the effectiveness of TiO2/UV

because bicarbonate and carbonate, as well as phosphate and sulfate, will compete with

NOM and scavenge HO* radicals (Crittenden et al., 2005).

Hydrogen peroxide and oxygen addition may increase the effectiveness of ROS

production (HO* and O2

*-, respectively) and lead to enhanced degradation of

contaminants (Toor and Mohseni, 2006; Bond et al., 2009; Zhao et al., 2008). That is, the

kinetics of degradation may increase but recalcitrant compounds may nonetheless remain

(Liu et al., 2008a; Liu et al., 2008b). Also, because adsorption onto the TiO2 surface is

important for NOM degradation, high concentrations of H2O2 may compete with NOM

for adsorption sites (Liu et al., 2008a; Liu et al., 2008b).

2.5.2 Reaction Kinetics

The rate of degradation of NOM is a second order reaction dependent on the production

of electron/hole (e-/h

+) pairs and concentration of DBP precursors. TiO2 concentration

and UV dose are two dominant factors in e-/h

+ pair production. When the concentration

of TiO2 does not change within an experiment, it is typically incorporated into the rate

constant and a pseudo first order reaction rate is determined based on UV dose, where

UV dose (mJ/cm2) is the product of the UV fluence rate (mW/cm

2) and irradiation time

(seconds).

The production of e-/h

+ pairs can be approximated by measuring the production of HO

*

and a pseudo first order reaction rate can be determined based on HO*

production. This is

helpful to compare with other AOP processes that utilize the HO*. A pseudo first order


26

reaction rate can also be determined for the apparent rate of degradation of contaminants

although adsorption and degradation pathways may complicate calculations made from

experimental observations.

A pseudo first order reaction rate can be described by Eqs 2.19 to 2.24. Empirical

determination of reaction rates is common since the mechanisms of degradation are

complicated and not fully understood.

Reaction

Where R = NOM (DBP precursor) (2.19)

Reaction rate [ ] , units of M/s

Where k = reaction rate constant (1/s) (2.20)

Rate Law [ ]

[ ] (2.21)

Integrated Rate Law

[ ] [ ]

Where [R]0 is [R] at t = 0

(2.22)

Linear plot to determine k

([ ]

[ ] ) ,

Where slope of line = -k (2.23)

half-life ⁄

( )

(2.24)

The pseudo first order reaction rate constant should steadily increase with increasing

hydroxyl radical production, and the actual second order reaction rate constant can be

determined with the following equation:

[ ] (2.25)

In the above equation, a plot of k vs hydroxyl radical concentration will give the actual

second order rate constant. An increase in the concentration of TiO2 should result in a

directly proportional increase in hydroxyl radical production. A point is reached when the

TiO2 has a screening effect which dominates reaction kinetics by obscuring the UV dose.

Also, with higher TiO2 concentrations, particles may aggregate and decrease the number

of active surface sites. Typically, optimal TiO2 concentration has been reported in the

range 0.75 to 1 g/L (Ghaly et al., 2011; Liu et al., 2008b), however this will be dependent

on the type of catalyst, source water quality, and the UV fluence rate.

The hydroxyl radical production can be determined with a probe compound that has a

known second order degradation reaction rate constant based on its own and the hydroxyl


27

radical concentration. Various probes have been investigated for AOPs to determine

hydroxyl radical production including sucralose, methylene blue, and parachlorobenzoic

acid (Keen and Linden, 2013). Each has advantages and disadvantages in terms of factors

such as detection, interferences, detection limits, and photo stability.

The UV dose is a measure of the radiant flux (mJ/s) with time (s) typically normalized to

the surface area (cm2) of the water being irradiated. As the photocatalytic reaction

proceeds at a constant radiant flux with time, the reaction rates described above can be

determined. The radiant flux can also vary depending on the light source. Natural solar

light will typically have a lower intensity than industrially produced light sources for

municipal water treatment plants. Studies show that at low radiant fluxes, the reaction

rate increases linearly with radiant flux. Similarly as with TiO2 concentration however, a

plateau is reached. Here, it has been attributed to an excess of photogenerated species

with the mass transfer of contaminants to the photogenerated species limiting the reaction

rate (Malato et al., 2009).

Adsorption is an important factor in TiO2 photocatalysis, and reaction kinetics based on

available adsorption sites and adsorption rate has been determined to model the initial

first order degradation reaction rate with the Langmuir-Hinshelwood (L-H) adsorption

model. These values can be determined empirically and are dependent on pH, alkalinity,

and other general water quality parameters. The initial reaction rate would be:

[ ]

[ ] (2.26)

Where: = reaction rate constant based on fraction of surface covered by NOM

= fraction of surface covered by NOM

[ ] initial concentration of NOM

NOM adsorption constant

2.5.3 Configurations for TiO2 Photocatalysis

TiO2 photocatalysts are typically composed of anatase and rutile crystalline phases.

Degussa® (Evonik) P25, also known as Aeroxide® TiO2 P25 is an industry standard

TiO2 nanoparticle powder that is composed of anatase and rutile at a ratio of 70:30 or


28

80:20 which has been shown to give optimum photo activity (Ohtani et al., 2010; Ohno et

al., 2001). Advances in nanotechnology have brought about nanowires (cylindrical and

long), nanorods (cylindrical and short), and nanobelts (flat) which still have nanoscale

dimensions (e.g. diameter, width) but have increased lengths generally with a length to

width ratio of approximately 10:1 to 1,000:1 (Hu et al., 2011). These longer dimensions

may increase the quantum efficiency of TiO2 by encouraging e-/h

+ migration to the

surface. However, since photocatalysis occurs at or the surface of TiO2, a large surface

area is desirable making nanoparticles, P25 particularly, the most efficient TiO2 materials

yet fabricated (Hu et al., 2011). Metallic ions such as silver can be used to increase the

quantum yield of TiO2/UV by conducting electrons away from the TiO2 material (Li et

al., 2008) and silver particularly may impart some disinfection capability (Dobrovic et al.,

2012).

TiO2/UV requires UVA light for activation, but lower energy light may also produce e-/h

+

pairs with sufficient energy to create ROS and degrade contaminants. TiO2 can be doped

with different elements such as nitrogen, boron, iodine, and silver to sensitize it to lower

energy light in the visible range (Cho and Choi, 2002). Doping involves incorporating

elements into the TiO2 crystalline lattice. This occurs because the band gap between the

valence and conduction band decreases, either by an increase in energy of the valence

band or a decrease in energy of the conduction band. Not only would manufactured

lamps require less power input, but solar applications of TiO2/UV would be able to draw

on more of the available sunlight since visible light makes up approximately 40 to 50 %

of the sun’s spectrum (Malato et al., 2009).

Photoelectrocatalysis (PEC) is another technique that is used to increase the quantum

efficiency of TiO2. The catalyst is fixed on a conductive substrate and the photogenerated

electrons are driven to a cathode by an external voltage. For example, TiO2/Ti and

RuO2/Ti can be employed as the anode and cathode, respectively (Li et al., 2011). Recent

research shows that the overall differences between PEC and photocatalysis are not

significant (Egerton, 2011).


29

TiO2/UV reactors for TiO2 powder in suspension have frequently been set up as batch

reactors where the TiO2 material in added to a container of water, irradiated for some

time, and then the TiO2 separated from the treated water. In sequencing batch reactors

(SBRs), multiple batch reactors are applied in a non-concurrent way to achieve

continuous flow. SBRs are not typical in drinking water treatment facilities and more

commonly seen applied in wastewater treatment. The scaled step up of an SBR with

nanomaterial powders may be difficult for a large scale WTP in terms of product

recovery and required “downtime” for filling and decanting/filtering the TiO2 slurry.

However, for smaller, decentralized systems or as an emergency response plan for

disaster relief, a batch TiO2/UV reactor may be ideal. A continuously stirred tank reactor

(CSTR) is another common TiO2/UV configuration where the TiO2 slurry is circulated

within annular tubes irradiation from the inside out and filtration is used at the outlet to

remove TiO2 particles.

Immobilizing TiO2 powders on a support has been extensively researched and has gained

some significant ground, not just for water treatment but also air treatment and keeping

surfaces clean (Hadnadjev et al., 2010). Processes such as the sol gel method (Hatat-

Fraile et al., 2012) have been developed to coat materials. Through various deposition

methods, TiO2 is coated onto smooth surfaces and also porous surfaces to create filters.

There has also been some progress in the development of free standing TiO2 membranes

with nanowires (Hu et al., 2011). Filters, depending on their resulting pore size, can be

used for various types of filtration used in the water treatment industry including micro-,

ultra-, nano-, or reverse osmosis (0.1-10, 0.001-0.1, 0.0005-0.0015, and <0.001 µm,

respectively) membrane filtration. These membranes are typically made from organic

polymers (e.g. cellulose acetate, polycarbonate, and polysulfone) or ceramics. Ceramic

filters are particularly suited for coating with TiO2 by the sol gel method because a

covalent bond can be created between the TiO2 and the ceramic material. Ultrafiltration is

growing in popularity in municipal water treatment systems, and one group fabricated

titania-silica membranes with a 60 nm pore size using the sol-gel method (Mrowiec-

Biaon et al., 2004).


30

Membrane filters are used in drinking water treatment for the reduction of pathogens,

dissolved organic matter, and salts. With a TiO2 membrane, the pore size and chemical

characteristics of the surface will dictate treatment efficiency. The TiO2 may increase

reduction and provide degradation to prevent fouling. Particles that might otherwise have

been retained may pass through when degraded to smaller compounds with less affinity

for the TiO2 membrane surface. Irreversible fouling and reversible fouling can be

managed by a TiO2 membrane filter and is an attractive application of the technology

since backwashing of membranes to manage fouling is a considerable cost. Numerous

configurations have been proposed for TiO2 membrane filters (Liu et al., 2013).

2.5.4 Degradation of DBP Precursors

TiO2/UV can be used as an AOP to remove recalcitrant anthropogenic or natural organic

matter which persists through a conventional drinking water treatment process or as a

stand-alone pre-treatment followed by disinfection. TiO2 has been shown to degrade

DBPs and reduce DBP formation by altering or mineralizing precursors, similar to other

AOPs (Liu et al., 2008b). Research in advanced oxidation processes that rely on the

hydroxyl radical, including TiO2/UV, show that HO* will preferentially degrade

polysaccharides and humic components of NOM (Huang et al., 2008; Mori et al., 2013;

Liu et al., 2010), by their affinity for TiO2 and reactivity with it and ROS. These large,

aromatic, and functionalized compounds are oxidized to short chain aldehydes and

ketones (including acetaldehyde, n-propanal, and n-butanal), which are then oxidized to

carboxylic acids. A cyclic process ensues when the carboxylic acid is degraded to CO2

and a carbon-centered radical and the carbon-centered radical is then degraded to an

alcohol that is then oxidized to shorter chain ketones and aldehydes such as formaldehyde

and acetone. With sufficient time, the cycle repeats until the organic matter is mineralized

or recalcitrant compounds remain. Reaction intermediates or recalcitrant compounds can

then react with the disinfectant to form DBPs. These organic intermediates/recalcitrant

compounds can be considered direct TiO2/UV by-products. In a study by Liu et al.,

humics, polysaccharides, and building blocks as measured by the LC-OCD decreased

from TiO2/UV while low molecular weight acids and neutrals were recalcitrant fractions,

absorbing at wavelengths less than 230nm (Liu et al., 2010).


31

Reactive intermediate degradation products may originate from innocuous compounds

causing an increase in DBP fp from TiO2/UV (Richardson et al., 1996). Some researchers

have hypothesized that this occurs because the fragmentation of aromatic structures

within NOM exposes sites for chlorine attack and DBP formation. Given enough time

however, research shows that the amount of DBP precursors significantly drops and

remaining recalcitrant compounds are much less aromatic and less reactive with chlorine

(Phillipe et al., 2010; Richardson et al., 1996; Mori et al., 2013; Liu et al., 2010). Lui et

al. conducted experiments using two natural water sources in Australia with P25

nanoparticle powder at 0.1 g/L in suspension followed by chlorination (Liu et al., 2008a;

Liu et al., 2008b). THM fp and HAA5 fp and specific THM fp steadily decreased with

irradiation while specific HAA5 fp following the dark adsorption step and 30 min

irradiation was higher than only the dark adsorption step (Liu, 2008b). Longer irradiation

times brought about decreases in specific HAA5 fp. The intermediate degradation

products of TiO2 photocatalysis can be HAA precursors however recalcitrant compounds

are less so.

Kent et al. also found that THM fp and HAA fp upon chlorination decreased for samples

treated with TiO2/UV prior to chlorination (Kent et al., 2011). While these studies

correlated treatment time with DBP fp, Gerrity et al., correlated DBP fp with energy

input and found that extended treatments greater than 80 kWh/m3 dramatically decreased

THM fp (Gerrity et al., 2009).

Some compounds adsorb onto TiO2, degrade, and then de-absorb. DBP fp may also

increase when reactive intermediates are de-sorbing faster than reactive original or

intermediate NOM are adsorbing. The degradation of NOM can also make it easier or

harder to detect by analytical methods. These mass transfer and degradation pathway

characteristics of TiO2/UV make it difficult to determine precursor degradation kinetics.

DBP formation potentials based on concentration and % reduction are often used

however these too will depend on chlorination parameters.


32

The hydroxyl radicals formed through TiO2/UV can also oxidize bromide or iodine

similarly to ozonation and other AOPs to form bromo- and iodo- organic compounds like

bromo-THMs, HAAs or HANs (Liu et al., 2013).

It is difficult to compare TiO2/UV and other DBP precursor reduction experiments in

literature because methodologies vary and treatment efficiencies are site specific but

studies suggest that TiO2/UV has the potential be as effective or more effective. It is also

difficult to compare between different TiO2/UV experiments because of the differences in

light source, TiO2 concentration, TiO2 type, and reactor configuration, and how treatment

is reported, e.g. UV dose (mJ/cm2), energy (kWh/m

3), photon flux (E/s), irradiation time

(min). A few recent studies have compared experimental results from actual chlorinated

water supplies, including one which compared pre-treatment with ozonation, coagulation,

ozonation followed by coagulation, and TiO2/UV. The treatments were for the most part

comparable except for a few DBPs; trichloroacetic acid was higher in water treated with

TiO2/UV, and bromochloroacetonitrile was only found in water treated with TiO2/UV;

while chloral hydrate was significantly lower in water treated with TiO2/UV (Bekbolet et

al., 2005). Studies by Gerrity et al. and Phillippe et al. compared bench-scale enhanced

coagulation to TiO2/UV and found that photocatalysis is superior for THM fp reduction

(Gerrity et al., 2009; Phillipe et al., 2010). Table 2-3 compares TiO2 photocatalysis to

other common DBP precursor reduction technologies in terms of treatment effectiveness,

chemical and energy inputs, and other concerns.

Numerous configurations have been proposed for TiO2/UV; it can be used as a stand-

alone treatment prior to chlorination or in conjunction with other unit treatment

processes. TiO2/UV after coagulation has been shown to improve DOC and UV254

reduction compared to coagulation alone (Uyguner et al., 2007).TiO2 photocatalysis has

also been shown to increase the biodegradability of NOM, making it an ideal treatment

prior to biofiltration for NOM reduction (Phillipe et al., 2010; Liu et al., 2008a).


33

Table 2-3: Comparison of TiO2/UV and other DBP Precursor Reduction Technologies

Treatment DOC

Reduction

THM fp

Reduction

HAA fp

Reduction

Chemical

Inputs Energy Inputs Other Concerns

(Enhanced)

Coagulation 22-72%

a 37-84%

b 15% to 78% j

Iron or

aluminum salts

and pH

adjustment

Mixing Sludge production

Ozone 0-13%b

-36% to

29%b

80%k O3

Ozone generation and

diffusion Operator safety

(Ozone and)

Biofiltration 3-25%

b

-36% to

10%b

-11% to 28% j O3 Ozone generation and

diffusion

Operator safety, cold

weather operation

Nanofiltration 70-95%c

96-99%d

60-100%e

67% to 97% j Chemical

cleaning

Pumping against

transmembrane pressure Pretreatment required

TiO2/UV

75%f

60%g

94%h

90%i

96%f

53%g

95%i

75%f

May add O2,

H2O2

UV lamp, mixing,

and/or pumping against

transmembrane pressure

Long contact time required

a(Crittenden et al., 2005; Bekbolet et al., 2005; Krasner et al., 2012)

b(Krasner et al., 2012)

c(Amy et al., 1990; De la Rubia et al., 2008)

d(Itoh et al., 2001)

e(De la Rubia et al., 2008)

f120 minutes contact time, 0.1 g/L TiO2 (Liu et al., 2008b)

g10 minutes contact time, 0.1 g/L TiO2 (Phillipe et al., 2010)

hNanowire membrane (Zhang et al., 2008)

i320 kWh/m

3 UV dose, 0.4 g/L TiO2 (Gerrity et al., 2009)

j(Bond et al., 2009)

k (Jacangelo et al., 1989)


34

2.6 Research Gaps

The use of TiO2 photocatalysis is widely being researched for incorporation into water

treatment processes but the industry standard P25 which has the best performance record

still falls short as an economically viable technology. The time required for treatment and

product recovery are challenges that the technology faces. The current research studied

innovative materials to enhance the performance of TiO2, through visible light sensitivity,

increased quantum efficiency (reaction kinetics), and immobilization of the photocatalyst.

The body of knowledge of the effects of TiO2 photocatalysis on subsequent DBP

formation is in its infancy stages, with work having been done mostly with THMs and

HAAs; major DBP classes formed during chlorination from the reaction of NOM with

chlorine. The current research also examined the formation potential of HAN, HNM and

HK, DBP classes that are re-emerging as contaminants of interest and have not been

studied as extensively as THMs and HAAs.

The current work utilized a solar simulator as the light source, where many research

projects have typically used LP or MP UV mercury vapor lamps or UVA lamps.

TiO2/UV is an attractive technology for rural areas or as disaster relief where natural

energy sources may be the only ones available. Investigating the effectiveness of

TiO2/UV using only UVA-Vis found in natural sunlight is an excellent addition to the

current body of knowledge using manufactured lamps. Not only is this a viable low-tech

solution to drinking water treatment, but UVA and visible light are lower in energy than

the UVB and UVC light in LP and MP lamps.

This research also considered both short-term exposures (up to a few minutes)

representative of flow-through treatment systems, and longer term exposures of up to 30

min or more that may be more representative of batch reactor systems. The majority of

research projects to-date has focused on longer irradiation times (1 – 2 hours). The

irradiation times used in the current research were more representative of typical design

conditions for drinking water treatment systems. Many drinking water supplies in Ontario

rely on surface water (river, lake) as their raw source water. This thesis may assist

engineers in designing for future system upgrades, alterations, and new construction.


35

3 MATERIALS AND METHODS

The general structure of the experiments conducted in the current research is illustrated in

Figure 3-1. Raw water was mixed with TiO2 in dark conditions for NOM adsorption to

the TiO2 surface to occur and then the water/TiO2 slurry was irradiated. Aliquots of the

raw water, water following dark adsorption only, and water following dark adsorption

and irradiation were taken for NOM characterization. The degradation of DBP precursors

was studied by monitoring dissolved organic carbon (DOC), multi-wavelength and 254

nm UV absorbance, and changes in NOM fractions via liquid chromatography- organic

carbon detection (LC-OCD). Aliquots of the water samples were also taken and

chlorinated following the Uniform Formation Condition chlorination test. Aliquots from

these chlorinated samples were analyzed for trihalomethanes (THMs), haloacetic acids

(HAAs), haloacetonitriles (HANs), a trihalonitromethane (THNM), and haloketones

(HKs). The individual DBPs included trichloro-, bromodichloro-, dibromochloro-, and

tribromomethane; monochloro-, dichloro-, trichloro-, monobromo-, dibromo-,

bromochloro-, bromodichloro-, dibromochloro- and tribromoacetic acid; trichloro-,

dichloro-, bromochloro-, and dibromoacetonitrile; trichloronitromethane; and 1,1-

dichloro-2- and 1,1,1-trichloropropanone. The disinfection by-products were measured

by gas chromatography with electron capture detection.

Figure 3-1: General Schematic of Experiments


36

3.1 Materials

The apparatus and reagents used in the experiments are listed in Table 3-1 and Table 3-2,

respectively.

Table 3-1: Apparatus

Apparatus Supplier, Product Number Use

Solar Simulator Photo Emission Tech. Inc. (Camarillo, CA,

USA), SS150AAA TiO2/UV

Fritsch Ultrasonic

Cleaner Laborette 17 Laval Lab Inc. (Laval, QC) TiO2/UV

2511B-75 Aspiration

Pump 115 V Fisher Scientific (Ottawa, ON), 0105510 TiO2/UV

0.45 µm x 47mm Supor®

PES membrane filters

VWR (Mississauga, ON), CA28147-640

and CA28147-468 TiO2/UV

TiO2 membrane support

for batch test University of Waterloo (Waterloo, ON) TiO2/UV

Analytical balance

(+/- 1 mg) Ohaus (Florham Park, NJ), AP210

TiO2/UV, synthetic

water prep, DOC

analysis,

1 L amber bottles Cole-Parmer (Montreal, QC), RK-34607-40 Synthetic water prep,

DOC analysis

TOC analyzer

OI Analytical (College Station, TX, USA),

Aurora model 1030 with autosampler

model 1088

DOC analysis

1 cm quartz cuvette Agilent Technologies (Mississauga, ON),

5061-3387 UV254 analysis

Hewlett Packard 8452A

Diode Array UV

spectrophotometer

Agilent Technologies (Mississauga, ON), UV254 analysis

LC-OCD University of Waterloo, Waterloo, ON NOM characterization

500 mL or 250 mL amber

bottles

Cole-Parmer (Montreal, QC), RK-99540-32

or RK-99540-31

TiO2 treated water

storage

125 mL amber bottles Cole-Parmer (Montreal, QC), 99535-30 UFC chlorination test

DR/2500

Spectrophotometer

Hach Company (Mississauga, ON),

5900000 UFC chlorination test

Thermo Scientific, Orion

Star A111 pH meter Cole-Parmer (Montreal, QC), RK-58825-04 UFC chlorination test

Incubator Precision Scientific, Model 805 UFC chlorination test

1.8mL amber glass GC

vials and caps with septa

Chromatographic Specialties Inc.,

C58002W

THM, HAA, HAN,

THNM, HK analysis

Hewlett Packard 5890

Series II GC-ECD Agilent Technologies (Mississauga, ON)

THM, HAA, HAN,

THNM, HK analysis

DB 5.625 capillary

column

Agilent Technologies (Mississauga, ON),

1225631

THM, HAA, HAN,

THNM, HK analysis

Diazomethane generator Sigma Aldrich (Oakville, ON), Z411736 HAA analysis

PTFE-faced silicone

septum Sigma Aldrich (Oakville, ON), Z411760 HAA analysis


37

Table 3-2: Reagents

Reagent, Purity Supplier, Product Number Use

Milli-Q® Water Ultrapure water prepared in

laboratory all procedures

Aeroxide® TiO2 P25 powder,

99.5%

Sigma Aldrich (Oakville, ON),

718467-100G TiO2/UV

TiO2 innovative nanostructured

powders

University of Waterloo (Waterloo,

ON)

TiO2/UV

Otonabee River water Peterborough Utilities Services Inc. TiO2/UV

Synthetic river water Prepared in laboratory TiO2/UV

Deionized water Prepared in Laboratory

TiO2/UV, synthetic

water prep, UFC

chlorination test,

Alginic Acid Sigma-Aldrich (Oakville, ON),

A7003-100G Synthetic water prep

Suwannee River NOM RO isolation, (International Humic

Substances Society, 2R101N) Synthetic water prep

Compressed nitrogen gas N2,

Ultra High Purity Praxair (Mississauga, ON), DOC analysis

200 µg/mL trihalomethanes in

methanol, 96.5 – 99.9%


48746 THM Analysis

2000 µg/mL EPA 551B

Halogenated Volatiles Mix,

89.9 – 99.9 %


48046

HAN, THNM, HK

Analysis

1,2-dibromopropane (10,000 µg/L

in hexane),

Ultra Scientific (Kingstown, RI,

USA), PPS-400

THM, HAN,

THNM, HK

Analysis

P5, 5% methane, 95% argon

(Ultrapure) Praxair (Mississauga, ON)

THM, HAA, HAN,

THNM, HK analysis

Helium, Ultra High Purity Praxair (Mississauga, ON) THM, HAA, HAN,

THNM, HK analysis

2000 µg/L Haloacetic acid in

MTBE, 96.0 – 99.9%


49107-U HAA analysis

N-methyl-N-nitroso-p-toluene

sulphonamide (Diazald)

[CH3C6H4SO2N(CH3)NO], 99%

Spectrum (New Brunswick, NJ,

USA), M2272-25G HAA analysis

2,3,4,5-tetrafluorobenzoic acid,

99%


326267 HAA analysis

The SS150AAA Solar Simulator from Photo Emission Tech., Inc. was the light source in

the experiments and is pictured in Figure 3-2. Its electromagnetic radiation spectrum

shown in Figure 3-3 matched the natural solar radiation spectrum at approximately 108

mW/cm2 “one sun” light intensity (300 – 1100 nm) (Chawla, 2014), of which

approximately 13.4 mW/cm2 was available for TiO2/UV. Outlined in Figure 3-3 is the


38

UV-Vis photoactive range (300-424 nm) of the TiO2 nanomaterials studied. At one sun

intensity, the solar simulator UVA fluence rate was slightly higher than the UVA fluence

rate (3 mW/cm2) of the sun (Malato et al., 2009; Chawla, 2014). It did not have

appreciable output below the 300 nm (Chawla, 2014).

Figure 3-2: Solar Simulator

Figure 3-3: Spectral Radiation of Solar Simulator

The innovative TiO2 nanomaterials were fabricated by project partners at the University

of Waterloo. They provided characterization information, including scanning electron

microscope (SEM) images and band gap data provided in Figure 3-4 and Table 3-3,

respectively. The anatase, anatase doped with nitrogen at 5% mass and anatase doped

with boron at 5% mass were fabricated using the sol-gel method (Hatat-Fraile et al.,

2013; Mendret et al., 2013). Nitrogen doping and boron doping was accomplished using

urea and boric acid as nitrogen and boron sources, respectively following procedures

described by Azouani, 2009. Nitrogen and boron doped products were fabricated in order

to sensitize anatase to visible light. As shown in Table 3-3, the band gap of nitrogen was

0

0.1

0.2

0.3

0.4

0.5

0.6

300 500 700 900 1100

Inte

nsi

ty (

mW

/cm

2)

Wavelength (nm)

Solar Simulator Spectral Distribution

424


39

lower than anatase however the band gap of boron was the same as anatase (Hatat-Fraile,

2014; Liang, 2014). Nanobelts (NB) were fabricated using a modified technique

described by Liang et al., 2013a and 2013b. Nanobelts were fabricated from P25 however

the manufacturing process created nanobelts mostly composed of less photoactive rutile

phase (Liang, 2014). The Ag@SiO2@TiO2/P25 product was a 1:99 mixture of

Ag@SiO2@TiO2 triplex core-shell photocatalyst and P25 (Liang, 2014). The

Ag@SiO2@TiO2 was fabricated using modified technique described by Zhang et al.,

2013 and is composed of a silver core coated with silica which is then coated with TiO2.

The silica coating prevents ionization of silver and subsequent loss to the water matrix.

The silver facilitates electron transport and sensitizes the TiO2 to visible light (Zhang et

al., 2013; Liang, 2014). Characterization and methods of synthesis for the nanobelts and

P25 thin film in preliminary experiments were not provided.

Figure 3-4: As prepared (a) NB, (b) Ag@SiO2@TiO2 (c) 1% Ag@SiO2@TiO2/P25,

(d) Anatase, (e) Anatase-N, and (f) Anatase-B


40

Table 3-3: Characteristics of TiO2 Materials

TiO2 material Band gap

(eV)

Wavelength

(nm)

Solar Simulator UV Fluence Rate

at Water Surface for =< band

gap (mW/cm2)

P25 3.03 409 9.50

Nanobelts 2.95 420 11.5

Ag@SiO2@TiO2/P25 3.03 409 9.50

Anatase 3.2 387 5.31

Anatase-N 3.15 394 6.60

Anatase-B 3.2 387 5.31

TiO2 photocatalytic experiments were conducted with model and real river waters. Model

river water (‘synthetic water’) was produced in the lab while the natural surface water

was sourced from the Otonabee River (‘Otonabee water’). Otonabee water was obtained

from the Peterborough Utilities Commission’s (PUC’s) Peterborough water treatment

plant (WTP) in Peterborough, Ontario and was couriered to the laboratory and kept

refrigerated at 4 oC until use. The water supplied by the PUC for the preliminary

experiments was obtained from the Peterborough WTP intake during multiple sampling

events from January to April 2014 when there was no prechlorination for zebra mussel

control. In June 2014 when the TiO2 experiments testing the innovative TiO2 materials

were conducted, the PUC was pre-chlorinating water at the water intake for zebra mussel

control so water was obtained from the river bank, and filtered with 0.45 µm PES

membrane filters to remove particulate matter. An image of the residue is provided in

Figure 3-5. It was homogenized and stored in one carboy to reduce variability in raw

water quality between experiments. The Otonabee water characteristics are given in

Table 3-4.

Also provided in Table 3-4 are the synthetic water characteristics. To prepare the

synthetic water, three stock solutions were first prepared; a salt mixture, a calcium

sulphate solution, and a humic and Alginic acid mixture. In a 1 L volumetric flask,

0.2264 g of calcium chloride dehydrate (CaCl2*2H2O, ≥99.0%), 0.8356 g of magnesium

chloride hexahydrate (MgCl2*6H2O, 99.0-101.0%), and 0.0412 g of sodium nitrate

(NaNO3, ≥99.0%) were dissolved in deionized (DI) water (Milli-Q® for preliminary

experiments). The salt mixture was stored at room temperature. In a 1 L volumetric flask,


41

0.2954 g of calcium sulphate dihyrate (CaSO4*2H2O, 98%) was dissolved and also stored

at room temperature. The stock humic and alginic acid mixture was prepared by

dissolving 0.0256 g of Suwannee River NOM and 0.0532 g of alginic acid into a mixture

of 75 mL DI water and 0.25 mL of freshly prepared 1.00 M sodium hydroxide (NaOH,

97.0%), and then transferring to a 100 mL volumetric flask and diluting to 100 mL. This

stock was stored at 4 oC. Stock solutions were typically stored for no longer than 2

weeks. To prepare the synthetic water, 100 mL of the salt mixture, 333 mL of the calcium

sulphate mixture, and 10 mL of the humic and alginic acid mixture, and 0.126 g of

sodium bicarbonate (NaHCO3, 99.5-100.5%) were mixed in a 1 L volumetric flask. The

synthetic water was stored at 4 oC for typically 1 week.

Figure 3-5: Residue from Otonabee Water Filtration


42

Table 3-4: Synthetic and Otonabee River Water Characteristics

Water

source

Ca2+

mg/L

Mg2+

mg/L

Na+

mg/L

Cl-

mg/L

NO3-

mg/L

Br-

mg/L

SO42-

mg/L

Hardness

(as CaCO3)

mg/L

(CO3)TOT

mg/L

Alkalinity

(as CaCO3)

mg/L

NOM

mg/L

Alginic

Acid

mg/L

DOC

mg/L pH

Synthetic

water 29.1

a 9.99

a 36.2

a 40.0

a 3.00

a N/A 55.0

a 114

a 90

a 117

b 2.56

a 5.32

a 3

c 8.2

a

Otonabee

water - - 6.46

d - - <0.011

e - 88.0

f - 84

b - - 5.6

f 8.0

d

a Synthetic water recipe

b Measured in laboratory with aliquot of raw water sample

c Approximate average concentration measured with TOC analyzer during experiments

d(PUC, 2013)

e (Woodbeck, 2007)

f Peterborough WTP influent sourced from Otonabee water (City of Peterborough Environmental Protection Laboratory, 2014)

“N/A” Not applicable, bromide was not added to the synthetic water

“-“ Not available


43

3.2 Experimental Protocols

3.2.1 TiO2 Photocatalytic Procedures

Bench scale batch experiments included a series of operational steps as shown in Figure

3-6, where TiO2 was added to water, the slurry irradiated, and then the TiO2 recovered.

This process might represent a full scale system where water is treated in batches or a

sequencing batch reactor (SBR) where multiple batch reactors are applied in a non-

concurrent way for continuous flow of treated water. The light intensity of the solar

simulator was set to 100 mW/cm2 “one sun” intensity (400 – 1100 nm) at the water

surface, with approximately 13.4 mW/cm2 (300 – 424 nm) available for TiO2/UV. Spatial

variation was calibrated to below 2 % by measuring intensity within the irradiation field

with a photo detector and multi volt meter and adjusting the lamp orientation in the x, y, z

axes accordingly (Photo Emission Tech. Inc., 2012).

IDLE

FILL

REACT

FILTER

Figure 3-6: Batch Experimental Set-up

Each experiment that was tested for disinfection by-product formation potential post

chlorination was replicated 4 four times; each as independent experiments. Either one or

two replicates were used for chlorine demand tests and the remaining two or three


44

replicates were used to determine DBP fp. All replicates were analyzed for UV254 and

DOC.

Six proof-of-concept experiments were initially conducted to observe TiO2/UV

degradation of DBP precursors. Treatments included 30 min of dark adsorption followed

by 0, 30, 60, 90 and 120 min of irradiation. The 30 min of dark adsorption was

determined to be the maximum adsorption of DOC and UV254 with P25 in a continuously

stirred suspension of synthetic water at a concentration of 0.5 g/L in experiments by other

project participants at DWRG. The 30 min maximum adsorption was also observed for

experiments by project participants at the University of Waterloo. The experimental

conditions of these six proof-of-concept experiments are provided in Table 3-5.

Table 3-5: Preliminary Proof-of-Concept Experiments

No. Source water TiO2 material TiO2 concentration

(g/L) TiO2 configuration

1 Synthetic water P25 0.5 suspension

2 Otonabee River P25 0.5 suspension


4 Otonabee River P25 0.15 thin film

5 Otonabee River nanobelts 0.15 suspension


TiO2 powder was measured using the analytical balance and added to 200 mL of water in

a 250 mL beaker to give the desired TiO2 concentration. Four stir plates were placed

within the solar simulator’s illumination field and the irradiance measured at the water

surface to confirm one-sun intensity. Some of the beakers were outside the solar

simulator’s calibrated illumination field by approximately 1 cm however the irradiance

was checked and was within the allowed 2 % variability. A stir bar was placed inside

each beaker and the speed setting was set to medium. The rotations per minute (RPM)

was not measured, however, consistency between samples was achieved by using the

same setting and ensuring a vortex with depth of about 1 cm was maintained. The

samples were left in the dark for 30 min, and then the solar simulator shutter was opened

for illumination to occur. Following treatment, the samples were filtered through 0.45 µm

hydrophilic polyethersulfone (PES) membrane filters to remove particulate matter


45

(including TiO2) and stored in the fridge at 4 oC in appropriate containers for DOC and

LC-OCD (40 mL glass vials) and UV254 and DBP fp (250 mL or larger glass amber

bottles) analysis typically within 7 days. Raw water was also filtered and stored similar to

the treated samples to be analyzed as a control. DOC and UV254 were measured for all

four replicates. Two of the replicates were used to determine chlorine demand of the

water, and two were used to determine DBP fp.

A series of preliminary tests with synthetic river water ‘synthetic water’ and P25 powder

were completed to observe the effects of TiO2 concentration and different dark adsorption

times. The TiO2/UV protocol for these experiments was generally the same as the proof-

of-concept experiments described above. In these experiments however, a stock slurry of

TiO2 was made at a concentration of 10 g/L and then added to the water samples. The

stock TiO2 slurry for the dark adsorption samples was additionally sonicated for 5 min to

ensure distribution of the TiO2 nanomaterial throughout the sample and prevent

clumping. Details of these experiments are provided in Table 3-6.

Table 3-6: Preliminary Optimization Experiments

Experiment TiO2 concentration

(g/L)

Dark Adsorption

(min)

Irradiation Time

(min)

TiO2

concentration

0.005, 0.05, 0.1, 0.2,

0.5 10

0, 0.5, 1, 2, 5, 10, 15,

30, 60

Dark

Adsorption 0.1 0, 1, 2, 5, 10 0, 1, 30

The TiO2 concentration experiment included many irradiation times in order to determine

pseudo first order reaction rates and optimize the efficiency of the treatment process. The

dark adsorption experiment irradiation times (1 and 30 min) were chosen to represent

what might typically be expected as treatment times for a TiO2 membrane filtration

system or large-scale municipal WTP and small-scale, low-tech, decentralized WTP,

respectively.

Following the preliminary experiments, a set of experiments were conducted to test

industry standard Aeroxide® P25 and newly developed TiO2 nanomaterials with

synthetic and Otonabee River water ‘synthetic water’ and ‘Otonabee water’, respectively.


46

The protocol from the preliminary experiments was modified for these experiments. A

TiO2 working stock solution of 5 g/L was prepared by measuring TiO2 powder into Milli-

Q® water and sonicating for 5 minutes, followed by stirring with stir bar and stir plate for

the duration of the experiments for which it was used. A 4 mL aliquot of the stock

solution was added to 196 mL of water in a 250mL beaker using an Eppendorf pipette to

create a 0.1 g/L working TiO2 concentration. A one minute dark adsorption time was

used, followed by 0, 1, 15 and 30 min of irradiation. One of the replicate was used to

determine chlorine demand, and three replicates were analyzed for DBP fp.

3.2.2 UFC Chlorination Test

The disinfection by-product formation potential of water treated with TiO2/UV was

determined generally following the methodology of the Uniform Formation Condition

chlorination test (Summers et al., 2013). However, samples and chlorine solution were

not buffered, since the water pH was approximately 8 for both the synthetic and

Otonabee waters. The UFC conditions are: contact time of 24 +/- 1 hr, temperature of

20.0 +/- 1.0 oC, pH of 8.0 +/- 0.2, and a free chlorine residual of 1.0 +/- 0.4 mg/L after 24

hr.

Chlorine demand-free 125 mL amber bottles were prepared by filling bottles with

deionized water and one Pasteur pipette of liquid Javex® bleach. Bottles were left for at

least one hr and typically 24 hr. Before use, the bottles were emptied and rinsed with

deionized water twice and distilled water once. A working solution of chlorine was

prepared by adding 1 mL of sodium hypochlorite (NaOCl, 10-15 %) stock solution into

99 mL of Milli-Q® in a 125 mL amber bottle, and stored in the fridge. The concentration

of the chlorine working solution was measured before each chlorine demand and UFC

test with a Hach kit three times and averaged to ensure an accurate value was obtained

and used for dosing samples. One or two 125 mL bottles were collected from each

sample type and the pH was measured. Bottles were dosed (at different concentration

levels if two were used) based on anticipated chlorine demand to obtain a 24 hr chlorine

residual of 1 mg/L and the time of dosing was noted. The free chlorine concentration and

pH were measured and then the samples, headspace free, were placed in an incubator at


47

20 oC for 24 hr. After 24 hr the chlorine residual, temperature, and pH were measured.

The required chlorine dose to obtain a 1 mg/L residual was determined. This amount was

then spiked into the remaining two or three replicates, and the procedure to determine

chlorine demand was followed, except that chlorine concentration was not measured

immediately following the spike. Also, after the 24 hr contact time and pH and chlorine

residual measurements, samples were quenched with L-ascorbic acid. Samples were then

immediately analyzed for DBPs, or stored in the fridge in 40 mL amber vials headspace

free for THM fp and HAN fp analysis and 125 mL amber bottles for HAA fp analysis.

3.3 Analytical Methods

3.3.1 Water Quality Parameters

The pH and temperature of water samples were measured during the chlorine demand

and UFC chlorination tests. The pH meter calibration was checked with standard pH

buffer solutions of 4, 7 and 10 prior to each use, and re-calibrated when the pH deviated

by +/-0.2 from the buffer. Samples were continuously stirred during pH measurement

with a magnetic stir bar and plate. Temperature was measured with a standard non-

mercury glass thermometer.

3.3.2 Chlorine Residual

Chlorine concentrations were determined following the DPD colorimetric Standard

Method 4500-Cl G, (APHA, AWWA, WEF, 2012). A Hach Kit was used and blanked

with the sample prior to testing for chlorine concentration. Samples with a chlorine

concentration greater than 2 mg/L were diluted with Milli-Q® to 0 to 2 mg/L to fall

within the Hach Kit range.

3.3.3 Trihalomethane, Haloacetonitrile, Halonitromethane and Haloketone Analysis

The analysis of trihalomethanes (chloroform or trichloromethane (TCM),

bromodichloromethane (BDCM), chlorodibromomethane (CDBM), bromoform or

tribromomethane (TBM)), haloacetonitriles (bromochloroacetonitrile (BCAN),


48

dibromoacetonitrile (DBAN), dichloroacetonitrile (DCAN), trichloroacetonitrile

(TCAN)), chloropicrin (CP), and haloketones (1,1-dichloro-2-propanone (DCP) and

1,1,1-trichloropropanone (TCP)) was performed according to Standard Method 6232B

(APHA, AWWA, WEF, 2012). These compounds were extracted from the environmental

water samples in one extraction with methyl-tert-butyl-ether (MTBE, ≥99.0%) and then

their concentration determined via gas chromatography with electron capture detection

(GC-ECD). Instrumentation details and operation conditions are provided in Table 3-7.

All preparative steps were performed in a fume hood wearing appropriate person

protection equipment.

Table 3-7: THM and HAN GC-ECD Instrumentation and Operating Conditions

Parameter Description

Model Hewlett Packard 5890 Series II Plus GC-ECD

Column DB 5.625 capillary column

Injector Temperature 200oC

Detector Temperature 300oC

Temperature Program

40 oC for 4.0 min

4 oC/min temperature ramp to 95

oC


oC

Carrier Gas helium

Flow Rate 1.2 mL/min at 35 oC

Makeup Gas 5% methane, 95% argon 23.1 mL/min

Samples were quenched of any chlorine residual with L-ascorbic acid and stored at 4 oC

for up to 14 days in 40 mL vials headspace free prior to THM and HAN analysis. A 20

mg/L working solution of THMs was prepared by diluting concentrated stock solution

(200 µg/mL THM in methanol, 96.5 to 99.9 %) in methanol (, ≥99.9 %). A 20 mg/L

working solution of the halogenated volatile mix (HAN, CP, and HK) was prepared by

diluting concentrated stock solution (2000 µg/mL EPA 551B Halogenated Volatiles Mix,

89.9 to 99.9 %) in acetone (, ≥99.9 %). The internal standard (1,2-dibromopropane (1,2-

DBP)) working solution was prepared by diluting concentrated stock solution (10,000

µg/L in hexane) in MTBE to 100 mg/L. Approximately 30 min prior to use,

environmental samples were taken from the refrigerator and DBP and internal standard

working solutions were taken from the freezer to bring them to room temperature.


49

Calibration curves were prepared using 7 standards at a range of concentration from 2 to

140 µg/L for THM and 2 to 64 µg/L for the halogenated volatile mix. Method detection

limits (MDLs) were determined with the preparation of 9 samples at 2 µg/L for both

THMs and the halogenated volatile mix. Quality control charts were prepared for running

check standards with 9 samples at 40 µg/L for THMs and 10 µg/L for the halogenated

volatile mix. Running check standards were prepared alongside environmental samples

at a frequency of 1 per 10.

A 25 mL aliquot of each sample was transferred into a clean 40 mL vial and 50

(preliminary tests) or 20 µL (other tests) of 1,2-DBP was added as an internal standard.

One tsp of sodium sulphate (Na2SO4, ACS Grade) was added to the samples to increase

extraction efficiency. Then, 4 mL of MTBE were added to extract DBPs using a bottle-

top dispenser and samples caped with Teflon®-lined silicon septa and screw cap.

Samples were shaken vigorously for approximately 30 s and placed on their sides while

the process was repeated for all samples. Then the samples were shaken for 2 min and

left upright in a rake for 60 min for phase separation. Without disturbing the water layer,

approximately 1.8 mL of the MTBE layer was removed with a Pasteur pipette and placed

in a 1.8 mL GC vial. Samples were then analyzed using the GC-ECD.

3.3.4 Haloacetic Acid Analysis

Haloacetic acids (monochloroacetic acid (MCAA), dichloroacetic acid (DCAA),

trichloroacetic acid (TCAA), bromochloroacetic acid (BCAA), bromodichloroacetic acid

(BDCAA), dibromochloroacetic acid (DBCAA), monobromoacetic acid (MBAA),

dibromoacetic acid (DBAA), and tribromoacetic acid (TBAA)) were analyzed following

the micro liquid-liquid extraction, gas chromatography separation with electron capture

detection method found in Standard Method 6251 B found in Standard Methods (APHA,

AWWA, WEF, 2012). GC instrumentation and operating conditions are listed in Table

3-8. All procedures were completed within a fume hood wearing appropriate person

protection equipment.

Samples were quenched of any chlorine residual with L-ascorbic acid and stored at 4 oC

for up to 14 days in 125 mL or larger amber bottles prior to HAA analysis. A 20 mg/L


50

working solutions of HAA was prepared by diluting concentrated stock solution (2000

µg/L HAA in MTBE, 96.0 to 99.9 %) in MTBE (≥99.0 %). A working solution of

internal standard 2,3,4,5-tetrafluorobenzoic acid (TFBA) was prepared by dissolving

TFBA powder (99 %) in MTBE to a concentration of 2000 mg/L. Approximately 30 min

prior to use, environmental samples were taken from the fridge and HAA and internal

standard working solutions were taken from the freezer to bring them to room

temperature. An HAA calibration curve was prepared using 7 standards at a range of

concentration from 2 to 60 µg/L. MDL were determined with the preparation of 9

samples at 2 µg/L. Quality control charts were prepared for running check standards with

9 samples at 10 µg/L. Running check standards were prepared alongside environmental

samples at a frequency of 1 per 10.

Table 3-8: HAA GC-ECD Instrumentation and Operating Conditions

Parameter Description

Model Hewlett Packard 5890 Series II Plus GC-ECD

Column DB 5.625 capillary column

Injector Temperature 200 oC

Detector Temperature 300 oC

Temperature Program

37 oC for 10.0 min

2.5 oC/min temperature ramp to 65

oC


oC


oC

205 oC hold for 7 min

Carrier Gas helium

Flow Rate 1.2 mL/min at 35 oC

Makeup gas 5% methane, 95% argon 23.1 mL/min

A 20 mL aliquot of sample was transferred into a clean 40 mL vial and 20 µL of TFBA

internal standard was added. Then, one tsp of oven dried sodium sulphate (Na2SO4, ACS

Grade) was added to increase extraction efficiency from the water phase. Three mL of

concentrated sulphuric acid (H2SO4, >98 %) was added to ensure HAAs remained

protonated; which also assisted in the extraction from water phase. Five mL of MTBE

was added and then the samples were shaken by hand for 4 min. Samples were left for 60

min for phase separation and then 1.5 mL of the MTBE layer was extracted with an

Eppendorf pipette into a 1.8 mL GC vial. Extracts were cooled in the freezer for at least 7


51

min, and then 150 µL of diazomethane collected in MTBE was added and samples were

left for derivatization for at least 48 hr before analysis by GC-ECD.

Diazomethane was produced with the MNNG diazomethane generation apparatus. The

apparatus was set in an ice water bath. MTBE (2.6 mL) was added to the outer tube of the

generator and approximately half an inch of Diazald (N-methyl-N-nitroso-p-

toluenesulfonamide, CH3C6H4SO2N(CH3)NO, 99 %) was added to the inner tube. The

Diazald was covered by approximately 0.5 mL of methanol (≥99.9 %) and the set-up

allowed to cool for 10 min. A gas tight syringe was used to slowly add 600 µL of 20%

sodium hydroxide solution (prepared by diluting sodium hydroxide NaOH, 97.0% in

Milli-Q®) to the inner tube. Diazomethane was allowed to form for about 45 minutes in

the ice bath and then was transferred to an amber vial for storage of up to two weeks in

the freezer.

3.3.5 Natural Organic Matter (DOC, UV254, LC-OCD)

All water samples were filtered with Supor® 0.45 µm PES membrane filters following

TiO2 photocatalysis to remove suspended organic matter and TiO2. Samples were stored

for a maximum of 7 days at 4 oC prior to DOC and UV254 analysis in amber bottles to

avoid exposure to light and atmosphere. SUVA was calculated as the UV254 (absorbance

of UV with wavelength of 254 nm) normalized to DOC concentration.

The dissolved organic carbon concentration of water samples was determined based on

Standard Method 5310 D: Wet-Oxidation Method (APHA, AWWA, and WEF, 2012). In

this method, inorganic carbon was first purged as carbon dioxide (CO2) by first acidifying

the sample to pH 2 and then purging with nitrogen gas. Then, persulphate was used to

oxidize the organic carbon to CO2 in an autoclave at temperature from 116 to 130 oC and

the CO2 formed quantified by non-dispersive infrared spectrometry. A 5 % phosphoric

acid solution and 1.0 mg/mL potassium hydrogen phthalate stock solutions were prepared

by diluting phosphoric acid (H3PO4, 85%) and potassium hydrogen phthalate (C8H5KO4,

≥99.95 %) with Milli-Q®, respectively. A 100 g/L sodium persulphate solution was

prepared by dissolving sodium persulphate (Na2(SO4)2, ≥98 %) pellets in Milli-Q®. The

potassium hydrogen phthalate stock solution was acidified to <2 with sulphuric acid


52

(H2SO4, >98 %). A calibration curve was prepared with the potassium hydrogen

phthalate stock solution using 5 standards in the range of 0.625 to 10 mg/L, and running

check standards were prepared at a concentration of 2.5 mg/L at a rate of 1 per 10

samples. Calibration and running check standards were also acidified to <2 with

sulphuric acid. Calibration, running check standards and water samples were prepared in

40 mL amber vials for analysis with the Aurora 1030 TOC analyzer.

The UV absorbance (190 to 1100 nm, at a step of 1 nm) of samples was measured based

on Standard Method 5910 B (APHA, AWWA, and WEF, 2012). Samples were pre-

filtered with Supor® 0.45 µm PES membrane filters to remove interference by particulate

matter and an aliquot analyzed with the spectrophotometer in a 1cm quartz cuvette. The

spectrophotometer was zeroed to Milli-Q®.

Samples were shipped via courier to the University of Waterloo for analysis by liquid

chromatography with organic carbon detection (LC-OCD). A 40 mL aliquot was placed

in a 40 mL vial after filtration with Supor® 0.45 µm PES membrane filter and stored at 4

oC prior to analysis.

3.3.6 UV Fluence Rate

The nanomaterials used in the experiments are photoactive at wavelengths ≤ 424 nm and

so these wavelengths were used to calculate the “UV” fluence rate. UV fluence rate at the

water surface was averaged based on the water sample depth for experiments using a

TiO2 concentration of 0.1 g/L through a procedure outlined by Bolton et al., (2003) and a

modified Bolton® Excel Spreadsheet for fluence calculations using a medium-pressure

lamp with a suspension depth of greater than 2 cm (Bolton 2004). Fluence rate at the

water surface was corrected for a water factor (based on UV absorbance with water

depth), divergence factor (based on distance between water surface and lamp source) and

reflection factor (RF). A Petri factor (spatial variation of UV Fluence rate on water

surface) was not calculated since the spatial variation was very low (less than 2%).

Fluence rate at the water surface for wavelengths from 300 to 424 nm was 13.4 mW/cm2

(Chawla, 2014). with the solar simulator when set at “one sun” intensity. A reflection

factor of 0.975 was used, based on a flat water surface. The distance between the water


53

surface and lamp source was 125 cm and the water sample depth was 6.0 cm. The

absorbance coefficients of a 0.1 g/L P25 TiO2 slurry in Milli-Q® for 300 to 424 nm

wavelengths at 1 nm increments were determined using the UV-Vis spectrometer. Then

an average adsorption coefficient was determined in 5 nm wavelengths increments to

input into Bolton® Excel Spreadsheet. The solar simulator spectral distribution in the 300

to 424 nm wavelength range was calculated based on data supplied by the solar simulator

manufacturer (Chawla, 2014). The product of the water factor (WF) and divergence

factor (DF) was calculated through a series of numerical integrations using Microsoft

Excel®. The average fluence rate based on sample depth (refer to Eq. 3.1) was

determined and the UV dose (mJ/cm2) for irradiations times was calculated as the product

of UV fluence rate (mW/cm2) and irradiation time (s).

Where RF=reflection factor

WF= water factor

DF = divergence factor

(3.1)

3.4 Statistical Analysis of Data

3.4.1 Analysis of Variance

Analysis of Variance (ANOVA) is a statistical technique for analyzing the variation due

to different sources. A two factor ANOVA was completed at 95 % confidence level with

TiO2 material as one factor (with 6 levels) and UV dose the other factor (with 5 levels)

for both Synthetic and Otonabee waters using Minitab®. Normal distribution and equal

variance was assumed and standard deviations were pooled. Each response variable was

calculated as a % reduction since the raw water quality varied between source waters and

batches of source water. Refer to Table 3-9 for ANOVA parameters.

Table 3-9: ANOVA Parameter Description

Parameter Description Levels

Factor 1 TiO2 Material P25, NB, Ag@SiO2@TiO2/P25, anatase,

anatase-N, anatase-B

Factor 2 UV dose control, 0, 1, 15, 30

Response Precursors DOC, UV254

Response DBPs THM fp, HAA fp, HAN fp, HNM fp, HK fp


54

The ANOVA calculations determined whether there was an effect from each factor

separately and any interaction between the two factors, for each response. The

calculations used in ANOVA are given in Eqs 3.2 to 3.5.

( )

Where y = measured response factor

μ = overall mean of response factor (concentration)

α = effect of treatment factor A (TiO2 type)

β = effect of treatment factor B (irradiation time)

(αβ) = effect of combined interactions of factors A and B

ε = random error

i = ith

level of factor A (6 levels)

j = jth

level of factor B (5 levels)

t = replicate (1, 2 or 3)

(3.2)

The null hypothesis for factor A (TiO2 type),

(3.3)

is rejected, and TiO2 type is deemed to have a significant impact on the response factor, if

at least one αi is not equal to zero.

The null hypothesis for factor B (treatment time),

(3.4)

is rejected, and treatment time is deemed to have a significant impact on the response

factor, if at least one βj is not equal to zero.

The null hypothesis for the interaction of factors A (TiO2 type) and B (treatment time),

( ) For all i, j (3.5)

is rejected, and interactions between TiO2 type and treatment time are deemed to have a

significant impact on the response factor, if at least one (αβ)ij is not equal to zero.


55

3.4.2 Coefficient of Determination

The coefficient of determination R2, alternatively known as the square of the sample

correlation coefficient (r)2, was calculated for the correlation of DBP precursors and

DBPs. The formula for R2 is provided below. It measures the percentage of variability in

the y’s (dependent variable) explained by the x’s (independent variable). It was also used

to check the calibration curves.

( ) ∑ ( ̅)( ̅)

√∑ ( ̅) ∑ ( ̅)

(3.6)

3.5 QA/QC Measures

3.5.1 Analytical QA/QC

THM, HAA and HAN running check standards were analyzed and plotted on quality

control charts as per Standard Method 1020 (APHA, AWWA, WEF, 2012) to ensure

accuracy of experimental results. The charts are appended as Figure 10-2 to Figure 10-6.

Laboratory procedures, GC-ECD chromatogram integrations, and/or new calibration

standards were prepared for subsequent experiments if any of the following trends were

observed:

• 2 consecutive measurements outside the control limits of Mean ± 3×standard

deviation(s) (upper control limit (UCL) and lower control limit (LCL));

• 3 out of 4 consecutive measurements were outside of mean ± 2×standard deviation(s)

(upper warning limit (UWL) and lower warning limit (LWL));

• 5 out of 6 consecutive measurements were outside of mean ± standard deviation(s);

• 5 out of 6 consecutive measurements were following a trend of increasing or

decreasing;

• 7 consecutive measurements were > the mean, or 7 consecutive measurements were <

the mean.


56

The standard deviations were calculated by analyzing 7 to 9 individually prepared

standard standards in sequence along with a new calibration curve. The QA/QC charts

were centered on the theoretical check standard concentration.

Method detection limits were also determined for the DBPs based on 7 to 9 individually

prepared standard standards in sequence along with calibration data.

A quality assurance and quality control method was also followed for dissolved organic

carbon (DOC) detection with the OI Analytical Aurora model 1030 (Aurora). A

calibration curve was prepared with each run on the Aurora and if the R2 value was less

than 0.995, a new calibration curve was prepared before sample analysis. Running check

standards at 3 mg/L from a separate source from the calibration standards were analyzed

every 10 environmental samples to verify calibration. If the results were out of

acceptance range (± 10% from expected concentration), a new calibration curve was

prepared and all samples back to last good check standard were reanalyzed. A reagent

blank was run after each check standard. If the reagent blank was greater than the method

detection limit (MDL), a new calibration curve was prepared and all associated samples

reanalyzed.

3.5.2 Experimental QA/QC

The TiO2/UV experiments were completed in quadruplicate in order to have duplicate or

triplicate DBP analysis for each treatment. Average values with error bars are provided in

the results section to show variability among sample replicates. Data from all replicates

was input to Minitab® for ANOVA analysis to determine the statistical significance of

the data based on sample variation. Blank MTBE and Milli-Q® samples were analyzed

during analysis with the GC-ECD and TOC analyzer after running check standard

samples. An internal standard was spiked into all environmental samples and running

check standards during DBP analysis to account for differences in extraction efficiency

between samples. Area response ratios were used to calculate actual DBP concentrations,

as per sample calculations in Appendix 10.3.1.


57

Raw and treated water samples were stored in opaque carboys and amber bottles,

respectively at 4 oC to preserve NOM. Holding times were generally kept to within

standard methods except in some circumstances. For example, DBP fp in Otonabee raw

water batches was highly variable so one batch of Otonabee water was obtained and

homogenized for subsequent experiments testing the innovative TiO2 nanomaterials. This

batch was stored for approximately one month instead of 7 days. Other reagents were

also stored as required for preservation or safety purposes.


58

4 PRELIMINARY TESTS AND TYPICAL DATA SETS

Six proof-of-concept bench-scale tests were completed to evaluate the trihalomethane

(THM) and haloacetic acid (HAA) formation potential (fp) in TiO2/UV treated water

following the UFC chlorination test. TiO2 treatment efficiency was examined based on its

concentration, type, configuration, and the water source employed in the test.

4.1 Overview of Experiments

Batch experiments utilized synthetic river water ‘synthetic water’ (laboratory-grade water

spiked with Suwannee River natural organic matter (NOM), Alginic acid, and inorganic

ions as described in Chapter 3) and Otonabee River water ‘Otonabee water’. The

Otonabee water was sourced from the intake of the Peterborough Utilities Commission

(PUC) Water Treatment Plant (WTP) intake during multiple sampling events.

A SS150AAA Solar Simulator matched the natural solar electromagnetic radiation

spectrum at approximately 108 mW/cm2 “one sun” light intensity (300 – 1100 nm).

Industry standard Aeroxide® P25 TiO2 was applied in both suspended form and

immobilized as a thin film, and TiO2 nanobelts were applied in suspension. As a control,

one sample was irradiated with no TiO2 to confirm photolysis was not a major contributor

to DBP precursor reduction. TiO2 powder was added to each 200 mL water sample and

continuously stirred throughout the experiment. The samples first underwent 30 min of

dark adsorption followed by irradiation for 0, 30, 60, 90 or 120 min. A raw water control

was prepared for each experiment utilizing a new batch of water and followed the same

handling procedures as the treated samples. Following treatment, the samples were

filtered through 0.45 µm Supor® polyethersulfone (PES) membrane filters to remove the

TiO2. The dissolved organic carbon (DOC) concentration and UV254 absorbance of the

filtrate were measured to investigate the degradation and reduction of DBP precursors,

where DOC is a general measurement of NOM and UV254 absorbance is correlated humic

substances which are often the NOM fraction most correlated to DBP fp. The samples

from one experiment were analyzed with liquid-chromatography organic carbon

detection (LC-OCD) to determine the degradation and reduction of NOM fractions by


59

TiO2/UV. Each experiment produced four replicate samples, all of which were analyzed

for DOC and UV254 while two replicates were analyzed for DBP fp. The uniform

formation condition (UFC) chlorination test, which employs a chlorine residual of 1

mg/L after 24 hr at a pH of 8 and temperature of 20 oC was followed to produce DBPs.

4.2 NOM Reduction

The first experiment conducted was with P25 at a concentration of 0.5 g/L suspended in

synthetic water. The results are provided in Figure 4-1. The UV254 of the synthetic water

decreased after 30 min of dark adsorption to TiO2 and continued to decrease with

increasing irradiation time. The DOC concentration of the water decreased during the

dark adsorption step, but was observed to fluctuate thereafter with irradiation time.

Although the DOC concentration decreased from approximately 3 mg/L to 2 mg/L (a 30

% reduction) the UV254 was reduced from approximately 0.04 cm-1

to 0.006 cm-1

(an 85

% reduction). DOC is a measure of the total organic carbon concentration of the water

and does not reflect the degradation of compounds into intermediate products. Its

decrease can be inferred to be from the adsorption of organic compounds to the TiO2

surface or mineralization from photocatalysis. The UV254 absorbance of the water

measures the aromaticity of the NOM and generally reflects the degradation of humic

substances, but does not elucidate the character of the degradation products.

Figure 4-1: DOC and UV254 in Synthetic Water Treated with 0.5 g/L P25 in

Suspension

0.00

0.05

0.10

0.15

0.20

0.0

1.0

2.0

3.0

4.0

5.0

6.0

Control 0 30 60 90

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)

Irradation Time (min)

DOC

UV254

Dark

Adsorption Only


60

The UV-Vis absorbance of the synthetic water treated with P25 TiO2/UV at a

concentration of 0.5 g/L is shown in Figure 4-2, enlarged in the region of 254 nm with

the inset showing the UV-Vis absorbance over the entire range measured (190-1100 nm).

Natural surface water typically has a relatively featureless UV-Vis absorbance spectrum,

as observed with the raw synthetic water (control), in which the absorbance simply

decreases with increasing wavelength. It appears that TiO2/UV treatment degrades

compounds absorbing at >240 nm wavelengths. The absorbance spectra of the waters

treated with 60 and 90 min of TiO2/UV are similar suggesting that the limit for impacts

on UV-Vis absorbance had been reached between the 30 and 60 min of treatment. Similar

results were observed by Liu et al., in which two natural surface waters were treated with

TiO2/UV at 0.1 g/L in suspension with a blacklight blue fluorescent lamp (maximum

emission at 365 nm) in an annular reactor for irradiation times ranging from 0 to 150 min

(Liu et al., 2010).

Figure 4-2: UV-Vis Absorbance of Synthetic water Treated with P25 TiO2/UV

The next experiment followed the same methodology as the first experiment but used

Otonabee water as the water source. Results of DOC and UV254 absorbance

measurements are provided in Figure 4-3. The DOC concentration and UV254 absorbance

of the raw Otonabee water control was approximately double (6 mg/L) and four times

0

0.05

0.1

200 250 300 350 400

UV

-Vis

Ab

sorb

an

ce (

1/

cm)

Wavelength (nm)

Control

0 min

30 min

60 min

90 min


61

(0.176 cm-1

) that of the raw synthetic water control in the experiment immediately

previous to this one. This Otonabee water experiment included an additional irradiation

time (120 min) to determine the extent of degradation of recalcitrant compounds with

longer irradiation time. The UV254 absorbance of the Otonabee water decreased with 30

min of TiO2 dark adsorption and continued to decrease with irradiation. With dark

adsorption the DOC concentration of the Otonabee water decreased, as expected. With

subsequent 30 min of irradiation, the DOC concentration increased. It can be inferred that

NOM desorbed from the TiO2 surface and it was suspected to be caused by a change in

the character of the NOM from partial degradation and/or a change in the chemical

character TiO2 surface (Liu et al., 2008a; Fujishima and Zhang, 2006). Changes to NOM

may also make it more amenable to detection by the TOC analyzer. With further

irradiation to 60 min, the DOC concentration of the Otonabee water decreased and

continued to decrease with longer irradiation times. The DOC concentration and UV254

absorbance were reduced by approximately 30 % and 80 % following 90 min of

irradiation similar to the synthetic water experiment immediately previous to this one,

with slightly less % reduction in UV254 absorbance. With 120 min of irradiation, the %

reduction of DOC concentration and UV254 absorbance in the Otonabee water improved

to 40 % and 85 %, respectively.

Figure 4-3: DOC of Otonabee Water Treated with 0.5 g/L P25 in Suspension

0.00

0.05

0.10

0.15

0.20

0.0

1.0

2.0

3.0

4.0

5.0

6.0

Control 0 30 60 90 120

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)

Irradiation Time (min)

DOC

UV254

Dark Adsorption

Only


62

To better understand the degradation of NOM with TiO2/UV, one sample replicate of

each treatment from this Otonabee water experiment was analyzed by LC-OCD. The

results are shown in Figure 4-4. With TiO2 dark adsorption the humic and biopolymer

portions of NOM decreased. With additional 30 min of irradiation, overall DOC

concentration increased with the humic, building blocks, and LMW acids and neutrals

fractions increasing in concentration while the biopolymers decreased (compared to the

dark adsorption only sample). It can be inferred that following the subsequent 30 min of

irradiation, some humics and biopolymers that had adsorbed to the TiO2 surface during

dark adsorption were degraded into building blocks and low molecular weight (LMW)

acids and neutrals, desorbed from the TiO2 surface, and were detected in the water

matrix. With subsequently longer irradiation times, the humic portion continued to

decrease along with overall DOC concentration while building block and LMW neutral

and acid fractions increased or remained constant. During these longer irradiation times

more humics were degraded to building blocks, LWM neutrals and acids and some

mineralization occurred. The UV254 absorbance of the water is superimposed onto the

LC-OCD results in Figure 4-4. The UV254 absorbance followed approximately the same

trend as the humic fraction. The coefficient of determination was 0.960 when correlating

average UV254 absorbance and the humic fraction concentrations.

Figure 4-4: NOM Fractions in Otonabee Water Treated with 0.5 g/L P25 in

Suspension

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.0

1.0

2.0

3.0

4.0

5.0

6.0

Control 0 30 60 90 120

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Co

nce

ntr

ati

on

(m

g/

L)


LMW Acids

LMW Neutrals

Building Blocks

Humics

Biopolymers

UV254

Dark Adsorption

Only


63

The next experiment was conducted with Otonabee water and P25 similar to the

previously described experiment; however, the TiO2 was applied at a concentration of

0.15 g/L instead of 0.5 g/L to observe the changes in degradation with changes to TiO2

concentration. Results are shown in Figure 4-5. With the higher concentration of TiO2,

there was a greater degree of NOM reduction through dark adsorption. However, with

irradiation, the difference in UV254 absorbance and DOC reduction between 0.5 g/L and

0.15 g/L of TiO2 was not proportional; suggesting that degradation by TiO2/UV was

being influenced by other factors than TiO2 concentration and there was diminishing

return by increasing TiO2 concentration from 0.15 to 0.5 g/L. Previous studies have

reported an optimal catalyst loading of 0.75 to 1 g/L attributed to increasing opacity of

the TiO2 suspension and aggregation of TiO2 particles (Ghaly et al., 2011) however

optimal TiO2 concentration will be dependent on factors such as water quality,

contaminant of interest, TiO2 type, configuration, and UV fluence rate.

Figure 4-5: DOC and UV254 in Otonabee Water Treated with TiO2/UV at 0.5 and

0.15 g/L in Suspension

The next two experiments investigated the effects of TiO2 configuration and type. P25

immobilized as a thin film on a stainless steel mesh and TiO2 nanobelts were provided by

the University of Waterloo. The concentration of TiO2 in these experiments was 0.15 g/L

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.0

1.0

2.0

3.0

4.0

5.0

6.0

Control 0 30 60 90 120

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)


DOC (0.5 g/L P25)

DOC (0.15 g/L P25)

UV254 (0.5 g/LP25)

UV254 (0.15 g/L P25)

Dark Adsorption

Only


64

and the results are compared to P25 in suspension at 0.15 g/L in Figure 4-6. The TiO2

coated mesh was suspended in the 200 mL water sample with a stainless steel support

with about 1 cm of water above it. The water was continuously stirred with stir bar and

stir plate similar to previous experiments however it was noted that the circulation of

water above the mesh was not significant. This experiment was meant to simulate the

application of TiO2 as a thin film or filter but fluid dynamics were not representative. The

DOC concentrations in samples treated with the TiO2 thin film were higher than the

Otonabee raw water control. There was probably an organic carbon contamination in

these samples, possibly from the TiO2 material, stainless steel mesh and/or support

structure. The P25 in suspension was more effective at reducing DOC concentration and

UV254 absorbance compared to the nanobelts. Although nanobelts increase the quantum

efficiency of photocatalysis, they have decreased surface area compared to nanoparticles.


Suspension, P25 Immobilized as a Thin Film, and NB in Suspension

The last experiment in this preliminary data set was a replicate of the experiment treating

Otonabee water with P25 in suspension at 0.5 g/L. This was repeated to include a short

irradiation time (5 min) and to test the degradation of NOM and DBP precursors with

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.0

1.0

2.0

3.0

4.0

5.0

6.0

7.0

8.0

9.0

10.0

Control 0 30 60 90 120

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)


DOC (P25 Suspension)

DOC (Immobilized ThinFilm)

DOC (NB Suspension)

UV254 (P25Suspension)

UV254 (ImmobilizedThin Film)

UV254 (NBSuspension)

Dark

Adsorption Only


65

photolysis only (no TiO2). This sample was irradiation with the solar simulator for 120

min and processed similar to the other treatments. The DOC concentration and UV254

absorbance are plotted in Figure 4-7. The treatment with solar irradiation only had

slightly higher DOC concentration and UV254 absorbance compared to the raw water

control. It is suspected that NOM did not undergo significant photolysis and the increase

seen in DOC concentration and UV254 absorbance was caused by random error in the

experiment from variability in the natural water source quality. It may also have been due

to changes to NOM which made it more susceptible to detection by the TOC analyzer.

The increase in UV254 absorbance may have been due to TiO2 nanoparticles that passed

through the 0.45 µm Supor® PES membrane filter and obstructed light during

measurement of UV254 absorbance with the UV-Vis spectrometer. There was a decrease

in DOC concentration and UV254 absorbance in Otonabee water during the TiO2 dark

adsorption step. Following the dark adsorption step, there was in increase in DOC

concentration and UV254 absorbance with subsequent 5 min of irradiation. An increase to

the hydrophilicity of TiO2 upon irradiation, as observed by some researchers (Fujishima

and Zhang, 2006) may have caused aromatic compounds to desorb from the TiO2 surface.

With continued irradiation, DOC and UV254 absorbance decrease as NOM was

mineralized and aromatic structures degraded both by direct e-/h

+ degradation and

oxidation by reactive oxygen species (ROS) in solution.


Suspension (Duplicate Experiment)

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.0

1.0

2.0

3.0

4.0

5.0

6.0

Control 120(Solaronly)

0(Dark Ads.

Only)

5 30 60 90 120

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)


DOC

UV254


66

4.3 DBP fp Reduction

Following treatment with TiO2 the water samples were filtered with 0.45 µm Supor®

PES membrane filters. Two replicates of each treatment level were utilized to determine

the chlorine demand of the water at that treatment level and calculate the required

chlorine spike concentration for a 1 mg/L chlorine residual after 24 hr. The remaining

two samples underwent the UFC chlorination test and analysis for THM and HAA fp.

The reduction of THM and HAA fp upon chlorination with TiO2 treatment prior to

chlorination was examined based on water source, and TiO2 concentration, type, and

configuration. Results for DBP fp are given as averages of the two replicates tested, with

high and low values as error bars.

In all of the preliminary experiments chloroform (trichloromethane – TCM) was the only

THM detected and dichloroacetic acid (DCAA) and trichloroacetic acid (TCAA) were

the only HAAs detected. The THM concentration reported is thus solely from TCM while

the HAA concentration is the sum of the DCAA and TCAA concentrations. The synthetic

water and Otonabee water did not have elevated bromide concentrations and brominated

DBPs were not expected. Monochloroacetic acid quickly converts to DCAA or TCAA in

the presence of chlorine and its absence in the water post chlorination was also expected.

Figure 4-8 presents the THM fp results from the synthetic water treated with P25 TiO2 in

suspension at a concentration of 0.5 g/L and chlorination. The raw synthetic water had a

THM fp of approximately 120 µg/L. After a significant decrease in THM fp (70 %

reduction) during the TiO2 dark adsorption step, there was an increase in THM fp with

subsequent 30 min of irradiation. The increase in THM fp may be from precursors that

desorbed from the TiO2 surface or formed in solution during photocatalysis. The THM fp

generally decreased with longer irradiation times and after 90 min of irradiation, the

THM fp was slightly lower than the THM fp following dark adsorption only (74 %

overall reduction). The ‘specific’ (sp) THM fp (THM fp normalized to DOC

concentration) is also plotted on Figure 4-8. The sp THM fp accounts for the fraction of

the organic matter that contains those particular DBP precursors. Figure 4-8 demonstrates

that THM precursors were preferentially reduced in this experiment from dark adsorption


67

only and photocatalysis. In this particular experiment, the TiO2 had a strong affinity for

the THM precursors in the synthetic water under dark adsorption. Liu et al. and other

research groups have studied the use of TiO2 as an adsorbent to remove humic acids from

aqueous solutions under dark conditions, with photocatalysis used to regenerate the TiO2

(Liu et al., 2014).

Figure 4-8: THM fp in Synthetic Water Following Treatment with P25 at 0.5 g/L in

Suspension and Chlorination

The THM fp results of the four preliminary experiments conducted with Otonabee water

are plotted in Figure 4-9 . These experiments compare the effects of TiO2 concentration,

configuration, and type. The filtered and chlorinated raw Otonabee River water DBP fp

fluctuated between different batches of water obtained from the PUC and can be

attributed to the variability in water quality in the Otonabee River during different water

collection days. The concentration of THM varied from approximately 250 to 400 µg/L

and was higher than the raw synthetic water (120 µg/L). The THM fp was higher

compared to synthetic water but generally the Otonabee water treated with P25 at a

concentration of 0.5 g/L in suspension followed the same trends in reduction as the

synthetic water treated with P25 at a concentration of 0.5 g/L in suspension, albeit with

0

5

10

15

20

25

30

35

40

45

50

0

100

200

300

400

Control 0 30 60 90

Av

g. C

on

c. /

Av

g. D

OC

(µ

g/

mg

)

Av

g. C

on

c. (

µg

/L

)


THMfp

SpTHMfp

Dark

Adsorption Only


68

less THM fp reduction (34 % compared to 70% with dark adsorption only and 45 %

compared to 74% following 90 min or irradiation). As per the Langmuir-Hinshelwood

model, initial DBP precursor concentration affects degradation kinetics. It was expected

that Otonabee water THM fp % reduction would be less than synthetic water because of

the higher initial THM fp, or different than the synthetic water due to differences in water

quality parameters that also affect the Langmuir-Hinshelwood model and TiO2/UV in

general.

The Otonabee water treated with P25 at lower concentration (0.15 g/L compared to 0.5

g/L) had less THM fp reduction with dark adsorption (22% compared to 34 %) but

comparable THM fp reduction following 90 min or irradiation (43% compared to 45%).

Thus the THM fp exhibited similar diminishing returns with increasing TiO2

concentration (0.15 to 0.5 g/L) as the DOC concentration earlier discussed.

Treatment with the P25 thin film and nanobelts generally was not as effective as the P25

powder in suspension. The THM fp % reduction after 120 min of irradiation with the thin

film and nanobelts was 22 and 3 %, respectively.

Figure 4-9: THM fp in Otonabee River Following Treatment with TiO2/UV and

Chlorination

0

50

100

150

200

250

300

350

400

450

500

Control 0 30 60 90 120

Av

g. C

on

c. (

µg

/L

)


P25 (0.5 g/L Suspension)

P25 (0.15 g/ L Suspension)

P25 (Immobilized)

NB (0.15 g/L Suspended)

Dark Adsorption

Only


69

HAA fp results of the preliminary experiments are provided in Figure 4-10 and Figure

4-11 for synthetic and Otonabee water, respectively. The control raw synthetic water

HAA fp was approximately 70 µg/L. The control filtered and chlorinated raw Otonabee

water HAA fp was higher than the synthetic water and varied considerable with each

batch of water obtained from the Otonabee River similar to the THM fp (70 to 150 µg/L).

The trends in HAA fp with TiO2 treatment are similar to those observed with THM fp.

Figure 4-10: HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV at

0.5 g/L in Suspension and Chlorination


Chlorination

0

5

10

15

20

25

30

35

40

45

50

0

20

40

60

80

100

120

140

160

180

200

Control 0 30 60 90 120

Av

g. C

on

c./

Av

g. D

OC

(µ

g/

mg

)

Av

g. C

on

c. (

µg

/L

)


HAAfp

SpHAAfp

Dark

Adsorption Only

0

50

100

150

200

Control 0 30 60 90 120

Av

g. C

on

c. (

µg

/L

)


P25 (0.5g/L Suspension)

P25 (0.15 g/L Suspension)

P25 (0.15 g/L Immobilized)

NB (0.15 g/Lsuspension)

Dark Adsorption

Only


70

The replicate experiment testing the DBP fp during solar irradiation with no TiO2 showed

that the THM fp and HAA fp did not change significantly compared to the raw water

control. The THM fp for the raw water control and irradiation only (no TiO2) samples

was 311 +/- 15 µg/L and 312 +/- 10 µg/L, respectively while the HAA fp was 137 +/- 7

µg/L and 132 +/- 1 µg/L, respectively.

4.4 TiO2 Configurations

A summary of the preliminary DBP results, namely the changes in THM and HAA

formation potential (fp) following 60 min of photocatalysis as determined by the UFC

chlorination test, are given in Table 4-1. Results are compared after 60 min rather than

the full treatment time (90 or 120 min) because recalcitrant compounds may cause

inaccurate findings after longer treatment times. The % reduction is given to account for

the variability in raw water quality between batches of collected natural source water and

laboratory prepared synthetic water. Not unexpectedly, the concentrations of the

precursors of these DBPs are different in different types of samples, and their responses

to photocatalysis are also different. Comparison of the results for treating synthetic water

and Otonabee water with suspended photocatalyst at a concentration of 0.5 g/L shows

that there were approximately 10 % more THM and HAA precursors’ reduction in the

synthetic water relative to the corresponding precursors in Otonabee water. It is also

reasonable to see that higher concentrations of photocatalyst reduce THM fp and HAA fp

more (compare the two concentrations of photocatalyst (0.5 and 0.15 g/L) in suspension

used to treat Otonabee water). Also shown in Table 4-1 are results for the ‘thin film’

experiment with 0.15 g/L P25 immobilized onto a support structure. A negative %

reduction in the results illustrates the need to be aware of optimizing mixing around the

immobilized P25 and/or reducing carbon leaching from the support structure, issues of

focus for work moving forward with immobilized photocatalysts. The nanobelts were less

effective in reducing the concentration of DBP precursors compared to P25.


71

Table 4-1: Summary % Reduction of THM and HAA fp in Preliminary

Experiments Following 60 min of TiO2/UV Treatment and Chlorination

Avg. %

Reduction

Synthetic

Water

Otonabee

water

Otonabee

water

Otonabee

water

Otonabee

water

0.5 g/L P25 0.5 g/L P25 O.15 g/L P25 O.15 g/L P25 0.15 g/L NB

In

Suspension

In

Suspension

In

Suspension

Immobilized

as Thin Film

In

Suspension

THM fp 46% 38% 30% -3% 17%

HAA fp 85% 74% 27% -17% 4%

4.5 Optimum TiO2 Concentration

As per Eq. 2.25 in Chapter 2 provided below, a pseudo first order reaction rate constant

was assumed for DOC and UV254 reduction based on a constant hydroxyl radical

concentration. The reaction rate constants for DOC and UV254 reduction in synthetic

water were determined at P25 TiO2 concentrations ranging from 0.005 to 0.5 g/L as per

Eg. 2.23 in Chapter 2 also provided below.

Pseudo first order reaction

rate constant

[ ] Where kactual = actual second order reaction

rate constant

(2.25)

Linear plot to determine k

(

) ,

Where slope of line = -k

R = DOC or UV254 at time t

Ro= DOC or UV254 in the control

(2.23)

The concentration of hydroxyl radicals should be proportional to the TiO2 concentration

until other factors limit the production of hydroxyl radicals, such as the screening effect

of TiO2 particles or their agglomeration and subsequent decrease in surface area. The P25

TiO2 concentrations in suspension were chosen based on concentrations used by other

researchers and observations made in preliminary experiments. TiO2/UV experiments

were completed with 10 min of dark adsorption and 0, 0.5, 1, 2, 5, 10, 15, 30, 60 min of

irradiation. The TiO2 was added to the water samples from a 10 g/L stock slurry to obtain

the required TiO2 concentration in suspension. Each treatment was duplicated and the

results averaged. The raw water control sample was plotted at time of 0 min while the


72

dark adsorption sample was not included in the analysis since the reaction rate constant is

based on TiO2 irradiation and the production of hydroxyl radicals. Due to the effects of

adsorption and desorption, DOC and UV254 appeared to increase following dark

adsorption with short irradiation times. The apparent increase in DOC and UV254 between

the dark adsorption only treatment and short irradiation times may be related to changes

to the TiO2 itself or changes in NOM character upon degradation (Liu et al., 2008b;

Fujishima and Zhang, 2006). These data points were not plotted so as to obtain a general

trend of the overall degradation of DOC and UV254 with irradiation time from hydroxyl

radical oxidation. The plot for TiO2/UV at a concentration of 0.1 g/L is shown in Figure

4-12. The degradation of DOC and UV254 follow pseudo first order reaction kinetics with

a coefficient of determination of 0.970 and 0.993, respectively. The UV254 reaction rate

constant was larger than DOC suggesting aromaticity is degraded faster than NOM

mineralization occurs.

Figure 4-12: Determining Reaction Rate Constant for TiO2/UV at 0.1 g/L

A summary of the calculated reaction rate constants is provided in Table 4-2. The 0.1 g/L

TiO2 concentration appeared to be the most effective concentration tested. The DOC and

UV254 reaction rate constants increased by approximately an order of magnitude from

0.005 g/L to 0.05 g/L TiO2. Thus, TiO2 was applied at an effective concentration in both

y = -0.0087x - 0.0414 R² = 0.9704

y = -0.0319x - 0.0102 R² = 0.9929

-2.5

-2

-1.5

-1

-0.5

0

0 10 20 30 40 50 60 70

Ln

(R/

Ro

)


DOC

UV


73

of these experiments and the concentration of TiO2 could be increased without

diminishing returns. The DOC and UV254 pseudo first order reaction rate constants

almost doubled in magnitude from 0.05 to 0.5 g/L TiO2. Thus, there were diminishing

returns in increasing the TiO2 concentration an order of magnitude from 0.05 to 0.5 g/L

and TiO2 was not effective at 0.5 g/L. Therefore, an optimal TiO2 concentration would

be between 0.05 and 0.5 g/L. This was expected based on preliminary experiments with

TiO2 at a concentration 0.15 and 0.5 g/L.

At a TiO2 concentration of 0.1 g/L, the DOC reaction rate constant was comparable to the

DOC reaction rate constant at a TiO2 concentration of 0.05 g/L while the UV254 reaction

rate constants approximately doubled. Although there was variability in the experimental

data, it appeared that there were diminishing returns in increasing TiO2 concentration

from 0.1 g/L to 0.2 g/L and 0.5 g/L. Therefore, 0.1 g/L was chosen for subsequent

experiments. Other studies have suggested higher optimal catalyst doses (Ghaly et al.,

2011) but may have made those conclusions based on increases in the reaction rates and

not whether there were diminishing returns. The TiO2 concentration of 0.1 g/L was also

chosen to compare results with other studies that also worked with a TiO2 concentration

of 0.1 g/L (Liu et al, 2010; Liu et al, 2008a; Liu et al, 2008b). This research group chose

this concentration rather than 1 g/L in order to investigate degradation mechanisms which

would be slower and easier to observe at the lower TiO2 concentration.

Table 4-2: Pseudo First Order Reaction Rate Constants

TiO2 Concentration

(g/L)

DOC

k (1/s)

UV254

k (1/s)

0.005 .0011 .0043

0.005 (replicate) .0007 .0028

0.05 .0084 .0203

0.05 (replicate) .0086 .0247

0.1 .0087 .0414

0.2 .0082 .0288

0.5 .0135 .0349

0.5 (replicate) .0147 .0412


74

The reaction rate constants determined were based on the bulk organic measurements of

DOC concentration and UV254 absorbance which included the degradation of many

different types of NOM compounds each with its own characteristic second order

reaction rate constant with HO*. This may have caused the variability observed in the

data. Also, the synthetic water used for these tests would have affected hydroxyl radical

production due to alkalinity and pH. Another procedure to determine the optimum TiO2

concentration would be to measure the degradation of a probe compound with known

second order reaction rate constant with hydroxyl radicals. There are many probe

compounds that are being investigated for this purpose, including methylene blue,

parachlorobenzoic acid, and sucralose (Keen and Linden, 2013). This test was not

included in the current research and was out of scope of the current research objectives.

The peak for DOC concentration with TiO2/UV occurred at approximately 15 minutes for

the experiment with 0.1 g/L TiO2 in suspension. Irradiation times of 0, 1, 15 and 30 min

were chosen for subsequent experiments to observe reduction by dark adsorption (0 min),

the potential maximum DOC concentration that might be expected during irradiation (15

min), and to simulate potential scale up scenarios with a short irradiation time (1 min)

expected in a flow through system (e.g. membrane filtration) and a long irradiation time

(30 min) possible with a batch reactor.

4.6 Optimum TiO2 Dark Adsorption Time

The adsorption of NOM onto the TiO2 surface constantly changes throughout the

irradiation period of a batch reactor, and having maximum dark adsorption prior to

irradiation may not significantly affect the results of subsequent photocatalytic

degradation that relies on both direct e-/h

+ pair and ROS degradation. It may be more

important for a reactor where the TiO2 is used as an adsorbent and regenerated with

irradiation. The maximum adsorption occurred at shorter adsorption times (about 1min

for DOC as shown in Figure 4-13) when a sonicated TiO2 stock solution was added to a

water sample compared to the addition of TiO2 powder to a water sample without

sonication (30 min in the preliminary proof of concept experiments). Sonication may

break up clumps of TiO2 nanoparticles and expose more surface area for adsorption. The


75

effects of different adsorption times on subsequent reduction in DOC concentration and

UV254 absorbance during irradiation were investigated using the same methodology used

in Section 4.4 however the TiO2 stock solution was sonicated and added to the water

samples. It was determined that dark adsorption time did not appreciably affect

subsequent irradiation effectiveness, as shown in Figure 4-13 for DOC concentration.

Results for UV254 absorbance are included in the Appendix. For subsequent experiments,

a dark adsorption time of 1 min was chosen. A shorter dark adsorption time may be

representative of an actual full scale water treatment operation, where it might occur

when TiO2 is added and mixed into the water matrix prior to irradiation.

Figure 4-13: DOC in Synthetic Water Following Treatment with P25 TiO2/UV

under Various Dark Adsorption and Irradiation Times

4.7 UV Fluence Rate

The solar simulator lamp spectral flux was provided by the manufacturer. Table 4-3

provides a summary of the relative lamp flux for the wavelengths of interest (300 to 424

nm). The absorbance coefficients were determined for a P25 TiO2 suspension at 0.1 g/L

in Milli-Q® with the UV-Vis Spectrometer. Actual environmental samples have

additional absorbance from the water matrix but this additional absorbance was

significantly less than the effects of the TiO2 itself for Otonabee and synthetic water and

0

0.5

1

1.5

2

2.5

3

0 1 30

Av

g. D

OC

Co

nc.

(m

g/

L)


0 ads

1 ads

2 ads

5 ads

10 ads


76

was not considered for the purpose of determining UV fluence rate through the water

sample. The average absorbance coefficients (5 nm wavelength increments) for the TiO2

slurry in Mili-Q® were between 1.6 to 3.6 cm-1

in the 300 to 424 nm range and are

provided in Table 4-3. In comparison the absorbance coefficients in the 300 – 424 nm

range for synthetic and Otonabee water were between 0.003 to 0.05 cm-1

and 0.006 to

0.08 cm-1

, respectively.

Table 4-3: UV Fluence Rate Raw Data Calculations

Wavelength

(nm)

Lamp

Flux

(rel)

N( )

Absorbance

Coefficient

(cm-1

)

WF x DF

Lamp

Flux x

WFxDF

300-304 5.2 3.57 0.02015 0.10

305-309 4.9 3.54 0.02032 0.10

310-314 4.7 3.63 0.01982 0.09

315-319 5.8 3.58 0.02011 0.12

320-324 5.1 3.58 0.02012 0.10

325-329 5.6 3.37 0.02133 0.12

330-334 3.2 3.42 0.02107 0.07

335-339 3.8 3.26 0.02208 0.08

340-344 5.0 3.08 0.02332 0.12

345-349 7.5 3.01 0.02387 0.18

350-354 17.8 2.83 0.02543 0.45

355-359 35.7 2.66 0.02702 0.97

360-364 55.0 2.55 0.02818 1.55

365-369 59.9 2.43 0.02954 1.77

370-374 57.1 2.32 0.03097 1.77

375-379 54.8 2.22 0.03239 1.77

380-384 60.6 2.17 0.03303 2.00

385-389 68.2 2.10 0.03414 2.33

390-394 76.4 2.02 0.03562 2.72

395-399 82.5 1.93 0.03719 3.07

400-404 78.4 1.87 0.03844 3.02

405-409 74.3 1.79 0.04003 2.98

410-414 75.8 1.74 0.04135 3.13

415-419 74.9 1.68 0.04276 3.20

420-424 77.8 1.62 0.04419 3.44

TOTAL 1000

35.25

Weighted

average

WFxDF =

0.0352


77

The water sample depth was 6 cm during TiO2/UV experiments and the distance between

the lamp and the surface of the water was 125 cm. The product of the water factor (WF)

and divergence factor (DF) was determined through a series of integrations using the

procedure outlined by Bolton et al., (2003) and the modified Bolton® Excel Spreadsheet

for fluence calculations using a medium-pressure lamp with a suspension depth of greater

than 2 cm (Bolton 2004). The results are provided in Table 4-3. The weighted product of

the water factor and divergence factor was determined for each 5 nm wavelength based

on lamp flux and then a weighted average WFxDF value was determined.

The UV fluence rate at the water surface was 13.4 mW/cm2. A reflection factor of 0.975

was used, based on a flat water surface (Bolton 2004). An experiment was completed to

determine if the vortex used in the samples affected the degradation of methylene blue

compared to a flat water surface, and no significant difference was observed (refer to

Figure 10-16 appended) therefore the reflection factor was assumed to be appropriate.

The corrected UV fluence rate, as shown in Table 4-4, was 0.46 mW/cm2. As expected,

light did not penetrate deep into the water sample. This is important to consider when

designing TiO2/UV reactors for TiO2 suspensions. Large surface areas, shallow depths

and long irradiation times may be required in TiO2 slurries. Immobilized TiO2 reactors

allow for increased light penetration through the water surface but provide less surface

area for adsorption and contact between NOM and e-/h

+ pairs and ROS. This is also an

important factor to consider when comparing results in literature based on UV dose when

it is not specified if the UV dose provided was measured at the water surface or corrected

for reflection, water and divergence. For this reason, measuring actual performance of the

TiO2/UV based on hydroxyl radical production is beneficial and has the added advantage

of being able to be compared to other AOPs that do not use UV.

Table 4-4: Average UV Fluence Rate Calculations

Correction

Factor Value UV Fluence Rate

Value

(mW/cm2)

Radiometer reading at the surface = 13.4

RF = 0.975

True incident irradiance entering the water = 13.0

WFxDF = 0.035

Avg. unweighted irradiance through the water = 0.46


78

The UV dose (mJ/cm2) for TiO2/UV irradiation times was calculated as the product of the

corrected UV fluence rate (mW/cm2) and irradiation time (s). The UV Dose at 0, 1, 15

and 30 min were determined to be 0, 28, 414 and 827 mJ/cm2, respectively. The 1 min

irradiation time was chosen as a short irradiation time representative of a flow through

membrane filter, overflow thin film reactor or other large scale application. The

associated UV dose of 28 mJ/cm2 is similar to typical UV doses applied during UV

disinfection. The 30 min irradiation time was chosen as a long irradiation time

representative of a potential sequencing batch reactor (SBR), continuously stirred tank

reactor (CSTR) or small scale batch treatment application. The associated UV dose of

827 mJ/cm2 is proximate to a typical UV dose applied during AOPs that utilize UV.

4.8 Summary of Preliminary Results

Preliminary results showed that there is a potential for various forms of TiO2

photocatalysis to reduce DBP precursors prior to chlorination, and % reduction is

affected by source water quality, TiO2 concentration and configuration. Generally raw

Otonabee River water had a higher concentration of DBP precursors and synthetic water

DBP fp % reduction was greater than Otonabee River water. Increasing TiO2

concentration from 0.005 to 0.5 g/L increased the degradation rate of NOM however at

approximately 0.1 g/L a point of diminishing returns was reached. TiO2 suspension

performed better than a TiO2 thin film, likely due to the increased contact between TiO2

and NOM. Industry standard Aeroxide® P25 DBP fp % reduction was higher compared

to the newly fabricated nanobelts from the University of Waterloo, re-affirming the

effectiveness of P25, a nanoparticle mixture of anatase and rutile crystalline phases, as a

photocatalyst (Hurum et al., 2003; Ohn et al., 2001; Ohtani et al., 2010)

Preliminary experiments also estimated optimal TiO2 concentration, dark adsorption time

and irradiation times for subsequent experiments. The equivalent corrected UV dose

(based on reflection, water and divergence factors) was also determined for the chosen

irradiation times. The transmission of light through water samples with TiO2 at a

concentration of 0.1 g/L was low due to turbidity and absorbance by the TiO2 particles

and thus TiO2/UV reactors would benefit from large surface areas for irradiation.


79

5 EFFECTS OF TIO2/UV ON DBP FORMATION IN A MODEL

RIVER WATER

The objective of the experiments described in this chapter was to compare the DBP

formation and precursor reduction with industry standard Aeroxide® P25 and innovative

TiO2 nanomaterials using lab-controllable synthetic water. The new materials included

P25 mixed with a silver based product, nanobelts, anatase, nitrogen doped anatase, and

boron doped anatase. The overall hypotheses were that TiO2/UV would either increase or

decrease DBP fp, and that different types of TiO2 would have varying degrees of effect.

Both short-term exposures (up to a few minutes) representative of flow-through treatment

systems, and longer term exposures of up to 30 min or more that may be more

representative of batch reactor systems were studied.

5.1 Overview of Experiments and Results

Batch experiments were conducted to test changes in THM, HAA, HAN, HNM, and HK

formation potential (fp) in ‘synthetic water’ (lab water augmented to model river water)

treated by TiO2/UV and chlorination. Synthetic water was used in this set of experiments

because its composition and expected DBP fp was known, limiting the effects of

variations in water quality typical of natural sources.

The methodology from preliminary experiments was generally followed, using a TiO2

concentration of 0.1 g/L in suspension. The TiO2 was added to the water sample from a

stock 5 g/L slurry that was sonicated for 5 min and continuously stirred throughout the

course of the experiment to reduce agglomeration of TiO2 powder in the water sample.

Preparing a sonicated stock solution was expected to be a great improvement to the bench

scale batch experiment both for treatment efficiency and replicability since the TiO2

suspension was observed to be much more consistent from this method rather than adding

the TiO2 powder directly to the water sample.

Following treatment, the samples were filtered through 0.45 µm Supor® polyethersulfone

(PES) membrane filters to remove TiO2. To understand DBP control with TiO2/UV by

NOM precursor degradation the dissolved organic carbon (DOC) concentration and


80

UV254 absorbance were monitored. DOC was measured to investigate the general effects

of TIO2/UV on NOM (adsorption to TiO2, degradation and mineralization) while UV254

absorbance was studied as a surrogate for humic substances, which are often the DOC

fraction most correlated to DBP fp. The uniform formation condition (UFC) test, which

employs a chlorine residual of 1 mg/L after 24 hr at a pH of 8 and temperature of 20 oC

was followed for DBP fp.

Two experimental factors were manipulated, TiO2 nanomaterial type and UV dose. TiO2

nanomaterials included industry standard Aeroxide® P25 (P25), newly fabricated P25

nanobelts (NB), P25 mixed with 1% Ag@SiO2@TiO2 triplex core-shell photocatalyst

(Ag@SiO2@TiO2/P25), sol-gel synthesized anatase (anatase), nitrogen-doped anatase

with 5% nitrogen by mass (anatase-N), and boron-doped anatase with 5 % boron by mass

(anatase-B) developed by project partners (Liang, 2014; Hatat-Fraille, 2014). Although

P25 was the predecessor of the NB, the fabrication process that grew the nanoparticles

along one length to nanobelts also affected the crystalline phase of the TiO2 and the NB

were largely composed of the less photoreactive rutile phase (Liang, 2014).

A SS150AAA Solar Simulator was the light source for the experiments. The average UV

fluence rate was determined to be 0.459 mW/cm2 (300 to 424 nm) as shown in Chapter 4.

Water samples with TiO2 in suspension were irradiated for either: 0, 1, 15 or 30 min after

1 min of dark adsorption. The UV dose for each irradiation time was determined as the

product of the average UV fluence rate and irradiation time. Irradiation times of 1, 15 and

30 min corresponded to UV doses of 28, 414 and 827 mJ/cm2, respectively (see Chapter

4 for calculations). A UV dose of 28 mJ/cm2 was representative of a short irradiation time

that might be expected in a TiO2 flow-through membrane filter or thin film reactor

configuration or other large scale application. Typical UV disinfection as currently

practiced applies a dose similar in magnitude (approximately 40 mJ/cm2) albeit at shorter

germicidal UV wavelengths. A UV dose of 827 mJ/cm2 is representative of a longer

irradiation time that would be practical in batch reactors where the TiO2 is applied as a

suspension in the contaminated water and irradiated. Advanced Oxidation Processes

(AOPs) with H2O2/UV typically apply a UV dose of approximately this magnitude or

greater but they must be applied at shorter UV wavelengths according to the adsorption


81

capability of H2O2. A UV dose of 414 mJ/cm2 is also in the lower end of the range of UV

doses applied during H2O2/UV AOPs. It was studied because DOC and DBP fp were

observed to peak around this UV dose associated with an irradiation time of 15 min

during preliminary experiments.

A UV dose of 0 mJ/cm2 corresponds to no irradiation with 1 min of dark adsorption. It

may represent a reactor where TiO2 is employed as an adsorbent and regenerated through

photocatalysis. This measurement may also be useful as a control to compare batch

systems which utilize a new or regenerated mass of TiO2 per unit water irradiated to

treatments with irradiation in a flow through system. However, the pseudo-equilibrium

adsorption of NOM to the TiO2 surface under dark adsorption would be different to the

pseudo-equilibrium adsorption of NOM onto the TiO2 surface in an irradiated reactor

because of factors such as TiO2 surface quality and NOM degradation. Bench scale flow

through CSTRs, thin films, or membrane filters would be more suitable for elucidation of

full scale CSTRs, thin films, or membrane filters, respectively. It is likely though that

these TiO2/UV flow through systems might also benefit from a regeneration step where

the TiO2 filter or thin film is flushed with clean water while being irradiated if

performance decreases from loss of adsorption or treatment capacity similar to

backwashing in a typical water treatment filtration unit.

At the beginning of each experiment testing a TiO2 nanomaterial, the DOC concentration,

UV254 absorbance and DBP fp were determined in the raw synthetic water. Any

variability in each batch of raw synthetic water was accounted for by determining %

reduction. This allowed for direct comparison between TiO2 nanomaterials and UV dose.

All treatments were completed independently in quadruplicate. One replicate was used to

determine chlorine demand for that treatment while the remaining three underwent the

UFC chlorination test. The DOC, UV254, and DBP fp results presented herewith are for

the three replicates used in the DBP fp test.

HAN, chloropicrin, and HK were not detected in any synthetic water samples, either raw

or treated. HAN, chloropicrin and HK are typically present at concentrations an order of

magnitude less than THM and HAA and the synthetic water NOM was not expected to be


82

high in nitrogeneous compounds (HAN and chloropicrin precursors) such as those found

in wastewater influenced source waters.

A two way ANOVA was completed at a 95% confidence level (p < 0.05) with the

statistical program Minitab and the results are provided in Table 5-1. The effects of six

TiO2 types (P25, NB, Ag@SiO2@TiO2/P25, anatase, anatase-N, anatase-B) and five UV

doses (control, 0, 28, 414 and 827 mJ/cm2) on five response variables (% reduction of

UV254 absorbance, DOC concentration, SUVA, THM fp and HAA fp) were determined.

An example ANOVA output table (for the UV254 absorbance response variable) is

provided in the Appendix. For p values less than 0.05 the means were significantly

different between the levels of the factor (TiO2 types or UV dose) or the interaction

between the two factors. UV254 absorbance, DOC concentration, SUVA, and HAA fp all

had p values < 0.05. Thus TiO2 type, UV dose, and the interaction between these two

factors affected the UV254 absorbance, DOC concentration, SUVA, and HAA fp. The R2

value for these response variables was between 0.847 and 0.933 indicating that the

variation seen in these parameters was well correlated to the TiO2 type and UV dose and

there was not a lot of other noise or error in the experiment causing their variation. For

THM fp, the TiO2 type and UV dose significantly affected THM fp while the interaction

between TiO2 type and UV dose was not significant however the R2 value was low

(0.626) indicating noise and other error or factors caused variation in THM fp that may

have overshadowed the effects and interaction of TiO2 type and UV dose.

Table 5-1: Two Way ANOVA Results for Synthetic Water

Treatment Factor UV254 DOC SUVA THM fp HAA fp

TiO2 type p value <0.001 <0.001 <0.001 <0.001 <0.001

UV Dose p value <0.001 <0.001 <0.001 <0.001 <0.001

Interaction p value <0.001 <0.001 <0.001 0.339 <0.001

R2 0.919 0.847 0.933 0.626 0.904


83

5.2 NOM Reduction

Experiments were first performed using P25 and then repeated with the other

nanomaterials. Similarly results here are described for P25 first. More complete data sets

are in the Appendix and what follows are summaries of the main points.

The DOC concentration and UV254 absorbance of synthetic water following P25

treatment is shown in Figure 5-1. P25 adsorbed NOM during the dark adsorption step,

with a 35 % and 12 % decrease in DOC concentration and UV254 absorbance,

respectively. Following the dark adsorption step, the aromaticity of NOM generally

decreased with increasing irradiation with up to 52 % reduction following a UV dose of

827 mJ/cm2. At UV doses of 28 and 414 mJ/cm

2, the DOC concentration in the synthetic

water successively increased suggesting that original NOM or intermediate TiO2/UV

degradation products desorbed from the TiO2 or were more amenable to detection by the

TOC analyzer. At the larger UV dose of 827 mJ/cm2, the DOC concentration in the

synthetic water (2.4 mg/L) was lower compared to the UV dose of 414 mJ/cm2

suggesting that some NOM was mineralized however it was still higher than the DOC

concentration in the synthetic water treated with a UV dose of 28 mJ/cm2 (2.0 mg/L).

Figure 5-1: DOC in Synthetic Water Following P25 TiO2/UV Treatment

0

0.01

0.02

0.03

0.04

0.05

0.06

0.07

0

0.5

1

1.5

2

2.5

3

3.5

4

Control 0 28 414 827

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)

UV Dose (mJ/cm2)

DOC

UV254

Dark

Adsorpiton Only


84

During the Ag@SiO2@TiO2/P25 and anatase-N TiO2/UV experiments, UV254 absorbance

generally decreased following dark adsorption and increasing UV doses while it stayed

relatively unchanged for the anatase-B and NB TiO2/UV experiments (as shown in

Figure 10-7). The UV254 absorbance increased during the anatase experiment and it was

suspected to be due to a contamination since photocatalysis is not expected to

significantly increase the aromaticity of NOM. The affinity between NOM and P25

observed during dark adsorption was also observed with the Ag@SiO2@TiO2/P25 but

was not as prominent with the NB, anatase, anatase-N or anatase-B. The difference in

NOM adsorption and subsequent degradation with the various TiO2 materials may have

been due to differences in the attraction between the TiO2 surface and NOM, the

available surface area, and photoreactivity of the TiO2 material.

The SUVA values (UV254 normalized to DOC concentration) of synthetic water during

the P25 experiment are shown in Figure 5-2. The SUVA value of water following the

dark adsorption step was higher than the raw water control. This can be explained by the

larger decrease in DOC concentration compared to UV254 absorbance, suggesting

compounds other than humics preferentially adsorbed to P25. The synthetic water had a

pH of 8 and at this pH TiO2 and humics are negatively charged and may repel each other

(Liu et al., 2008a). With irradiation, SUVA decreased. During irradiation humics may

have been degraded by TiO2 e-/h

+ pair and by ROS in solution while the process of NOM

mineralization appeared to be slower.

Figure 5-2: SUVA in Synthetic Water Following P25 TiO2/UV Treatment

0

0.5

1

1.5

2

2.5

3

3.5

Control 0 1 15 30

Av

g. S

UV

A (

L/

mg

*m)

UV Dose (mJ/cm2)

Dark Adsorption Only


85

A summary of the % reduction of SUVA with UV dose is shown in Figure 5-3 for all the

TiO2 nanomaterials tested. At a UV dose of 28 mJ/cm2 the effectiveness of the TiO2

nanomaterials was anatase-N = anatase-B > NB > Ag@SiO2@TiO2/P25 = P25 > anatase.

The high negative SUVA % reduction observed in the anatase product was suspected to

be due to some contamination that absorbed UV254. SUVA % reduction remained at just

below 0 % for all treatments with NB. With larger UV doses SUVA % reduction

increased for anatase-N, Ag@SiO2@TiO2/P25, P25 and anatase. At a UV dose of 827

mJ/cm2 the effectiveness of the TiO2 nanomaterials was anatase-N > P25 >

Ag@SiO2@TiO2/P25 > anatase-B > NB > anatase. The SUVA % reduction was

approximately -30 to 20 % at a UV dose of 28 mJ/cm2 and 10 to 50 % at a UV dose of

827 mJ/cm2.

Figure 5-3: Synthetic Water SUVA % Reduction Following TiO2/UV Treatment

5.3 DBP fp Reduction

The disinfection by-product formation potentials (DBP fp) of the water samples were

determined following the Uniform Formation Condition (UFC) chlorination test. A

typical data collection sheet from the UFC chlorination test is appended. The

-100

-80

-60

-40

-20

0

20

40

60

0 200 400 600 800

Av

g. %

Re

du

ctio

n

UV Dose (mJ/cm2)

Anatase N

P25

P25/AgSiO2

Anatase B

NB

Anatase


86

concentrations of the DBPs were calculated using calibration curves prepared with

analytical standards and the area response ratio of DBPs to the internal standard. Plots of

typical calibration curves, complete calibration curve data, and sample calculations are

available in the Appendix. The coefficients of determination (R2) were determined to

ensure subsequent accuracy in DBP quantification. Most calibration curves R2 values

were > 0.99 while a few were slightly below this value. Method detection limits were

determined with most detection limits at 1 to 2 µg/L. Running check standards were

prepared at a frequency of 1 in 10 environmental samples and analyzed alongside the

environmental samples. MDL calculations and quality control charts are also provided in

the Appendix. HAN, HNM, and HK were not detected in any of the synthetic water

samples at quantifiable levels. TCP was not detected in the working standard so a

calibration curve and QA/QC chart could not be prepared for it. This may have been due

to the choice of organic solvent for the working stock solution.

The THM and HAA fp in synthetic water treated with P25 TiO2/UV and chlorination is

shown in Figure 5-4 and Figure 5-5, respectively. Chloroform (trichloromethane -

TCM), dichloroacetic acid (DCAA) and trichloroacetic acid (TCAA) were the only

THMs and HAAs measured in all synthetic water samples. The THM fp in raw synthetic

water was approximately 64 µg/L while the HAA fp was approximately 48 µg/L (sum of

DCAA fp and TCAA fp).

The THM and HAA fp in synthetic water decreased following the dark adsorption step,

inferring that TiO2 adsorbs DBP precursors. Following dark adsorption a low UV dose of

28 mJ/cm2 caused an increased in TCM, DCAA and TCAA fp. Additional irradiation to a

UV dose of 414 mJ/cm2 further increased TCM, DCAA and TCAA fp. As shown in

Figure 5-5, at 414 mJ/cm2 the DCAA fp was approximately the same as the raw water

control and TCAA fp was greater than the raw water control. It can be inferred that the

intermediate products of TiO2/UV may be more reactive with free chlorine than their

parent compounds, particularly with regards to HAA fp. Similar results have been

reported in literature (Daugherty et al., 2011; Philippe, 2010; Gerrity, 2009). At a UV

dose of 827 mJ/cm2 the HAA fp and THM fp was lower when compared to a UV dose of


87

414 mJ/cm2 and it can be derived that some reactive NOM had been degraded to benign

compounds.

Figure 5-4: THM fp in Synthetic Water Following Treatment with P25 TiO2/UV and

Chlorination

Figure 5-5: HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV and

Chlorination

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. C

on

c. (

µg

/L

)

UV Dose (mJ/cm2)

TCMfp

Raw WaterTCMfp

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. C

on

c. (

µg

/L

)

UV Dose (mJ/cm2)

DCAAfp

TCAAfp

Raw waterDCAAfp

Raw waterTCAAfp


88

Complete data sets for the THM and HAA fp in synthetic water following treatment with

the innovative TiO2 nanomaterials are appended as Figure 10-10 and Figure 10-12.

During the NB experiment, the synthetic water THM fp increased during the dark

adsorption step and during irradiation while the HAA fp decreased with TiO2/UV.

Nanobelts decrease electron recombination compared to nanoparticles and this should

increase reactivity however the nanobelts structure has a smaller surface area compared

to nanoparticles, decreasing the available sites for reactions to occur. Also, the nanobelts

were mainly composed of the rutile phase which is less photoreactive than anatase and

the combination of anatase and rutile found in P25. Since DOC and UV254 did not change

significantly it can be inferred that the NB reaction kinetics were slower compared to the

other materials. The THM fp and HAA fp of synthetic water after dark adsorption to the

other nanomaterials decreased similarly to P25. The THM fp with increasing UV doses

tended to decrease rather than peak at 414 mJ/cm2 (as seen with P25). The HAA fp with

Ag@SiO2@TiO2/P25 peaked at 414 mJ/cm2 and was higher compared to the raw water

control, while the HAAfp of anatase and anatase-N also peaked but did not exceed the

raw water HAA fp. Anatase-B treated synthetic water had decreasing HAA fp with

increasing UV doses.

In Figure 5-6 the THM fp was normalized to SUVA value in the water for the P25

TiO2/UV experiment. The THM fp/SUVA was lower than the raw water control for water

treated with dark adsorption and a UV dose of 28 mJ/cm2 and was successively higher

with UV doses of 414 mJ/cm2 and 827 mJ/cm

2 (exceeding raw water control at these two

higher UV doses). Since SUVA is a measure of the relative UV254 absorbance of the DOC

content of the water and not the absolute value of UV254 absorbance, it may not be an

accurate way to predict DBP fp. SUVA also may not accurately predict DBP fp during

TiO2/UV since intermediate TiO2/UV degradation products may be more reactive to

chlorine than original NOM. A graph of HAAfp/SUVA in synthetic water treated with

P25 (Figure 5-7) also exhibits the same trend. Here HAA fp is the sum of the DCAA fp

and TCAA fp.


89

Figure 5-6: THMfp/SUVA in Synthetic Water Following Treatment with P25

TiO2/UV and Chlorination

Figure 5-7: HAAfp/SUVA in Synthetic Water Following Treatment with P25


A similar THMfp/SUVA and HAAfp/SUVA was observed with the

Ag@SiO2@TiO2/P25 and anatase-N experiments. The THMfp/SUVA stayed relatively

constant and lower than the raw water control for anatase and anatase-B experiments.

0

5

10

15

20

25

30

35

40

45

50

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. C

on

c./

SU

VA

(µ

g*m

g*m

/L

2)

Av

g. C

on

c. (

µg

/L

)

UV Dose (mJ/cm2)

TCMfp

TCMfp/SUVA

Raw WaterTCMfpRaw WaterTCMfp/SUVA

0

5

10

15

20

25

30

35

40

45

50

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. C

on

c./

SU

VA

(µ

g*m

g*m

/L

2)

Av

g. C

on

c. (

µg

/L

)

UV Dose (mJ/cm2)

HAAfp

HAAfp/SUVA

Raw waterHAAfpRaw waterHAAfp/SUVA


90

The specific (sp) THM fp (THM fp normalized to DOC concentration) is graphed in

Figure 5-8. As previously mentioned, following the dark adsorption step the THM fp of

the synthetic water decreased. However, the specific THM fp increased with dark

adsorption suggesting that THM precursors were not preferentially adsorbed to the TiO2

surface. This was expected since the aromatic fraction of NOM did not adsorb to the TiO2

surface to the same extent as NOM generally during the dark adsorption step in this

experiment, and these substances are known THM precursors. With irradiation, the

specific THM fp decreases as humics were degraded by ROS and other NOM precursors

were degraded by direct e-/h

+ pair degradation or oxidation by ROS.

Figure 5-8: Sp THM fp in Synthetic Water Following Treatment with P25 TiO2/UV

and Chlorination

The sp HAA fp is graphed in Figure 5-9. Similar to sp THM fp, the sp HAA fp increased

in the synthetic water following dark adsorption. Although TiO2 can be used as an

adsorbent with regeneration by photocatalysis, this may not be the most effective way to

remove DBP precursors with TiO2. The sp HAA fp continued to increase with low UV

doses supporting the correlation of HAA precursors with intermediate TiO2/UV

degradation products in previous studies (Liu et al., 2008b; Daugherty et al., 2011).

0

5

10

15

20

25

30

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. S

pe

cifi

c C

on

c. (

µg

/m

g D

OC

)

Av

g. C

on

c. (

µg

/L

)

UV Dose (mJ/cm2)

TCMfp

SpTCMfp

Raw WaterTCMfp

Raw WaterSpTCMfp


91

Figure 5-9: Sp HAA fp in Synthetic Water Following Treatment with P25 TiO2/UV

and Chlorination

The sp THM and HAA results for synthetic water treated TiO2 was dependent on the type

of TiO2 materials used. Most notably, some materials did not have an increase in sp THM

or HAA upon dark adsorption.

Figure 5-10 provides a graphical representation of the % reduction of THM fp in

synthetic water with the different TiO2 nanomaterials at a UV dose of 28 mJ/cm2 and 827

mJ/cm2. Reduction efficiency at both 28 and 827 mJ/cm

2 for the various nanomaterials

was anatase > anatase-N = P25 = Ag@SiO2@TiO2/P25 > anatase-B > NB. Although

anatase outperformed P25, Ag@SiO2@TiO2/P25 and anatase-N in terms of DBP fp

reduction; P25, Ag@SiO2@TiO2/P25 and anatase-N may have performed better in terms

of HO* production and NOM degradation (as inferred by the greater SUVA reduction in

these products compared to anatase) but DBP reduction was compromised by reactive

intermediates. Generally the nanomaterials (except the nanobelts) reduced THM

precursors by 12 to 32 % at a UV dose of 28 mJ/cm2 and 16 to 42 at a UV dose of 827

mJ/cm2. The % reduction of THM fp increased with the larger UV dose for all the

nanomaterials (except the nanobelts). Raw synthetic water treated with NB TiO2/UV and

0

5

10

15

20

25

30

0

10

20

30

40

50

60

70

80

0 28 414 827

Av

g. S

pe

cifi

c C

on

c. (

µg

/m

g D

OC

)

Av

g. C

on

c. (

µg

/L

)

UV DOse (mJ/cm2)

HAAfp

SpHAAfp

RawwaterHAA

Raw waterSpHAA


92

chlorination had THM fp % increase of approximately 5 and 15 % for UV doses of 28

and 827 mJ/cm2, respectively. NB also did not significantly reduce DOC concentration or

UV254 absorbance and was not included in Figure 5-10.

Figure 5-10: THM fp % Reduction in Synthetic Water Following Treatment with

Various Nanostructured TiO2/UV and Chlorination

Figure 5-11 provides a graphical representation of the % reduction of HAA fp in

synthetic water with the different TiO2 nanomaterials at a UV dose of 28 mJ/cm2 and 827

mJ/cm2. Generally the nanomaterials reduced HAA precursors by 9 to 32 % at a UV dose

of 28 mJ/cm2 and by 9 to 37 % at a UV dose of 827 mJ/cm

2. The % reduction of HAA fp

increased with the larger UV dose for anatase, anatase-N, anatase-B and NB while it

decreased for P25 and Ag@SiO2@TiO2/P25. Reduction efficiency at the UV dose of 28

mJ/cm2 for the various nanomaterials was anatase = Ag@SiO2@TiO2/P25 > P25 >

anatase-N > anatase-B = NB. The % Reduction efficiency at the UV dose of 827 mJ/cm2

was anatase = anatase-N > anatase-B > NB = P25 = Ag@SiO2@TiO2/P25. The treatment

efficiency of P25 and Ag@SiO2@TiO2/P25 at the larger dose may have been

compromised by reactive intermediate products of photocatalysis.

0

5

10

15

20

25

30

35

40

45

50

28 827

Av

g. %

Re

du

ctio

n

UV Dose (mJ/cm2)

P25

P25AgSiO2

Anatase

Anatase-N

Anatase B


93

Figure 5-11: HAA fp % Reduction in Synthetic Water Following Treatment with


5.4 Summary of Results

TiO2/UV was effective for DBP precursor reduction, reducing both THM and HAA

precursors in synthetic water following dark adsorption and low and high UV doses. For

experiments with P25 and Ag@SiO2@TiO2/P25, the HAA fp was larger following the

827 mJ/cm2 UV dose compared to the 28 mJ/cm

2 dose. Otherwise, DBP fp decreased

with the larger UV dose for the remaining nanomaterials. Innovative nanostructured TiO2

materials were effective at reducing DBP precursors. Generally, it appeared that anatase,

anatase-N and Ag@SiO2@TiO2/P25 were comparable to P25 in treatment efficiency,

anatase-B was slightly less effective, and NB was the least effective. The

Ag@SiO2@TiO2/P25 was expected to be at least equal to or better than P25 since it was

largely composed (99 %) of P25. Although these two products did not significantly

outperform the innovative nanomaterials in DBP fp reduction, this was likely do to

reactive intermediates (most notably inferred by the increase in HAA fp with increasing

UV dose observed with P25 and Ag@SiO2@TiO2/P25) and may be reflective of very

good HO* production.

0

5

10

15

20

25

30

35

40

45

50

28 827

Av

g. %

Re

du

ctio

n

UV Dose(mJ/cm2)

P25

P25AgSiO2

Anatase

Anatase-N

anatase b

NB


94

6 EFFECTS OF TIO2/UV ON DBP FORMATION IN A NATURAL

RIVER WATER

The efficiency of TiO2 photocatalysis (TiO2/UV) in the degradation of disinfection by-

product (DBP) precursors in natural river water is presented in this chapter. The

experimental methodology described in Chapter 3 and 5 for synthetic river water

‘synthetic water’ experiments was repeated substituting the lab-prepared synthetic water

with Otonabee River water ‘Otonabee water’.

6.1 Overview of Experiments and Results

Otonabee water was treated with TiO2/UV then chlorinated to determine the effects of

nanostructured TiO2 photocatalysis on THM, HAA, HAN, HNM, and HK formation

potential (fp). Two experimental factors were manipulated; TiO2 material type (6 levels)

and UV dose (5 levels). TiO2 nanomaterials included industry standard Aeroxide® P25

(P25), newly fabricated nanobelts (NB) made from P25, P25 mixed with 1%

Ag@SiO2@TiO2 triplex core-shell photocatalyst (Ag@SiO2@TiO2/P25), sol-gel

synthesized anatase (anatase), sol-gel synthesized anatase doped with 5% nitrogen by

mass (anatase-N), and sol-gel synthesized anatase doped with 5% boron by mass

(anatase-B) developed by project partners (Liang, 2014; Hatat-Fraile, 2014).

The UV doses included 0, 28, 414 and 827 mJ/cm2 and each included 1 min of TiO2 dark

adsorption prior to the UV dose. A raw water control was also included and underwent

the same handling procedures as the other treatments. A UV dose of 28 mJ/cm2 was

representative of a reactor configuration with a short contact time between the TiO2, UV

light and water such as a flow-through TiO2 membrane filter or over-flow thin film. A

UV dose of 827 mJ/cm2 was representative of a reactor configuration with a long contact

time between the TiO2, UV light and water such as a batch reactor, SBR or CSTR. A dark

adsorption time of 1 min was chosen based on preliminary experiments that showed that

a range of dark adsorption times did not significantly affect subsequent NOM degradation

at 28 to 827 mJ/cm2 UV doses. It may be representative of a potential initial mixing

scenario prior to irradiation. A UV dose of 0 mJ/cm2 with dark adsorption only was also


95

representative of a reactor configuration where TiO2 is used as an absorbent and

regenerated through irradiation. The 414 mJ/cm2

dose was included because DOC and

DBP fp peaked at this dose in preliminary experiments and testing this treatment would

potentially give “worst case” concentrations.

The Otonabee water was obtained from the river bank proximate the Peterborough

Utilities Commission (PUC) Water Treatment Plant (WTP) in June of 2014. Enough

water was taken during one sampling event for all the experiments to reduce variability in

raw water quality observed during preliminary experiments. The raw water was filtered

with 0.45 µm Supor® PES membrane filters to remove surface contaminants such as

algae that would be less prevalent at the point of water intake for a drinking water supply

which is typically located a distance away from the river bank and below the water

surface. Any variability in the raw water was accounted for by including a raw water

control for each TiO2 type experiment and NOM and DBP fp % reduction was calculated

for TiO2/UV treatment. This allowed for comparison between TiO2 experiments and with

synthetic water experiments.

Each treatment was completed independently in quadruplicate. Following TiO2 treatment,

the samples were filtered through 0.45µm Supor® PES membrane filters to remove the

TiO2. The DOC concentration and UV254 absorbance were measured as surrogates for

NOM and humics, respectively. One replicate was used to determine the chlorine demand

of that treatment and calculate the required chlorine spike for a 24 hr 1 mg/L chlorine

residual. The remaining three samples were then chlorinated following the Uniform

Formation Condition (UFC) chlorination test which employs a chlorine residual of 1

mg/L after 24 hr at a pH of 8 and temperature of 20 oC. Samples were not buffered by the

natural pH of the water remained within 8 +/- 0.4. The results presented in this chapter

are of the three replicates which underwent the UFC chlorination test.

HANs, chloropicrin, and HKs were not detected in any raw or treated Otonabee water

samples, similar to synthetic water. HAN, chloropicrin and HK are typically present at

concentrations an order of magnitude less than THM and HAA and the Otonabee water


96

NOM was not expected to be high in nitrogeneous compounds (HAN and chloropicrin

precursors) such as those found in wastewater influenced source waters.

A two way ANOVA was completed at a 95% confidence level (p < 0.05) with the

statistical program Minitab® for response variables UV254, DOC, SUVA, THM fp and

HAA fp and the results are provided in Table 6-1 . For p values < 0.05 the means were

significantly different between the levels of the factor (TiO2 type or UV dose) or the

interaction between the two factors. UV254, DOC, SUVA, and HAA fp had p values <

0.05. Thus TiO2 type, UV dose, and the interaction between these two factors affected the

UV254, DOC, SUVA, and HAA fp of the water. The R2 value for these response variables

was between 0.933 and 0.997 indicating that the variation seen in these parameters was

well correlated to the TiO2 type and UV dose and there was not a lot of other noise or

error in the experiment causing their variation. The TiO2 type and UV dose significantly

affected the THM fp mean concentration while the interaction between TiO2 type and UV

dose was not significant however the R2 value was low (0.587) indicating noise and other

error or factors caused variation in THM fp that may have overshadowed the effects of

and interaction between TiO2 type and UV dose.

Table 6-1: Two-way ANOVA for Otonabee Water

Treatment Factor UV254 DOC SUVA THM HAA

TiO2 type - p <0.001 <0.001 <0.001 0.004 <0.001

UV Dose - p <0.001 <0.001 <0.001 <0.001 <0.001

Interaction - p <0.001 <0.001 <0.001 0.179 <0.001

R2 0.997 0.933 0.985 0.587 0.948

6.2 Otonabee Water Quality

The Peterborough Utility Commission (PUC) source water for the Peterborough Water

Treatment Plant (WTP) was the Otonabee River. The PUC provided data on the raw

water quality at their intake and treated effluent (PUC, 2013; City of Peterborough

Environmental Protection Laboratory, 2014). The Otonabee water temperature fluctuated

annually from 0.0 to 28.4 oC with an annual average of 10.5

oC in 2013. The water

treatment process acidifies the raw water pH (approximately 8) and sodium silicate was

added to adjust pH to 7.1. The primary disinfection dose of chlorine for 2013 ranged


97

between 2.3 and 3.1 mg/L where doses at the higher end of the range were added in the

summer when the water temperature was warmer. From May to October, 0.5 mg/L of

chlorine was added at the WTP intake to control zebra mussel growth. Secondary

disinfection was dosed at approximately 0.1 to 0.5 mg/L to the plant effluent to maintain

minimum free chlorine residual in the distribution system of 0.2 mg/L. During the

treatment process at the WTP (coagulation, sedimentation, filtration and disinfection), the

alkalinity decreased by approximately 10 to 20 mg/L as CaCO3 and the DOC also

decreased by approximately 40 %. Raw water DOC ranged from 4.5 to 7.6 mg/L in 2013

and final DOC concentration ranged from 2.4 to 5.0 mg/L. UV254 absorbance was also

measured in 2013 and the average for raw water and plant effluent was 0.07 and 0.02 cm-

1, respectively. The PUC monitored total THM (TCM, TBM, BDCM and DBCM) and

HAA6 (MCAA, DCAA, TCAA, MBAA, DBAA, BCAA) quarterly and reported the

average concentration leaving the WTP and at a point along the distribution system

chosen to represent where concentrations may have been the highest in the city. Reported

values are summarized in Table 6-2. The summary includes the 3rd

quarter sample results

since they were highest and coincided with summer sampling season which was also the

season in which the raw water was taken for the experiment described in this chapter.

Table 6-2: THM and HAA Reported by the PUC (PUC, 2013)

Parameter

2013

Avg.

(µg/L)

2013

Avg.

(µg/L)

Past 10

year Avg.

(µg/L)

3rd

Quarter

(summer

2014) (µg/L)

THM WTP effluent 37.3 37.0 46 60

THM distribution system 74.0 68.5 76.2 125

HAA6 WTP effluent 28 - - 38

HAA6 distribution system 63.8 - - 100

“-“ Not available

6.3 NOM Reduction

Experiments were first performed with P25 then repeated with the other nanomaterials.

What follows here are summaries of the main observations, with P25 results presented

first followed by the other nanomaterials. More complete data sets and QA/QC results are

found in the Appendix.


98

The DOC concentration and UV254 absorbance of Otonabee water following P25

TiO2/UV are shown in Figure 6-1. Raw Otonabee River DOC concentration and UV254

absorbance were higher compared to raw synthetic water. Otonabee River had an

approximate DOC concentration of 5 mg/L while the synthetic water DOC (reported in

Chapter 5) had been about 3 mg/L. The UV254 absorbance of raw Otonabee and synthetic

water was approximately 0.15 and 0.05 cm-1

, respectively. Upon P25 dark adsorption,

DOC concentration and UV254 absorbance of Otonabee water decreased by

approximately 10 to 15 %. Following dark adsorption the UV254 absorbance of the

Otonabee water decreased with increasing UV dose to 0.07 cm-1

, while the DOC

concentration remained relatively constant at about 4.4 mg/L. It can be inferred that

humic substances adsorbed and were degraded by TiO2/UV, either or both by direct

TiO2/UV e-/h

+ pair degradation and reactive oxygen species (ROS) reaction in solution.

Although there did not appear to be DOC reduction with P25 photocatalysis, it is

expected that the interaction between NOM and TiO2/UV caused NOM degradation with

generally lower aromaticity as a result.

Figure 6-1: DOC in Otonabee Water Following P25 TiO2/UV Treatment

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0.18

0

1

2

3

4

5

6

Control 0 28 414 827

Av

g. U

V2

54 A

bso

rba

nce

(1

/cm

)

Av

g. D

OC

Co

nc.

(m

g/

L)

UV Dose (mJ/cm2)

DOC

UV254

Dark Adsorption

Only


99

The DOC concentration and UV254 absorbance of Otonabee water treated with the

various innovative TiO2 nanomaterials are shown in Figure 10-8 appended. The DOC

concentration for Otonabee water treated with NB TiO2/UV did not change significantly

with dark adsorption (a 2 % decrease) or irradiation. There was a 15 % decrease in UV254

absorbance with NB dark adsorption and it remained relatively constant with subsequent

irradiation. For the remaining nanomaterials the UV254 absorbance typically decreased

with dark adsorption and continued to decrease with increasing UV dose. The raw

Otonabee water samples during the anatase experiment had low UV254 absorbance

(average of 0.133 +/- 0.0014 cm-1

) and some experimental error may have thus caused the

dark adsorption and 28 mJ/cm2 UV dosed samples to appear to increase in UV254

absorbance (0.140 +/- 0.0007 cm-1

and 0.136 +/- 0.0004 cm-1

, respectively). Alternatively

the anatase product may have been contaminated with organic compound(s) that may

have leached from the anatase product into the water matrix and caused the UV254

absorbance to increase. The UV254 absorbance was also higher than the control in the

synthetic water treated with anatase. The DOC concentration of the Otonabee water

treated with innovative TiO2 nanomaterials followed a similar trend to P25, an

approximate 10 to 15 % decrease following the dark adsorption step and then relatively

unchanged during subsequent irradiation. Humics and other original NOM compounds

interacted with the innovative TiO2 materials and were degraded to intermediate products

with generally less aromaticity with increasing UV dose from 28 to 827 mJ/cm2.

The SUVA values (UV254 normalized to DOC concentration) of the raw and treated water

during the experiment with P25 are shown in Figure 6-2. The SUVA value of the raw

Otonabee water was approximately double that of raw synthetic water (3 and 1.7

L/mg*m, respectively). The Otonabee water SUVA value decreased slightly with P25

TiO2 dark adsorption and continued to decrease with increasing UV doses. It can be

inferred that aromatic substances (e.g. humics) were preferentially adsorbed and degraded

by P25 TiO2. In the synthetic water (chapter 5) experiment, the SUVA value increased

following the P25 dark adsorption step, and then decreased with irradiation. Although

aromatic substances were not preferentially adsorbed during the dark adsorption step in

this case, they were preferentially degraded by P25 photocatalysis.


100

Figure 6-2: SUVA in Otonabee Water Following P25 TiO2/UV Treatment

Otonabee water SUVA % reductions from treatment with innovative nanostructured

TiO2/UV are provided in Figure 6-3. The rates of reduction (slope of the line) for P25,

Ag@SiO2@TiO2/P25 and anatase were similar but initial and final % reductions are

different. The rates of SUVA % reduction were similar between NB, anatase-N and

anatase-B products, and lower than the other TiO2 products. Ignoring the negative %

reduction found with anatase that is suspected to be a contamination or other

experimental error (also observed in the synthetic water experiment with anatase), the %

reduction of SUVA from the various TiO2 nanomaterials studied was approximately 0 to

14 % and 10 to 45 % after a UV dose of 28 and 827 mJ/cm2, respectively. At a UV dose

of 28 mJ/cm2 the effectiveness of the TiO2 nanomaterials was P25 = NB > anatase-N >

Ag@SiO2@TiO2/P25 > anatase-B > anatase. At a UV dose of 827 mJ/cm2 the

effectiveness of the TiO2 nanomaterials was P25 > Ag@SiO2@TiO2/P25 > anatase-N =

anatase = NB > anatase-B.

0

0.5

1

1.5

2

2.5

3

3.5

4

Control 0 28 414 827

Av

g. S

UV

A (

L/

mg

*m)

UV Dose( mJ/cm2)

Dark Adsorption Only


101

Figure 6-3: SUVA % Reduction in Otonabee Water Following TiO2/UV Treatment

The interaction between NOM and the TiO2 surface in the dark and upon irradiation is

affected by many factors. The TiO2 surface and humics are both negatively charged at a

pH of 8 and may generally repel each other. Also, TiO2 may exhibit increased

hydrophilic properties upon irradiation thus repelling hydrophobic compounds such as

humics and aromatic substances. However, large molecules like humics contain many

functional groups that may individually interact with the TiO2 surface. Water chemistry

such as pH and the presence of ions and molecules in solution and the surface character

of the TiO2 affect the attraction between NOM and TiO2 as well as the production of

ROS. These and other factors such as UV dose and reactor configuration affect NOM

degradation by e-/h

+ pairs and ROS and reduction by adsorption. This explains the varied

results in literature, preliminary experiments, and experiments with synthetic and

Otonabee water using different nanostructured TiO2 materials.

6.4 DBP Reduction

The disinfection by-product (DBP) formation potentials (fp) of the Otonabee River raw

and treated water samples were determined post chlorination according to the same

-40

-30

-20

-10

0

10

20

30

40

50

60

0 200 400 600 800

Av

g. %

Re

du

ctio

n

UV Dose (mJ/cm2)

P25

AgSiO2/P25

Anatase N

Anatase

NB

Anatase B


102

methodology discussed in Chapter 3 and 5 to allow for comparison between treatment

efficiencies of the various nanostructured TiO2 materials in a model and natural river

water. Calibration and QA/QC data is appended. HANs, chloropicrin, and HKs were not

detected in any of the Otonabee water samples. Chloroform, dichloroacetic acid (DCAA)

and trichloroacetic acid (TCAA) were the only THMs and HAAs measured in all

Otonabee water samples, similar to the synthetic water samples. In the results that follow,

the THM reported concentrations are solely from chloroform while the HAA

concentrations are from DCAA and TCAA only.

As expected because of higher DOC and UV254, raw Otonabee water had higher THM fp

(98.9 +/- 5.0 µg/L to 114.0 +/- 1.8 µg/L between experiments) compared to raw synthetic

water (45.8 +/- 1.1 µg/L to 63.6 +/- 2.5 µg/L between experiments). The HAA fp of the

Otonabee water (88.4 +/-1.9 µg/L to 106 +/- 1.2 µg/L between experiments) was also

higher than the synthetic water (46.7 +/- 3.3 µg/L to 58.5 +/- 3.1 µg/L between

experiments). In both the raw synthetic and Otonabee waters the range of THM fp was

similar to but slightly higher than the HAA fp range.

The THM fp and HAA fp in Otonabee water treated with P25 TiO2/UV are presented

first followed by discussion comparing the innovative TiO2 nanomaterials. The Otonabee

water THM and HAA fp following treatment with P25 TiO2/UV and chlorination are

shown in Figure 6-4 and Figure 6-5, respectively. The THM and HAA fp was lower in

the samples treated with TiO2 dark adsorption and no irradiation compared to the raw

water control, inferring that TiO2 adsorbs THM and HAA precursors. Following dark

adsorption, a low UV dose of 28 mJ/cm2 increased TCM, DCAA and TCAA fp.

Additional irradiation to a UV dose of 414 mJ/cm2 further increased TCM, DCAA and

TCAA fp.

As shown in Figure 6-5, at 414 mJ/cm2 the DCAA and TCAA fp were greater than the

raw water control. As discussed in Chapter 2, the degradation of NOM follows a

sequence of successively smaller organic molecules with lower aromaticity until

mineralization occurs or recalcitrant compounds remain. The degradation products

interact with the TiO2 surface differently than parent compounds. THM and HAA fp may


103

have increased at the 28 and 414 mJ/cm2 UV doses due to desorption of DBP precursors

from the TiO2 surface and the degradation of benign NOM into reactive intermediates.

THMs are typically associated with hydrophobic precursors like humics, while HAAs are

typically associated with both hydrophobic and hydrophilic precursors. The more

substantial increase in DCAA and TCAA fp compared to TCM fp at the low UV doses

may be due this. With additional irradiation up to a UV dose of 827 mJ/cm2

the TCM fp

decreased to below the TCM fp of the water with UV doses of 28 and 414 mJ/cm2

inferring that some TCM precursors were degraded to benign compounds. The DCAA fp

increased further with additional irradiation up to 827 mJ/cm2 while the TCAA fp

decreased. It is expected that DCAA would have eventually decrease with additional

irradiation as reactive intermediate products would have been degraded to benign

compounds.

Figure 6-4: THMfp in Otonabee water Following Treatment with P25 TiO2/UV and

Chlorination

0

20

40

60

80

100

120

0 28 414 827

Av

g. C

on

c. (

µg

/L

)

UV Dose(mJ/cm2)

TCMfp

Raw WaterTCMfp


104

Figure 6-5: HAA fp in Otonabee water Following Treatment with P25 TiO2/UV and

Chlorination

Complete data sets for the different TiO2/UV nanomaterials are available in the

Appendix. The decrease in Otonabee water THM fp from P25 TiO2 dark adsorption was

seen with all the nanomaterials however only the Ag@SiO2@TiO2/P25 had a similar %

reduction (approximately 20 % reduction) while the other nanomaterials had less (5 to 13

% reduction). The Ag@SiO2@TiO2/P25 THM fp trend with UV doses was similar to

P25. The NB and anatase treated Otonabee water THM fp did not change significantly

between dark adsorption and low UV doses but did decrease with the higher 827 mJ/cm2

UV dose. Anatase-B and anatase-N treated Otonabee water had a decrease in THM fp

with dark adsorption and a UV dose of 28 mJ/cm2 but then the THM fp increased with

the larger UV doses.

The HAA fp in Otonabee water treated with the innovative TiO2 nanomaterials followed

the same trends as treatment with P25. Except for the nanobelts, all TiO2 nanomaterials

adsorbed HAA precursors and lowered the HAA fp in Otoanbee River water following

dark adsorption. The increase in HAA fp with NB dark adsorption may be due to

contamination since the lack of photocatalysis should have kept existing Otonabee River

0

20

40

60

80

100

120

0 28 414 827

Av

g.

Co

nc.

(µ

g/

L)

UV Dose (mJ/cm2)

DCAAfp

TCAAfp

Raw waterDCAAfp

Raw waterTCAAfp


105

NOM unaltered. Treated with the innovative TiO2 nanomaterials and UV doses of 28 and

414 mJ/cm2 the Otonabee River HAA fp increased with increasing UV dose and at 414

mJ/cm2 exceeded or was very proximate to the fp of the untreated water. At a UV dose of

827 mJ/cm2 the Otonabee water HAA fp either increased or decreased slightly compared

to the 414 mJ/cm2 dose.

In Figure 6-6 the P25 TiO2/UV treated Otonabee water THM fp is normalized to SUVA

value. The same THMfp/SUVA trend was observed with Otonabee water and synthetic

water treated with P25. The THM fp/SUVA was lower than the raw water control for

water treated with dark adsorption and a UV dose of 28 mJ/cm2. The THMfp/SUVA was

successively higher with UV doses of 414 mJ/cm2 and 827 mJ/cm2 and in both cases

exceeded the raw water control THMfp/SUVA. Although humic substances may be

degraded to compounds with less aromaticity, these and other (intermediate degradation)

compounds may be THM fp precursors. The Otonabee water treated with

Ag@SiO2@TiO2/P25 followed a similar THMfp/SUVA trend as with P25 however for

the remaining TiO2 nanomaterials the THMfp/SUVA fluctuated slightly above or below

the raw water control value.

Figure 6-7 displays the HAAfp/SUVA in Otonabee water treated with P25 TiO2/UV and

chlorination. Similar trends were present in HAAfp/SUVA compared to the

THMfp/SUVA and HAAfp/SUVA trends in synthetic water treated with P25. Except for

the NB experiment, the Otonabee River treated with innovative TiO2 nanomaterials

followed the same HAAfp/SUVA trend as with P25. The NB treated Otonabee water

HAAfp/SUVA increased upon dark adsorption and remained relatively stable with UV

doses.


106

Figure 6-6: THMfp/SUVA in Otonabee Water Following Treatment with P25


Figure 6-7: HAAfp/SUVA in Otonabee Water Following Treatment with P25


0

10

20

30

40

50

60

70

0

20

40

60

80

100

120

0 28 414 827

Av

g. C

on

c./

SU

VA

(µ

g*m

g*m

/L

2)

Av

g. C

on

c. (

µg

/L

)

UV Dose(mJ/cm2)

TCMfp

TCMfp/SUVA

Raw WaterTCMfp

Raw WaterTCMfp/SUVA

0

10

20

30

40

50

60

70

0

20

40

60

80

100

120

0 28 414 827

Av

g. C

on

c./

SU

VA

(µ

g*m

g*m

/L

2)

Av

g.

Co

nc.

(µ

g/

L)

UV Dose (mJ/cm2)

HAAfp

HAAfp/SUVA

Raw waterHAAfp

Raw waterHAAfp/SUVA


107

The specific (sp) THM fp and sp HAA fp (THM fp and HAA fp normalized to DOC,

respectively) are shown in Figure 6-8 and Figure 6-9, respectively. The sp THM fp of

Otonabee water remained relatively unchanged in the P25 treated samples and samples

treated with innovative materials, fluctuating approximately +/- 2 µg/mg. The sp HAA fp

of Otonabee water treated with P25 and innovative TiO2 nanomaterials exhibited an

increasing trend with increasing UV dose. These results suggest that the intermediate

degradation products of NOM by TIO2/UV were more likely to react with chlorine to

form HAAs than THMs.

Figure 6-8: THMfp/DOC in Otonabee Water Following Treatment with P25


Figure 6-9: HAAfp/DOC in Otonabee Water Following Treatment with P25


0

5

10

15

20

25

30

35

40

0

20

40

60

80

100

120

0 28 414 827A

vg

. Co

nc.

/ S

UV

A

(µg

*mg

*m/

L2)

Av

g. C

on

c. (

µg

/L

)

UV Dose(mJ/cm2)

TCMfp

TCMfp/DOC

Raw WaterTCMfpRaw WaterTCMfp/DOC

0

5

10

15

20

25

30

35

40

0

20

40

60

80

100

120

0 28 414 827

Av

g. C

on

c./

SU

VA

(µ

g*m

g*m

/L

2)

Av

g.

Co

nc.

(µ

g/

L)

UV Dose (mJ/cm2)

HAAfp

HAAfp/DOC

Raw waterHAAfpRaw waterHAAfp/DOC


108

Figure 6-10 provides a graphical representation of the % reduction of THM fp in

Otonabee water treated with P25 and innovative TiO2 nanomaterials as a UV dose of 28

and 827 mJ/cm2. Generally there was less reduction and more variability in the Otonabee

water compared to synthetic water. The nanomaterials reduced THM fp by approximately

5 to 23 % when irradiated with a UV dose of 28 mJ/cm2 and 4 to 25 % when irradiated

with a UV dose of 827 mJ/cm2. The % reduction for Ag@SiO2@TiO2/P25 and anatase-N

remained relatively constant between the two UV doses, while the % reduction for P25,

NB, and anatase increased and the % reduction for anatase-B decreased. % Reduction

efficiency at the 28 mJ/cm2 UV dose was anatase-B > P25 = anatase-N = anatase > NB =

Ag@SiO2@TiO2/P25. The % reduction efficiency at the 827 mJ/cm2 UV dose was P25 =

anatase = NB = anatase-N > Ag@SiO2@TiO2/P25 = anatase-B. The treatment

efficiencies of these TiO2 nanomaterials with respect to DBP fp may not be reflective of

their AOP capabilities due to reactive intermediate products and precursor adsorption to

the TiO2 surface.

Figure 6-10: THM fp % Reduction in Otonabee Water Following Treatment with


0

5

10

15

20

25

30

35

28 827

Av

g. %

Re

du

ctio

n

UV Dose (mJ/cm2)

P25

NB

AgSiO2/P25

Anatase

Anatase-N

Anatase B


109

Figure 6-11 provides a graphical representation of the % reduction of HAA fp in

Otonabee water treated with P25 and innovative TiO2/UV at UV doses of 28 mJ/cm2 and

827 mJ/cm2. Apart from the NB, the smaller UV dose reduced HAA precursors

(approximately 2 to 13 %). With the UV dose of 827 mJ/cm2, the intermediate products

of photocatalysis contributed to HAA fp with an increase in HAA fp of 1 to 23 %

approximately. This was markedly different than the HAA fp in synthetic water after

TiO2/UV with the 827 mJ/cm2 UV dose which either increased or decreased (based on

TiO2 type) compared to the lower 28 mJ/cm2 UV dose but still remained lower than the

raw water control. The effectiveness of these AOPs was compromised by the formation

of reactive intermediate products from benign original NOM. The effectiveness of the

various TiO2 nanomaterials at reducing HAA fp with a UV dose of 28 mJ/cm2 was

Ag@SiO2@TiO2/P25 = anatase-N > anatase = P25 >anatase-B > NB. For the UV dose

of 827 mJ/cm2 the % reduction efficiency (determined based on lowest negative %

reduction) was anatase-N = anatase-B > NB = anatase = Ag@SiO2@TiO2/P25 > P25.

Figure 6-11: HAA fp % Reduction in Otonabee Water Following Treatment with


-30

-25

-20

-15

-10

-5

0

5

10

15

20

28 827

Av

g. %

Re

du

ctio

n

UV Dose(mJ/cm2)

P25

NB

AgSiO2/P25

Anatase

Anatase-N

anatase b


110

The PUC Water Treatment Plant (WTP) reported THM and HAA concentration of 125

µg/L (worst case scenario) and 100 µg/L (worst case scenario), respectively during the

summer 2014 sampling event (coinciding to the approximate time of water sampling for

these experiments). The PUC secondary chlorination procedures (pH, temperature, 24 hr

chlorine residual) were not the same as the UFC chlorination test followed in these

experiments. Also, water quality parameters such as DOC and UV254 in the PUC WTP

effluent were different compared the raw Otonabee water used in the experiments.

Nonetheless, the THM and HAA concentrations at the worst case location in the

distribution system during the summer sampling event were similar to those found in the

raw Otonabee water. Generally, inferences can be made that the PUC WTP effluent

would experience reduction in THM and HAA during chlorination it had TiO2/UV as an

additional unit treatment process prior to chlorination, as long as the system was designed

according to UV dose and TiO2 type to optimize DBP precursor reduction.

6.5 Comparison of Results for Synthetic and Natural Waters

As expected, DOC concentration, UV254 absorbance, THM fp and HAA fp were higher in

raw Otonabee water compared to raw synthetic water. Also, % reduction was generally

lower in Otonabee water compared to synthetic water. The % DBP fp reductions for

Otonabee and synthetic water are provided in Figure 6-12. Otonabee water DBP fp %

reduction following TIO2/UV may have been lower due its higher concentration of NOM

and DBP precursors. The effects of NOM concentration on adsorption and reaction rates

are described by the Langmuir-Hinshelwood (L-H) adsorption model, which also

accounts for differences in water quality parameters that can affect adsorption kinetics

and subsequent degradation rates. Water quality parameters such as pH and alkalinity,

can affect the character of NOM on the TiO2 surface as well as the production of reactive

oxygen species. The pH of the Otonabee and synthetic waters were similar

(approximately 8) while the alkalinity of the synthetic and Otonabee water was 117 and

84 mg/L as CaCO3, respectively. Higher alkalinity may increase radical formation which

may be another cause in the higher DBP fp % reduction in synthetic water compared to

Otonabee water.


111

Figure 6-12: TiO2/UV Treatment Comparison between Otonabee and Synthetic

Water

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

P25

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

AgSiO2/P25

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

NB

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

Anatase

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

Anatase-N

-30

-20

-10

0

10

20

30

40

50

THM28

THM827

HAA28

HAA827

Avg

. %

Re

du

ctio

n

UV Dose (mJ/cm2)

Anatase-B


112

The DBP fp % reduction efficiencies between the TiO2 nanomaterials studied is also

shown in Figure 6-12 and the data is summarized in Table 6-3. The P25 and

Ag@SiO2@TiO2/P25 products are plotted side by side to compare these two products

since the Ag@SiO2@TiO2/P25 was largely composed of P25. The Ag@SiO2@TiO2/P25

treatment efficiency was similar to P25, and sometimes performed better than or worse

than P25 in terms of DBP fp. Next anatase and anatase-N are plotted side by side. For the

most part, anatase performed better than anatase-N. Also, the anatase product often

performed better than P25. The treatment efficiencies of these various nanomaterials

based on DBP fp may not be reflective of their performance as AOPs due to the

formation of reactive intermediate products from benign original NOM and adsorption of

NOM onto the TiO2 surface. By examining the changes to DOC concentration, UV254

absorbance and DBP fp during the dark adsorption step and with subsequent and

increasing UV doses, it was possible to generally infer the degradation of original

precursors, formation of new precursors and effects of adsorption to the TiO2 surface.

From these observations P25, Ag@SiO2@TiO2/P25, anatase and anatase-N appear to be

effective AOPs and mechanisms for DBP precursor reduction.

The last two graphs in Figure 6-12 show the NB and anatase-B products that had the

lowest DBP % reductions overall. These materials also generally provided lower DOC

concentration and UV254 absorbance % reductions. The efficiency of the NB most likely

suffered from the fabrication process that altered its crystalline structure from that of P25.

The anatase-B product, although it was fabricated to lower its bang gap energy, had the

same band gap as the anatase.

Experimental results of THM and HAA % reduction with P25 TiO2/UV in the current

research and from TiO2/UV studies by other researchers are summarized in Table 6-4.

Table 6-4 is a general comparison between the current research and research conducted

by other groups since factors such as TiO2 concentration, type and configuration, light

source (wavelength and fluence rate), UV dose, water quality, varied between

experiments.


113

Table 6-3: Summary of DBP % Reduction with TiO2/UV

DBP

(µg/L)

UV Dose

(mJ/cm2)

Water P25 Ag@SiO2

@TiO2/P25 NB Anatase Anatase-N Anatase-B

THM

28 Synthetic 21.8 +/- 3.2 19.7 +/- 2.1 -5.2 +/- 1.7 31.9+/-5.9 23.4+/-3.6 11.9+/-0.5

Otonabee 17.2+/-1.7 5.1+/-1.2 5.7+/-3.0 14.9+/-1.2 14.9+/-9.3 22.5+/-6.7

827 Synthetic 24.6 +/- 1.1 22.0 +/-0.7 -13.7 +/- 8.0 41.4+/-2.9 29.6+/-8.4 16.2+/-4.2

Otonabee 24.4+/-3.4 5.1+/-3.5 14.2+/-7.4 18.0+/-8.9 14.9 3.7+/-1.7

HAA

28 Synthetic 17.8+/-2.4 31.4+/-4.0 9.4+/-2.6 32.0+/-1.8 14.9 11.7+/-1.6

Otonabee 6.2+/-2.2 12.6+/-3.7 -8.2+/-1.0 6.3+/-1.7 12.3+/-0.9 1.6+/-1.2

827 Synthetic 13.2+/-2.7 9.0+/-2.7 15.1+/-1.5 35.8+/-1.2 32.4+/-5.5 19.3

Otonabee -23.4+/-3.9 -9.6+/-1.2 -6.7+/-2.7 -8.6+/-1.4 -0.6+/-7.2 -3.6+/-2.2

*Only 1 replicate if no standard deviation given

*Negative % reduction (i.e. increase in concentration) is bolded


114

Table 6-4: Comparison of THM and HAA fp % Reduction in TiO2/UV Studies

Reference

TiO2

Conc.

(g/L)

TiO2

Type

TiO2

Configuration

UV Light

Source /

Wavelength

UV Dose: Calculated

Equivalent (Reported as) Water Source

THM fp

% Reduction

HAA fp

%

Reduction

Currenta,b

Research 0.1 various

Batch

suspension

Solar Simulator/

300-424 nm

4.6 J/L (28 mJ/cm2 or 1 min

irradiation x 0.46 mW/cm2)

0.14 KJ (827 mJ/cm2 or 30

min irradiation)

Otonabee River,

ON

5 to 23%

4 to 24%

-8 to 13%

-23 to -1 %

Currenta,b

Research 0.1 various

Batch

suspension

Solar Simulator/

300-424 nm

4.6 J/L (28 mJ/cm2)

0.14 KJ (827 mJ/cm2)

Synthetic water

(Suwannee NOM)

-5 to 32 %

-14 to 41%

9 to 32 %

9 to 36%

Liu et al,

2008a 0.1 P25

Annular photo-

reactor 365 nm

12 KJ/L (20.3 µE/L*s, for 30

min irradiation)

Fluka Humic

Acid solution 42%

Liu et al.,

2008b 0.1 P25

Annular photo-

reactor 365 nm

12 KJ/L

(20.3 µE/L*s for 30 min)

Myponga

Reservoir, AU

43%

SpTHMfp:

13%

13%

SpHAAfp:

-29%

Liu et al.,

2010 0.1 P25

Annular photo-

reactor 365 nm

12 KJ/L

(20.3 µE/L*s for 30 min)

Australian

Surface Water

65%

SpTHMfp

6%

-

Gerrity,

2009 0.1 P25

Photo-Cat Lab®

Purifics®

254 nm (LP

lamp)

1.1 KJ/L

(0.3 kWh/m3)

Salt and Colorado

River

1 and -45%

reductions

-

Mori,

2013

0.3 µm

thick-ness

Sol-gel

Anatase Thin Film 350 nm peak

17 KJ/L (2 mW/cm2

at surface with 100cm2, 500

mL, 12 hr)

Swamp waters,

Japan 59-62 % 55-73%

Philippe,

2010 1 P25

Annular photo-

reactor

MP lamp

(630 W)

0.1 L,

1 min

glycine, D-

Mannose,

Resorcinol, tannic

acid

-800%, 25%,

95%, 92%,

respectively

-

Kent et al.,

2011b

0.001 P25 Batch

suspension 254 nm (LP) (50 mJ/cm

2) French river water 20% 90%

Daugherty

et al.,

2011

1 P25 PhotoCAT®

Lab, Purifics® 254 nm (LP)

18 KJ/L

(5 kWh/m3)

WTP settled

water, AZ

-106%

spTHMfp:

-153%

-132%

spHAAfp:

-200% aUV Dose normalized to water depth (0.459 mW/cm

2; at water surface intensity was 13.4 mW/cm

2)

bUV Fluence rate calculated accounting for water, divergence, reflection, petri factors

“-“ not measured


115

The experiment by Kent et al. applied a similar corrected UV dose (50 mJ/cm2) to the

current research at (28 mJ/cm2) and they had similar THM fp % reduction (20 %) but

much larger HAA fp % reduction (90 %) compared to the P25 results in the current

research for the 28 mJ/cm2 dose. Liu et al., worked with P25 at a concentration of 0.1 g/L

in suspension similar to the current research and following 30 min of irradiation had

higher THM fp % reduction (42 to 65 %) and similar HAA fp % reduction (13 %)

compared to the P25 experiment with synthetic water following 30 min of irradiation

(827 mJ/cm2 UV dose) in the current research (25 % and 13 %), respectively. The range

in THM and HAA fp % reduction reported in literature is large (-800 to 90 % reduction

for the experiments listed in Table 6-4) and although there is potential for TiO2/UV to be

a very effective DBP control strategy, planning in the form of bench and pilot scale

testing would be advised to optimize precursor reduction based on source water quality.

6.6 Summary of Results

TiO2/UV effectively reduced DBP precursors in source water and performance efficiency

was affected by the TiO2 type, water quality, DBP class, and UV dose. Generally, P25,

Ag@SiO2@TiO2/P25, anatase, and anatase-N had comparable % reductions (up to 41 %

reduction), anatase-B had lower overall % reductions (up to 23 %), and NB had the

lowest % reductions (up to 15 %). Generally there were larger % reductions in DBP

precursors in synthetic water compared to Otonabee water and there were greater %

reductions in THM precursors compared to HAA precursors. The efficiency between

low and high UV doses (28 and 827 mJ/cm2, respectively) was dependent on DBP class.

Typically there were higher THM % reductions following the higher UV dose. The

Otonabee water HAA fp increased with the larger UV dose for all nanomaterials while

the synthetic water HAA fp saw increases in HAA fp between the low and high UV doses

during the P25 and Ag@SiO2@TiO2/P25 experiments. The increase in HAA fp observed

may have been caused by reactive intermediate products of TiO2/UV; a crucial factor to

consider in the design of TiO2/UV processes for DBP precursor reduction. Less treatment

may actually be better for DBP control, unless a sufficiently large dose is applied to

degrade reactive intermediates.


116

7 CONCLUSIONS

Disinfection byproducts (DBPs) are the inadvertent result of drinking water disinfection

since natural organic matter (NOM) is ubiquitous in natural water systems used as a

source for drinking water and NOM contains precursors to DBPs. Chlorine remains

popular for primary and secondary disinfection of drinking water because it is reliable,

safe to handle, and cost effective. Coagulation and adsorption in conventional water

treatment systems, and membrane filtration and AOPs in newer systems employed for

DBP precursor control prior to chlorination, may be successfully augmented or replaced

by TiO2/UV. There are numerous potential advantages to TiO2/UV including:

concurrent disinfection,

degradation of recalcitrant compounds,

no chemical inputs required (other than a capital investment of TiO2),

ability to use UVA or solar energy where other AOPs such as UV/H2O2 require

wavelengths < 300 nm (UVB and UVC light),

no waste products formed (NOM degraded rather than separated),

can operate at ambient temperature and pressure, and natural drinking water pH.

Preliminary “proof-of-concept” experiments showed that water quality, TiO2 type,

concentration and configuration, and UV dose impacted DBP formation and served to

initiate the comparisons of impacts from innovative TiO2 nanomaterials, low and high

UV doses and water source that were examined in later experiments. The trihalomethane

(THM) and haloacetic acid (HAA) formation potential (fp) of treated water generally

decreased with increasing irradiation time and TiO2 concentration; although some

intermediate TiO2/UV degradation products may have been more reactive to chlorine

than parent compounds. Industry standard Aeroxide® P25 in suspension performed the

best from the configurations tested. The highest DBP reductions were observed at the

longest irradiation times (up to 87 % reduction). An optimal TiO2 concentration of 0.1

g/L was determined based on the analysis of DOC concentration and UV254 absorbance

reduction using pseudo first order reaction kinetics for TiO2 concentrations between

0.005 and 0.5 g/L. An optimal TiO2 dark adsorption time of 1 min was chosen based on


117

potential full scale reactor design constraints since short TiO2 dark adsorption times did

not appear to affect the extent of DOC concentration and UV254 absorbance reduction

during subsequent irradiation.

Subsequently, simulated solar photocatalytic experiments with industry standard P25 and

five innovative TiO2 nanomaterials (nanobelts, P25 mixed with silver, anatase, nitrogen-

doped anatase, and boron-doped anatase) were completed in batch reactors with a TiO2

concentration of 0.1 g/L in suspension using model and natural river water sources. The

overall hypotheses were that TiO2/UV would either increase or decrease DBP fp, and that

different types of TiO2 would have varying degrees of effect. These experiments also

employed both short-term exposures (UV dose of 28 mJ/cm2) representative of flow-

through treatment systems, and longer term exposures (UV dose of 827 mJ/cm2) that may

be more representative of batch reactor systems. The following may be considered to be

the major conclusions from this work:

• The relative formation of chlorinated, brominated, and nitrogeneous DBPs was

expected considering the low bromide and nitrogen levels in the waters tested.

Chloroform (trichloromethane – TCM) was the only THM detected and dichloroacetic

acid (DCAA) and trichloroacetic acid (TCAA) were the only HAAs detected.

Monochloroacetic acid quickly converts to DCAA or TCAA when exposed to elevated

concentration of chlorine so its absence was expected. HAN, HK and HNM were not

detected in any raw or treated samples.

• Some THM and HAA precursors adsorbed to TiO2, however subsequent irradiation

showed some potential to increase DBP fp. Additional precursors may have been

original NOM or intermediate degradation products that desorbed from the TiO2

surface or products of NOM oxidation by reactive oxygen species in solution. The

DBP fp at a low UV dose (28 mJ/cm2) was lower than, equal to, or greater than a high

UV dose (827 mJ/cm2) depending on water source, TiO2 type and DBP class.

Typically there were higher THM % reductions following the higher UV dose while

HAA fp had a tendency to increase with the larger UV dose.

• Generally, synthetic water exhibited a higher % reduction of DBP fp compared to

Otonabee water. Synthetic water experienced THM % reduction up to 41 % and HAA


118

% reduction up to 36 %. Otonabee River THM fp was reduced by up to 24 % and

HAA fp was reduced by up to 13%. Generally there were greater THM precursor

reductions compared to HAA precursor reductions. The results of the current research

were comparable to results published in literature.

• The performance of the nanostructured TiO2 materials was dependent on the response

variable (DOC, UV254, THM, HAA) water source (synthetic or Otonabee water) and

UV dose (0, 28, 414 or 827 mJ/cm2). Generally, P25, Ag@SiO2@TiO2/P25, anatase,

and anatase-N had comparable overall % reduction (up to 41 % reduction), anatase-B

had lower overall % reductions (up to 23 %), and NB had the lowest % reductions (up

to 15 %).

Although early in their testing, the results of this research were encouraging in regards to

the potential use of these alternative TiO2 materials for water treatment. Advancement in

treatment efficiencies emerging from innovations in material science and reactor design

continue to increase the feasibility of full scale systems.


119

8 RECOMMENDATIONS

TiO2 photocatalysis is a promising technology being researched today for disinfection

byproduct management in drinking water treatment Future research recommendations to

advance the body of knowledge of the effects of TiO2/UV on DBP formation include

possible investigations into:

Impacts of TiO2 type (quantum efficiency, visible light sensitivity, photo-

reactivity, surface area, adsorption),

Reactor configuration (membrane flow through filter, continuously stirred tank

reactor, thin films, batch),

AOP treatment efficiency based on hydroxyl radical production (using probe

compound)

Water quality impacts (alkalinity, pH, NOM fractions),

UV dose (low, intermediate, and high)

DBP class (nitrogenous species, re-emerging DBPs of concern),

TiO2 regeneration by irradiation and/or cleaning for reuse in multiple batch

treatments or for periodic regeneration in flow-through systems

Pilot scale experiments incorporating TiO2/UV within treatment process train.

The results of the current research showed that TiO2/UV compares favorably to other

DBP control management strategies that rely on DBP precursor reduction and are

consistent with similar studies reviewed in literature. As expected, the TiO2/UV

treatment systems tested were affected by factors such as water quality, TiO2 type and

concentration, reactor configuration, UV dose, and contaminant of interest. Moving

forward, bench and pilot scale testing will likely be needed to determine optimal

operational parameters for full-scale implementation.


120

9 REFERENCES

Aieta, E.M., Berg, J.D., 1986. A review of chlorine dioxide in drinking water treatment.

J. Am. Water Work. Assoc. 78, 62–72.

American Public Health Association (APHA), American Water Works Association

(AWWA), Water Environment Federation (WEF), 2012. Standard Methods for the

Examination of Water and Wastewater, 22nd ed. Baltimore, Maryland, USA.

Amy, B.G.L., Alleman, B.C., Cluff, C.B., 1990. Removal of dissolved organic matter by

nanofiltration. J Environ. Eng. 116, 200–205.

Azouani, R., 2009. Elaboration de Nouveaux Nanomatériaux Photocatalytiques Actifs

sous Rayonnement Visible. Université Paris 13.

Bekbolet, M., Uyguner, C.S., Selcuk, H., Rizzo, L., Nikolaou, A.D., Meric, S., Belgiorno,

V., 2005. Application of oxidative removal of NOM to drinking water and formation

of disinfection by-products. Desalination 176, 155–166.

Bolton, J.R., 2001. Ultraviolet Applications Handbook, 2nd ed. Photosciences Inc.,

Edmonton, AB, Canada.

Bolton, J.R., 2004. Germicidal Fluence (UV Dose) Calculations for a Low Pressure UV

Lamp. Bolton Photosciences Inc., Edmonton, Alberta, Canada.

Bolton, J.R., Cotton, C., 2008. The Ultraviolet Disinfection Handbook, 1st ed. American

Water Works Association, Denver, CO, USA.

Bolton, J.R., Linden, K.G., 2003. Standardization of methods for fluence (UV dose)

determination in bench-scale UV experiments. J. Environ. Eng. 209–215.

Bond, T., Goslan, E.H., Jefferson, B., Roddick, F., Fan, L., Parsons, S.A., 2009.

Chemical and biological oxidation of NOM surrogates and effect on HAA formation.

Water Res. 43, 2615–22.


121

Bond, T., Huang, J., Templeton, M.R., Graham, N., 2011. Occurrence and control of

nitrogenous disinfection by-products in drinking water - a review. Water Res. 45,

4341–54.

Buchanan, W., Roddick, F., Porter, N., 2006. Formation of hazardous by-products

resulting from the irradiation of natural organic matter: comparison between UV and

VUV irradiation. Chemosphere 63, 1130–41.

Chawla, M., 2014. Personal Communications. Photo Emission Tech. Inc., Camarillo, CA,

USA.

Chen, B., Lee, W., Westerhoff, P.K., Krasner, S.W., Herckes, P., 2010. Solar photolysis

kinetics of disinfection byproducts. Water Res. 44, 3401–9.

Cho, Y., Choi, W., 2002. Visible light-induced reactions of humic acids on TiO2. J.

Photochem. Photobiol. A Chem. 148, 129–135.

City of Peterborough Environmental Protection Laboratory, 2014. Water Quality

Analysis Work Order No: 2141025.

Crittenden, J.C., Trussel, R.R., Hand, D.W., Howe, K.J., Tchobanoglous, G., 2005.

MWH’s Water Treatment Principles and Design, 2nd ed. Wiley, Hoboken, NJ, USA.

De la Rubia, A., Rodriguez, M., Leon, V.M., Prats, D., 2008. Removal of natural organic

matter and THM formation potential by ultra- and nanofiltration of surface water.

Water Res 42, 714–722.

Daugherty, E., Mayer, B., Abbaszadegan, M., 2011. Analysis of photocatalysis

implementation in water treatment systems for the removal of DBP precursors, in:

Water Quality Technology Conference and Exposition 2011. American Water Works

Association, pp. 2516 – 28.

Dobrović, S., Juretić, H., Ljubas, D., Vrček, I.V., Avdičević, M.Z., Milić, M., Kopjar, N.,

Nemet, I., 2012. Genotoxicity and effects of nanosilver contamination in drinking

water disinfection. Water Sci. Technol. Water Supply 12, 829-836.


122

Egerton, T.A., 2011. Does photoelectrocatalysis by TiO2 work? J. Chem. Technol.

Biotechnol. 86, 1024–31.

European Commission, 1998. Drinking Water Directive, Council Directive 98/83/EC.

Fujishima, A., Zhang, X., 2006. Titanium dioxide photocatalysis: present situation and

future approaches. Comptes Rendus Chim. 9, 750–760.

Gerrity, D., Mayer, B., Ryu, H., Crittenden, J., Abbaszadegan, M., 2009. A comparison

of pilot-scale photocatalysis and enhanced coagulation for disinfection byproduct

mitigation. Water Res. 43, 1597–1610.

Ghaly, M.Y., Jamil, T.S., El-Seesy, I.E., Souaya, E.R., Nasr, R.A., 2011. Treatment of

highly polluted paper mill wastewater by solar photocatalytic oxidation with

synthesized nano TiO2. Chem. Eng. J. 168, 446–454.

Government of Singapore, 2008. Environmental Public Health Act (Chapter 95)

Environmental Public Health Quality of Piped Drinking Water, Regulations 2008.

Hadnadjev, M., Ranogajec, J., Petrovic, S., Markov, S., Ducman, V., Marinkovic-

Neducin, R., 2010. Design of self-cleaning TiO2 coating on clay roofing tiles. Philos.

Mag. 90, 2989–3002.

Hashimoto, K., Irie, H., Fujishima, A., 2005. TiO2 photocatalysis: a historical overview

and future prospects. Jpn. J. Appl. Phys. 44, 8269–85.

Hatat-Fraile, M., 2014. Personal communications.

Hatat-Fraile, M., Mendret, J., Rivallin, M., Brosillon, S., 2012. Photocatalytic membranes

for the Treatment of refractory organic pollutants. Procedia Eng. 44, 440–442.

Hatat-Fraile, M., Mendret, J., Rivallin, M., Brosillon, S., 2013. Effect of hydrodynamics

during sol–gel synthesis of TiO2 nanoparticles: from morphology to photocatalytic

properties. Chem. Eng. Res. Des. 91, 2389–2400.

Health Canada, 2012. Guidelines for Canadian Drinking Water Quality Summary Table.

Ottawa, Ontario.


123

Herrmann, J.-M., 2010. Photocatalysis fundamentals revisited to avoid several

misconceptions. Appl. Catal. B Environ. 99, 461–468.

Hoehn, R.C., Shorney-Darby, H., Neemann, J., 2010. Chlorine dioxide, in: White, G.C.,

Corporation, B.& V. (Eds.), White’s Handbook of Chlorination and Alternative

Disinfectants. Wiley, Hoboken, NJ, USA, p. 700.

Hrudey, S.E., 2009. Chlorination disinfection by-products, public health risk tradeoffs

and me. Water Res. 43, 2057–92.

Hu, A., Zhang, X., Oakes, K.D., Peng, P., Zhou, Y.N., Servos, M.R., 2011. Hydrothermal

growth of free standing TiO2 nanowire membranes for photocatalytic degradation of

pharmaceuticals. J. Hazard. Mater. 189, 278–85.

Huang, X., Leal, M., Li, Q., 2008. Degradation of natural organic matter by TiO2

photocatalytic oxidation and its effect on fouling of low-pressure membranes. Water

Res. 42, 1142–50.

Hurum, D.C., Agrios, A.G., Gray, K.A., Rajh, T., Thurnauer, M.C., 2003. Explaining the

Enhanced Photocatalytic Activity of Degussa P25 Mixed-Phase TiO2 Using EPR. J.

Phys. Chem. B 107, 4545–49.

Itoh, M., Kunikane, S., Magara, Y., 2001. Evaluation of nanofiltration for disinfection

by-products control in drinking water treatment. Water Sci. Technol. Water Supply 1,

233–243.

Keen, O.S., Linden, K.G., 2013. Re-engineering an artificial sweetener: transforming

sucralose residuals in water via advanced oxidation. Environ. Sci. Technol. 47, 6799–

6805.

Kent, F.C., Montreuil, K.R., Brookman, R.M., Sanderson, R., Dahn, J.R., Gagnon, G.A.,

2011. Photocatalytic oxidation of DBP precursors using UV with suspended and fixed

TiO2. Water Res. 45, 6173–80.


124

Kim, H.-C., Yu, M.-J., 2005. Characterization of natural organic matter in conventional

water treatment processes for selection of treatment processes focused on DBPs

control. Water Res. 39, 4779–89.

Krasner, S.W., Mcguire, M.J., Jacangelo, J.G., Patania, N.L., Reagan, K.M., Aieta, E.M.,

1989. The occurrence of disinfection by-products in US drinking water. J. Am. Water

Work. Assoc. 81, 41–53.

Krasner, S.W., Mitch, W.A., Westerhoff, P., Dotson, A., 2012. Formation and control of

emerging C- and N-DBPs in drinking water. J. AWWA 104, 582–595.

Krasner, S.W., Weinberg, H.S., Richardson, S.D., Pastor, S.J., Chinn, R., Sclimenti, M.J.,

Onstad, G.D., Thruston, A.D., 2006. Occurrence of a new generation of disinfection

byproducts. Environ. Sci. Technol. 40, 7175–85.

Krasner, S.W., Westerhoff, P., Chen, B., Amy, G., Nam, S.-N., Chowdhury, Z.K., Sinha,

S., Rittmann, B.E., 2008. Contribution of Wastewater to DBP Formation.

Kronberg, L., Holmbom, B., Reunanen, M., Tikkanen, L., 1988. Identification and

quantification of the Ames mutagenic compound 3-chloro-4-(dichloromethyl)-5-

hydroxy-2(5H)-furanone and of its geometric isomer (E)-2-chloro-3-(dichloromethyl)-

4-oxobutenoic acid in chlorine-treated humic water and drinking water extract.

Environ. Sci. Technol. 22, 1097–1103.

Li, A., Zhao, X., Liu, H., Qu, J., 2011. Characteristic transformation of humic acid during

photoelectrocatalysis process and its subsequent disinfection byproduct formation

potential. Water Res. 45, 6131–40.

Li, J.W, Yu, Z., Gao, M., Zhang, L., Cai, X., Chao, F., 1996. Effect of ultraviolet

irradiation on the characteristics and trihalomethanes formation potential of humic

acid. Water Res. 30, 347–350.

Li, Q., Mahendra, S., Lyon, D.Y., Brunet, L., Liga, M. V, Li, D., Alvarez, P.J.J., 2008.

Antimicrobial nanomaterials for water disinfection and microbial control: potential

applications and implications. Water Res. 42, 4591–602.


125

Liang, R., 2014. Personal communications.

Liang, R., Hu, A., Persic, J., Zhou, Y.N., 2013a. Palladium nanoparticles loaded on

carbon modified TiO2 nanobelts for enhanced methanol electrooxidation. Nano-Micro

Lett. 5, 202–212.

Liang, R., Hu, A., Li, W., Zhou, Y.N., 2013b. Enhanced degradation of persistent

pharmaceuticals found in wastewater treatment effluents using TiO2 nanobelt

photocatalysts. J Nanopart Res 15, 1990: 1–13.

Liu, S., Lim, M., Amal, R., 2013. Photocatalysis of natural organic matter in water:

characterization and treatment integration, in: Pichat, P. (Ed.), Photocatalysis and

Water Purification: From Fundamentals to Recent Applications. Wiley-VCH Verlag

GmbH & Co. KGaA., pp. 271–294.

Liu, S., Lim, M., Amal, R., 2014. TiO2-coated natural zeolite: rapid humic acid

adsorption and effective photocatalytic regeneration. Chem. Eng. Sci. 105, 46–52.

Liu, S., Lim, M., Fabris, R., Chow, C., Chiang, K., Drikas, M., Amal, R., 2008a.

Removal of humic acid using TiO2 photocatalytic process - fractionation and

molecular weight characterisation studies. Chemosphere 72, 263–71.

Liu, S., Lim, M., Fabris, R., Chow, C., Drikas, M., Amal, R., 2008b. TiO2 photocatalysis

of natural organic matter in surface water: impact on trihalomethane and haloacetic

acid formation potential. Environ. Sci. Technol. 42, 6218–23.

Liu, S., Lim, M., Fabris, R., Chow, C.W.K., Drikas, M., Korshin, G., Amal, R., 2010.

Multi-wavelength spectroscopic and chromatography study on the photocatalytic

oxidation of natural organic matter. Water Res. 44, 2525–32.

Love, S.A., Maurer-Jones, M.A., Thompson, J.W., Lin, Y.-S., Haynes, C.L., 2012.

Assessing nanoparticle toxicity. Annu. Rev. Anal. Chem. 5, 181–205.

Lyon, B.A, Dotson, A.D., Linden, K.G., Weinberg, H.S., 2012. The effect of inorganic

precursors on disinfection byproduct formation during UV-chlorine/chloramine

drinking water treatment. Water Res. 46, 4653–64.


126

Malato, S., Fernández-Ibáñez, P., Maldonado, M.I., Blanco, J., Gernjak, W., 2009.

Decontamination and disinfection of water by solar photocatalysis: recent overview

and trends. Catal. Today 147, 1–59.

Marhaba, T.F., Pipada, N.S., 2000. Coagulation: effectiveness in removing dissolved

organic matter fractions. Environ. Eng. Sci. 17, 107–115.

Mendret, J., Hatat-Fraile, M., Rivallin, M., Brosillon, S., 2013. Hydrophilic composite

membranes for simultaneous separation and photocatalytic degradation of organic

pollutants. Sep. Purif. Technol. 111, 9–19.

Ministry of Environmental Protection of the Government of the People’s Republic of

China, 2006. Standards for Drinking Water Quality, GB 5749-2006.

Mitch, W.A., Krasner, S.W., Westerhoff, P., Dotson, A., 2009. Occurrence and

Formation of Nitrogenous Disinfection By-Products.

Mori, M., Sugita, T., Mase, A., Funatogawa, T., Kikuchi, M., Aizawa, K., Kato, S., Saito,

Y., Ito, T., Itabashi, H., 2013. Photodecomposition of humic acid and natural organic

matter in swamp water using a TiO2-coated ceramic foam filter: potential for the

formation of disinfection byproducts. Chemosphere 90, 1359–65.

Mrowiec-Białoń, J., Jarzębski, A.B., Kholdeeva, O.A., Trukhan, N.N., Zaikovski, V.I.,

Kriventsov, V.V., Olejniczak, Z., 2004. Properties of the sol–gel TiO2–SiO2 oxidation

catalysts prepared using ethyl silicate 40 as a silica precursor. Appl. Catal. A Gen.

273, 47–53.

Narayan, R., 2010. Use of nanomaterials in water purification. Mater. Today 13, 44–46.

Norwood, D.L., Christman, R.F., Hatcher, P.G., 1987. Structural characterization of

aquatic humic material. 2. Phenolic content and its relationship to chlorination

mechanism in an isolated aquatic fulvic acid. Environ. Sci. Technol. 21, 791–8.

Nosaka, Y., Nosaka, A.Y., 2013. Identification and roles of the active species generated

on various photocatalysts, in: Pichat, P. (Ed.), Photocatalysis and Water Purification:

From Fundamentals to Recent Applications. Wiley-VCH Verlag GmbH & Co. KGaA.


127

Ohno, T., Sarukawa, K., Tokieda, K., Matsumura, M., 2001. Morphology of a TiO2

photocatalyst (Degussa, P-25) consisting of anatase and rutile crystalline phases. J.

Catal. 203, 82–86.

Ohtani, B., Prieto-Mahaney, O.O., Li, D., Abe, R., 2010. What is Degussa (Evonik) P25?

Crystalline composition analysis, reconstruction from isolated pure particles and

photocatalytic activity test. J. Photochem. Photobiol. A Chem. 216, 179–182.

Ontario Ministry of the Environment, 2008. Ontario Drinking Water Quality Standards,

O.Reg. 327/08.

Pehlivanoglu-Mentas, E., Sedlak, D.L., 2006. The fate of wastewater-derived NDMA

precursors in the aquatic environment. Water Res 40, 1287–92.

Peterborough Utilities Commission (PUC), 2013. 2013 Water Quality Report.

Philippe, K.K., Hans, C., MacAdam, J., Jefferson, B., Hart, J., Parsons, S., 2010.

Photocatalytic oxidation of natural organic matter surrogates and the impact on

trihalomethane formation potential. Chemosphere 81, 1509–16.

Photo Emission Tech. Inc., 2012. Solar Simulator Model # SS150AAA Operating

Manual. Camarillo, CA, USA.

Plewa, M.J., Muellner, M.G., Richardson, S.D., Fasano, F., Buettner, K.M., Woo, Y.-T.,

McKague, A. B., Wagner, E.D., 2008. Occurrence, synthesis, and mammalian cell

cytotoxicity and genotoxicity of haloacetamides: an emerging class of nitrogenous

drinking water disinfection byproducts. Environ. Sci. Technol. 42, 955–961.

Plewa, M.J., Wagner, E.D., Jazwierska, P., Richardson, S.D., Chen, P.H., McKague,

A.B., 2004a. Halonitromethane drinking water disinfection byproducts: chemical

characterization and mammalian cell cytotoxicity and genotoxicity. Environ. Sci.

Technol. 38, 62–68.

Plewa, M.J., Wagner, E.D., Richardson, S.D., Thruston, A.D., Woo, Y.-T., McKague,

A.B., 2004b. Chemical and biological characterization of newly discovered iodoacid

drinking water disinfection byproducts. Environ. Sci. Technol. 38, 4713–22.


128

Pressman, J.G., Richardson, S.D., Speth, T.F., Miltner, R.J., Narotsky, M.G., Hunter,

E.S., Rice, G.E., Teuschler, L.K., McDonald, A., Parvez, S., Krasner, S.W., Weinberg,

H.S., McKague, a B., Parrett, C.J., Bodin, N., Chinn, R., Lee, C.-F.T., Simmons, J.E.,

2010. Concentration, chlorination, and chemical analysis of drinking water for

disinfection byproduct mixtures health effects research: U.S. EPA’s four lab study.

Environ. Sci. Technol. 44, 7184–92.

Richardson, S., 2003. Disinfection by-products and other emerging contaminants in

drinking water. TrAC Trends Anal. Chem. 22, 666–684.

Richardson, S.D., Plewa, M.J., Wagner, E.D., Schoeny, R., Demarini, D.M., 2007.

Occurrence, genotoxicity, and carcinogenicity of regulated and emerging disinfection

by-products in drinking water: a review and roadmap for research. Mutat. Res. 636,

178–242.

Richardson, S.D., Thruston, A.D., Collette, T.W., Patterson, K.S., Lykins, B.W.J.,

Ireland, J.C., 1996. Identification of TiO2/UV disinfection byproducts in drinking

water. Environ. Sci. Technol. 30, 3327–34.

Selcuk, H., Sarikaya, H.Z., Bekbolet, M., Anderson, M. A, 2006. Bromate formation on

the non-porous TiO2 photoanode in the photoelectrocatalytic system. Chemosphere 62,

715–21.

Shorney-Darby, H., Harms, L.L., 2010. Chlorination of potable water, in: White, G.C.,

Black & Veatch Corporation (Eds.), White’s Handbook of Chlorination and

Alternative Disinfectants. Wiley, Hoboken, NJ, USA, p. 230.

Sohn, J., Amy, G., Yoon, Y., 2007. Process-train profiles of NOM through a drinking

water treatment plant. J. Am. Water Work. Assoc. 99, 145–153.

State of California, 2013. California Code of Regulations, Title 22, Division 4, Chapter

15.5 Disinfectant Residuals, Disinfecation Byproducts, and Disinfection Byproduct

Precursors.


129

Stevenson, F., 1994. Humus Chemistry Genesis, Composition, Reactions, 2nd ed. Wiley,

New York.

Summers, R.S., Hooper, S.M., Shukairy, H.M., Solarik, G., Owen, D., 2013. Assessing

DBP yield : uniform formation conditions. J. Am. Water Work. Assoc. 88, 80–93.

Toor, R., Mohseni, M., 2006. UV-H2O2 based AOP and its integration with biological

activated carbon treatment for DBP reduction in drinking water. Chemosphere 66,

2087–95.

US EPA, 1999. Alternative Disinfectants and Oxidants Guidance Manual.

US EPA, 2009. National Primary and Secondary Drinking Water Regulations EPA 816-

F-09-004.

Uyguner, C.S., Bekbolet, M., Selcuk, H., 2007. A comparative approach to the

application of a physico‐chemical and advanced oxidation combined system to natural

water samples. Sep. Sci. Technol. 42, 1405–19.

Valencia, S., Marín, J.M., Restrepo, G., Frimmel, F.H., 2014. Evaluation of natural

organic matter changes from Lake Hohloh by three-dimensional excitation-emission

matrix fluorescence spectroscopy during TiO2/UV process. Water Res. 51, 124–133.

Villanueva, C.M., Castaño-Vinyals, G., Moreno, V., Carrasco-Turigas, G., Aragonés, N.,

Boldo, E., Ardanaz, E., Toledo, E., Altzibar, J.M., Zaldua, I., Azpiroz, L., Goñi, F.,

Tardón, A., Molina, A.J., Martín, V., López-Rojo, C., Jiménez-Moleón, J.J., Capelo,

R., Gómez-Acebo, I., Peiró, R., Ripoll, M., Gracia-Lavedan, E., Nieuwenhujsen, M.J.,

Rantakokko, P., Goslan, E.H., Pollán, M., Kogevinas, M., 2012. Concentrations and

correlations of disinfection by-products in municipal drinking water from an exposure

assessment perspective. Environ. Res. 114, 1–11.

Weinberg, H.S., 1999. Disinfection byproducts in drinking water: the analytical

challenge. Anal. Chem. News Featur. 801A–808A.

Weinberg, H.S., Krasner, S.W., Richardson, S.D., Thruston, Jr., A.D., US EPA - National

Exposure Research Laboratory, 2002. The Occurrence of Disinfection By-Products (


130

DBPs ) of Health Concern in Drinking Water : Results of a Nationwide DBP

Occurrence Study. Athens, GA.

WHO, 2011. WHO Guidelines for Drinking-Water Quality 4th Ed, WHO Chronicle.

Woodbeck, M., 2007. Evaluating the potential genotoxicity of disinfected source waters.

Master’s thesis, University of Toronto, Ontario.

Yang, X., Shang, C., Westerhoff, P., 2007. Factors affecting formation of

haloacetonitriles, haloketones, chloropicrin and cyanogen halides during

chloramination. Water Res. 41, 1193–1200.

Zhang, X., Du, A.J., Lee, P., Sun, D.D., Leckie, J.O., 2008. TiO2 nanowire membrane for

concurrent filtration and photocatalytic oxidation of humic acid in water. J. Memb.

Sci. 313, 44–51.

Zhang, X., Zhu, Y., Yang, X., Wang, S., Shen, J., Lin, B., Li, C., 2013. Enhanced visible

light photocatalytic activity of interlayer-isolated triplex Ag@SiO2@TiO2 core–shell

nanoparticles. Nanoscale 5, 3359–66.

Zhao, Y.-Y., Boyd, J.M., Woodbeck, M., Andrews, R.C., Qin, F., Hrudey, S.E., Li, X.-F.,

2008. Formation of N-nitrosamines from eleven disinfection treatments of seven

different surface waters. Environ. Sci. Technol. 42, 4857–62.


131

10 APPENDICES

10.1 Experimental Data for Chapters 5 and 6

10.1.1 Calibration Data

10.1.1.1 DOC

The TOC Analyzer prepared calibration standards from a stock standard solution. The

calibration results of the experiments are provided in Table 10-1 and Table 10-2.

Table 10-1: DOC Calibration Data for Synthetic Water Experiments

Parameter P25 &

NB Ag@SiO2@TiO2/P25 Anatase

Anatase-

N

Anatase-

B

slope 4170 4017 4079 4094 4565

intercept 6169 7519 7510 4706 6682

R2 0.9999 0.9994 0.9996 0.9996 0.9992

Table 10-2: DOC Calibration Data for Otonabee Water Experiments

Parameter P25 & NB Ag@SiO2@TiO2/P25

and Anatase

Anatase-N and

Anatase-B

slope 4036 4493 4403

intercept 5744 7759 8063

R2 0.9997 0.9999 0.9998


132

10.1.1.2 THM

Seven THM calibration standards were prepared ranging from 0 to 140 µg/L. Results are

provided in Table 10-3. An example calibration curve is also provided for THM in

Table 10-3: THM Calibration Data

Parameter TCM BDCM CDBM TBM

Slope 108 19 22 53

Intercept -1.30 4.57 2.14 0.05

R2 0.984 0.991 0.995 0.999

Figure 10-1: THM Calibration Curves

Method detection limits (MDL) were also determined based on 9 samples at a

concentration of 2 µ/L. Results are provided in Table 10-4.

Table 10-4: THM MDL Results

Parameter TCM BDCM CDBM TBM

Avg. (µg/L) 5.65 5.58 3.57 3.32

s (µg/L) 0.37 0.06 0.08 0.89

t 2.897 2.897 2.897 2.897

MDL (µg/L) 1.08 0.19 0.25 2.58

y = 107.55x - 1.2957 R² = 0.9844

y = 19.221x + 4.5658 R² = 0.9914

y = 21.687x + 2.1427 R² = 0.995

y = 52.518x + 0.0481 R² = 0.9994

0

10

20

30

40

50

60

70

80

90

0.000 1.000 2.000 3.000 4.000 5.000

Co

nce

ntr

ati

on

(µ

g/

L)

Area Response

TCM BDCM

CDBM TBM


133

10.1.1.3 HAA

Eight HAA calibration standards were prepared ranging from 0 to 60 µg/L. Results are

provided in Table 10-5.

Table 10-5: HAA Calibration Data

Parameter MCAA MBAA DCAA TCAA BCAA DBAA BDCAA CDBAA TBAA

Slope 659 53 40 18 22 23 19 31 49

Intercept 1.59 -1.36 -1.16 -0.10 -0.09 -0.71 0.45 1.48 -0.12

R2 0.986 0.999 0.996 0.999 0.999 0.996 0.999 0.9998 0.998

Method detection limits (MDL) were also determined based on 9 samples at a

concentration of 2 µ/L. Results are provided in Table 10-6. An MDL for MCAA and

MBAA was not determined because their area response on the GC-ECD was low at 2

µg/L and was not detected for some of the samples. The resulting sample size was too

small to obtain a t value.

Table 10-6: HAA MDL Results

Parameter DCAA TCAA BCAA DBAA BDCAA CDBAA TBAA

Avg. (µg/L) 1.14 1.67 1.76 1.33 2.03 2.85 1.23

s (µg/L) 0.39 0.28 0.28 0.29 0.25 0.26 0.38

t 2.897 2.897 2.897 2.897 2.897 2.897 2.897

MDL (µg/L) 1.13 0.80 0.80 0.84 0.73 0.74 1.10


134

10.1.1.4 HAN, HNM, HK

Eight calibration standards were prepared ranging from 0 – 64 µg/L. Results are provided

in Table 10-7. TCP was not detected in any of the standard prepared although it was

listed as an analyte in the stock solution purchased from Sigma Aldrich. It may be due to

the fact that acetone was used as an organic solvent for the working stock solution instead

of the hexane in which the concentrated stock was delivered in.

Table 10-7: HAN Calibration Data

Parameter DCAN BCAN DBAN TCAN DCP CP

Slope 11.35 18.39 34.08 46.68 47.15 24.58

Intercept -3.14 -1.08 1.13 3.00 2.99 0.96

R2 0.976 0.995 0.998 0.990 0.991 0.998

MDL were also determined based on 9 samples at a concentration of 2 µ/L. Results are

provided in Table 10-8.

Table 10-8: HAN MDL Data

Parameter DCAN BCAN DBAN TCAN DCP CP

Avg. (µg/L) -0.56 1.23 3.15 3.56 3.48 2.97

s (µg/L) 0.35 0.26 1.17 0.13 0.20 0.31

t 2.897 2.897 2.897 2.897 2.897 2.897

MDL (µg/L) 1.00 0.74 3.40 0.37 0.59 0.90


135

10.1.2 QA/QC

10.1.2.1 DOC

The running check standards for the DOC analysis are provided in Table 10-9. All but

one check standard are within +/- 10 % of theoretical value of the check standard. The

one that failed was only 0.01 mg/L outside this limit. No corrective action was taken, and

subsequent check standards were within +/- 10 %. There was an increasing baseline in

the TOC analyzer. This increasing baseline was corrected for in the checks and the

environmental samples by determining the slope of the increasing baseline based on the

increasing concentration of the check standards and Milli-Q® samples. The slope was

used to correct the concentration of the samples based on their location in the queue.


136

Table 10-9: QA/QC Data for DOC Analysis

Date

yy/mm/dd Experiment

Number of

Preceding

Samples

Check

Number

in Queue

Check

Standard

Concentration

(mg/L)

Measured

Check

Standard

Concentratio

n (mg/L)

Reagent Blank

Measured

Concentration

(mg/L)

Comments

17/07/2014 P25 and NB 0

0.05

Synthetic 10 1 3 2.91

pass

10 2 3 2.92

pass

10 3 3 2.95

pass

10 4 3 3.02 0.07 pass

23/07/2014 P25AgSiO2 0

0.13


pass

8 2 3 3.14 0.15 pass

24/07/2014 Anatase 0

0.13


pass

8 2 3 3.03 0.14 pass

27/07/2014 Anatase-N 0

0.19

Synthetic 10 1 3 3.10 0.14 pass

10 2 3 3.19 0.15 pass

27/07/2014 Anatase-B 0

0.00

Synthetic 10 1 3 2.86 0.57 pass

10 2 3 2.95 0.23 pass

10 3 3 2.69 0.01

fail - by

.01mg/L

6 4 3 3.24 0.23 pass

05/08/2014 P25 and NB 0

0.24

Otonabee 10 1 3 3.02 0.33 pass

10 2 3 3.26 0.44 pass

6 3 3 3.14 0.30 pass

10 4 3 3.05 0.35 pass

14/08/2014 Ag &

Anatase 0

0.15


10 2 3 2.96 0.23 pass

10 3 3 3.03 0.20 pass

6 4 3 3.04 0.20 pass

10 5 3 2.99 0.24 pass

16/08/2014 A-B & A-N 0

0.13


10 2 3 2.89 0.16 pass

10 3 3 2.90 0.10 pass

6 4 3 2.95 0.16 pass


137

10.1.2.2 THM

Results of the running check standard for THM analysis are provided in Figure 10-2. The

control limits (CL) and warning limits (WL) are +/-3 and +/-2 standard deviations,

respectively. Generally the TCM and other THM check standards were within the quality

control parameters.

Figure 10-2: THM QA/QC Charts

32

34

36

38

40

42

44

46

48

Co

nce

ntr

atio

n (

µg/

L)

TCM Running Check Standards

Avg

CL

WL

Std.dev. +/- 1

P25 NB Ag & Anatase Anatase-N Anatase-B P25 & NB Ag & Anatase Anatase-B & Anatase-N SYNTHETIC WATER EXPERIMENTS OTONABEE RIVER WATER EXPERIMENTS

32

34

36

38

40

42

44

46

48

Co

nce

ntr

atio

n (

µg/

L)

BDCM Running Check Standards

Avg

CL

WL

Std.dev. +/- 1

P25 NB P25 & Anatase Anatase-N Anatase-B P25 & NB Ag & Anatase Anatase-B & Anatase-N SYNTHETIC WATER EXPERIMENTS OTONABEE RIVER WATER EXPERIMENTS

32

34

36

38

40

42

44

46

48

Co

nce

ntr

atio

n (

µg/

L)

CDBM Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


32

34

36

38

40

42

44

46

48

Co

nce

ntr

atio

n (

µg/

L)

TBM Running Check Standards

Avg

CL

WL

Std.dev. +/- 1



138

10.1.2.3 HAA

Results of the running check standard for HAA analysis are provided in Figure 10-3 and

Figure 10-4. The DCAA and TCAA check standards were found to be relatively

consistent, yet elevated from the theoretical check standard concentration. BCAA,

DBAA, BDCAA were also found to be on average elevated from the theoretical check

standard concentration. MCAA and TBAA were found to be lower than the theoretical

check standard and TBAA showed a significant decreasing trend. MBAA ranged from

the upper to lower control limit. The source of these problems was not determined. The

initial 9 running check standard that were prepared to make the QA/QC charts had a very

low standard deviation and subsequent variability in the check standards during

experimental extractions increased. The experiments were all run (TiO2/UV, UFC test

and DBP extraction and GC analysis) within approximately a one month time frame and

during that time, the GC was also used by other students. The stock solutions were

prepared from analytical standards one week before experiments were started and the

same stock solutions were used throughout the experiments.


139

Figure 10-3: HAA QA/QC Charts

4

6

8

10

12

14

16C

on

cen

trat

ion

(µ

g/L)

MCAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

MBAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

DCAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

TCAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1



140

Figure 10-4: HAA QA/QC Charts

4

6

8

10

12

14

16C

on

cen

trat

ion

(µ

g/L)

BCAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

DBAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

BDCAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

CDBAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


0

2

4

6

8

10

12

14

16

Co

nce

ntr

atio

n (

µg/

L)

TBAA Running Check Standards

Avg

CL

WL

Std.dev. +/- 1



141

10.1.2.4 HAN, HNM, HK

Although the analytical standards purchased from Sigma Aldrich listed 4 HAN, one

HNM (chloropicrin CP) and two HKs, only one HK (1,1-diochloro-2-propanone DCP)

was detected in the running check standards. The HAN QA/QC charts are provided in

Figure 10-5 and the THN and HK QA/QC charts are provided in Figure 10-6.

Figure 10-5: HAN QA/QC Charts

6

7

8

9

10

11

12

13

14

Co

nce

ntr

atio

n (

µg/

L)

DCAN Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


6

7

8

9

10

11

12

13

14

Co

nce

ntr

atio

n (

µg/

L)

BCAN Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


6

7

8

9

10

11

12

13

14

Co

nce

ntr

atio

n (

µg/

L)

DBAN Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


0

5

10

15

20

25

Co

nce

ntr

atio

n (

µg/

L)

TCAN Running Check Standards

Avg

CL

WL

Std.dev. +/- 1



142

Figure 10-6: HNM and HK QA/QC Charts

0

2

4

6

8

10

12

14

16

18

Co

nce

ntr

atio

n (

µg/

L)

DCP Running Check Standards

Avg

CL

WL

Std.dev. +/- 1


6

7

8

9

10

11

12

13

14

Co

nce

ntr

atio

n (

µg/

L)

CP Running Check Standards

Avg

CL

WL

Std.dev. +/- 1



143

10.1.3 Supplementary Data

10.1.3.1 Analysis of Variance

An example of the ANOVA output is given in Table 10-10, which shows the Minitab

output for UV254 in Synthetic Water.

Table 10-10: ANOVA Output for UV254 in Synthetic Water

Source DF SS MS F P

TiO2 type 5 29159.9 5831.98 72.44 <0.001

Treatment

Time 4 13425 3356.25 41.69 <0.001

Interaction 20 12209.7 610.48 7.58 <0.001

Error 60 4830.5 80.51

Total 89 59625.1

S 8.973

R2 91.90%

R2(adj) 87.98%

Entries in Table 10-10 are explained below.

Source: indicates the source of variation, either from the factor, the interaction, or error.

The total is the total variation from all sources.

DF: Degrees of freedom from each source. Equal to n – 1 where n is the number of levels

in each factor.

SS: Sum of squares between factors and within factors.

∑( )

MS: Mean squares, calculated by dividing the sum of squares by the degrees of freedom.

F: Calculated by dividing the factor MS by the error MS. It can be used to determine the

significance of a factor.

p: Used to determine if a factor is significant. It is typically compared against a desired %

confidence level. For example, 95% confidence would require a p value < 0.05.


144

10.1.3.2 NOM Characterization

Figure 10-7: DOC and UV254 in TiO2/UV Treated Synthetic Water

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

P25

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

NB

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

AgSiO2/P25

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

Anatase

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

Anatase-N

0

0.02

0.04

0.06

0.08

0.1

0

1

2

3

4

5

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose (mJ/cm2)

Anatase-B


145

Figure 10-8: DOC and UV254 in TiO2/UV Treated Otonabee Water

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

P25

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

NB

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

AgSiO2/P25

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

Anatase

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

Anatase-N

0

0.02

0.04

0.06

0.08

0.1

0.12

0.14

0.16

0

1

2

3

4

5

6

Control 0 28 414 827

UV

25

4 (

1/c

m)

DO

C (

mg/

L)

UV Dose(mJ/cm2)

Anatase-B


146

10.1.3.3 UFC Chlorination Test

A copy of a typical monitoring sheet for the UFC chlorination test is provided below.

Figure 10-9: UFC Chlorination Test Data for AgSiO2/P25 and Anatase TiO2/UV

Treated Otonabee Water

UFC condition: 24 +/- 1 hour, 20.0 +/- 1.0 oC, 8.0 +/- 0.2 pH, 1.0 +/- 0.4 mg/L residual

Trials: 1) P25AgSiO2 TiO2 at 0.1 g/L + Otonabee water + Solar Simulator Aug8th

2) Anatase TiO2 at 0.1 g/L + Otonabee River Water + Solar Simulator Aug 8th

Date: Aug 11th to 12th , 2014 Temperature: 20 oC incubator setting

Check a sample: 20 oC water temp of

Trial Sample ID (140211 +)pH before

spike

Target spike

conc (mg/L)

Volume of

stock to spike

in (mL)**

"Initial" Cl

conc

(mg/L)*

Time of spikepH after Cl

spikepH after 24h

Residual after

24h (mg/L)

***Cl

Demand for

24h (mg/L)

"Initial" Cl

conc for

DBP (mg/L)

*^Vol of

stock to

spike in

(mL)

ORW-Control-1 8.18 5.86 0.700 4.02 10:55:00 AM 8.33 8.21 1.12 2.90 3.90 0.679

P25AgSiO2-0-1 8.33 5.02 0.600 3.62 11:04:00 AM 8.42 8.34 1.00 2.62 3.62 0.600

P25AgSiO2-1-1 8.19 4.98 0.595 3.66 11:11:00 AM 8.32 8.22 0.87 2.79 3.79 0.616

P25AgSiO2-15-1 8.23 5.32 0.635 3.40 11:17:00 AM 8.33 8.22 0.75 2.65 3.65 0.682

P25AgSiO2-30-1 8.15 5.28 0.630 3.50 11:24:00 AM 8.27 8.15 0.66 2.84 3.84 0.691

Anatase-0-1 8.08 5.02 0.600 3.56 11:30:00 AM 8.25 8.16 0.87 2.69 3.69 0.622

Anatase-1-1 8.1 4.98 0.595 3.42 11:36:00 AM 8.25 8.16 0.75 2.67 3.67 0.638

Anatase-15-1 8.1 5.32 0.635 3.52 11:42:00 AM 8.27 8.15 0.76 2.76 3.76 0.678

Anatase-30-1 8.11 5.28 0.630 3.60 11:48:00 AM 8.27 8.15 0.88 2.72 3.72 0.651

* Taken immediately after spiking with stock solution; use this as your starting concentration for calculation of chlorine demand

** Determine by measuring stock concentration, = V2C2/C1

*** Calcuated using a stock concentration that is determined from the "initial"Cl conc

*^ =spike concentration for DBP*Volume of stock to spike in/"initial" Cl conc

Date: Aug 12th to 13th temp: 20 oC incubator setting

DBP formation Check a sample: 20.5 oC water temp of Stock concentration:

Trial Sample IDpH before

spike

Volume of

stock to

spike in

(mL)^

Time of spikepH after Cl

spike

Time of

quenchpH after 24h

Cl residual

after 24h

(mg/L)

diluted sample

measured

(mg/L)

actual

(mg/L)

ORW-control-2 8.14 0.690 1.30 8.33 2.30 8.16 0.82 Date: 11-Aug-14

ORW-control-3 8.19 0.690 1.38 8.36 2.34 8.23 1.09 1 1090

OTW-control-4 8.19 0.690 1.41 8.35 2.38 8.22 1.18 2 1050

P25AgSiO2-0-2 8.13 0.600 1.44 8.31 2.42 8.20 1.17 3 1000

P25AgSiO2-0-3 8.22 0.600 1.46 8.35 2.45 8.27 1.08 avg 1046.66667

P25AgSiO2-0-4 8.28 0.600 1.49 8.38 2.48 8.34 1.06 Date: 12-Aug-14

P25AgSiO2-1-2 8.20 0.615 1.52 8.31 2.51 8.26 0.98 1 980

P25AgSiO2-1-3 8.16 0.615 1.54 8.29 2.55 8.24 0.93 2 980

P25AgSiO2-1-4 8.28 0.615 1.57 8.37 2.58 8.30 1.11 3 1040

P25AgSiO2-15-2 8.11 0.670 2.00 8.25 3.01 8.18 0.85 avg 1000

P25AgSiO2-15-3 8.08 0.670 2.03 8.24 3.04 8.15 0.89

P25AgSiO2-15-4 8.12 0.670 2.05 8.26 3.07 8.18 0.95

P25AgSiO2-30-2 8.12 0.680 2.08 8.30 3.10 8.19 1.00

P25AgSiO2-30-3 8.12 0.680 2.11 8.30 3.13 8.19 1.05

P25AgSiO2-30-4 8.17 0.680 2.14 8.30 3.16 8.22 1.08

Anatase-0-2 8.16 0.620 2.19 8.31 3.19 8.23 1.11

Anatase-0-3 8.09 0.620 2.21 8.28 3.22 8.22 1.09

Anatase-0-4 8.15 0.620 2.23 8.30 3.25 8.22 1.12

Anatase-1-2 8.05 0.630 2.26 8.23 3.28 8.18 0.95

Anatase-1-3 8.11 0.630 2.28 8.26 3.31 8.15 1.12

Anatase-1-4 8.06 0.630 2.31 8.22 3.33 8.16 0.95

Anatase-15-2 8.02 0.670 2.33 8.20 3.36 8.12 0.98

Anatase-15-3 8.03 0.670 2.35 8.20 3.39 8.14 0.95

Anatase-15-4 8.06 0.670 2.38 8.24 3.42 8.17 1.06

Anatase-30-2 8.10 0.650 2.40 8.26 3.45 8.16 1.09

Anatase-30-3 8.06 0.650 2.42 8.24 3.48 8.14 0.99

Anatase-30-4 8.10 0.650 2.45 8.23 3.50 8.17 1.00

Chlorine Demand

1

2

1

2


147

10.1.3.4 THM fp

Figure 10-10: THM fp in Synthetic Water Following Treatment with TiO2/UV and

Chlorination

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

P25

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

NB

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

P25/AgSiO2

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-N

0

10

20

30

40

50

60

70

80

0

10

20

30

40

50

60

70

80

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-B


148

Figure 10-11: THM fp in Otonabee Water Following Treatment with TiO2/UV and

Chlorination

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

NB

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose(mJ/cm2)

P25

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

P25/AgSiO2

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-B

0

10

20

30

40

50

60

70

80

0

20

40

60

80

100

120

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-N


149

10.1.3.5 HAA fp

Figure 10-12: HAA fp in Synthetic Water Following TiO2/UV and Chlorination

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

P25

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

NB

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

AgSiO2/P25

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase -N

0

5

10

15

20

25

30

35

40

0

5

10

15

20

25

30

35

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-B


150


Chlorination

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

P25

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Nanobelt

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

AgSiO2/P25

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-B

0

5

10

15

20

25

30

35

40

0

10

20

30

40

50

60

70

0 28 414 827

Avg

. C

on

c./

SUV

A (

µg*

mg*

m/L

2)

Avg

. C

on

c. (

µg/

L)

UV Dose (mJ/cm2)

Anatase-N


151

10.2 Experimental Data for Preliminary Experiments

10.2.1 Optimal TiO2 Dark Adsorption Time

Figure 10-14: UV254 in Synthetic Water Following Treatment with P25 TiO2/UV

under Various Dark Adsorption and Irradiation Times

10.2.2 UV Fluence Rate

Figure 10-15: UV-Vis Absorbance of a 0.1 g/L TiO2 suspension in Milli-Q®

0

0.02

0.04

0.06

0.08

0 1 30

UV

25

4 (

1/

cm)


0 ads

1 ads

2 ads

5 ads

10 ads

0

0.5

1

1.5

2

2.5

3

3.5

4

4.5

200 300 400 500 600 700 800 900 1000 1100

Ab

sorb

ance

(cm

-1)

Wavelength (nm)


152

Figure 10-16: P25 TiO2/UV Methylene Blue Degradation with and without a Vortex

10.3 Sample Calculations

10.3.1 Determining DBP Concentration

From TCM calibration curve provided above,

y = 107.55x - 1.2957

Where: y = concentration (µg/L)

x = area response

The first replicate of synthetic water treated with P25 TiO2/UV for 28 mJ/cm2 had an area

response ratio of 0.475.

Where: area response ratio=

Therefore, the concentration of TCM was:

49.8 µg/L

0.7

0.9

1.1

1.3

1.5

1.7

1.9

2.1

-10 0 10 20 30

UV

66

5 a

bso

rba

nce

(1

/cm

)

Time (min)

No vortex

with vortex

control

Control


153

10.3.2 Determining UV Dose in Published TiO2/UV Studies

Liu et al., 2008a; Liu et al., 2008b; Liu et al., 2010

Given:

Photon flux = 20.3 µE/Ls (where E = Einstein = one mole of photons) Photon flux = 0.0000203 moles/Ls Wavelength = 0.000000365 m

Where:

U = energy of one mole of photons of specific wavelength (J/mol)

λ = wavelenth (m)

C = 299790000 (speed of light in vacuum (m/s)) H = 6.6261E-34 (plancks constant (Js))

NA = 6.02214E+23 (Avogadro's number ( mol-1))

Calculations:

U= 327743 U = Energy (J/mol) of one mole of 365 nm light UV flux 6.65 J/Ls = W/L UV dose 399 J/L (1 min) UV dose 11976 J/L (30 min)

Gerrity et al., 2009

UV dose = 5 kWh/m3 UV dose = 18000000 J/m3 UV dose = 18000 J/L

Mori et al., 2013

UV fluence rate= 2 mW/cm2 Irradiation time = 43200 sec UV dose = 86400 mJ/cm2 Surface area = 100 cm2 surface area

UV dose = 8640000 mJ Volume = 0.5 L (of water treated) UV dose = 17.28 KJ/L

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