‫דוח מסכם תלת שנתי‬
‫‪ 01/12/2010‬עד ‪31/11/2013‬‬
‫מענק מספר בקרן‪0455004054 :‬‬
‫‪1‬‬
‫חלק א '‪ :‬פרטים כלליים על המחקר‬
‫שם המחקר‬
‫סילוק שאריות תרופות ממי שתייה וקולחים‬
‫בעזרת תהליכי חמצון מתקדם‬
‫שמות החוקרים‬
‫דר' הדס ממן‪ ,‬דר' דרור אבישר‬
‫שם המוסד הראשי‬
‫אוניברסיטת תל אביב‬
‫תקופת ההתקשרות‬
‫‪ 3‬שנים‬
‫‪2‬‬
‫חלק ב '‪ :‬רקע מדעי כללי‬
‫מספר עבודות שנעשו בארה"ב ובאירופה הראו כי ברבים ממקורות המים העיליים הגדולים (נהרות)‬
‫ואף במאגרי מי תהום נמצאו ריכוזים מדאיגים של שאריות תרופות והורמונים כתוצאה מהזרמת‬
‫שפכים וקולחים מטוהרים המכילים מזהמים אלו באגני ההיקוות‪ ,‬בהחדרת קולחים למי‪-‬תהום‬
‫ובהשקיה‪ .‬מזהמים אלו‪ ,‬אשר מקורם במטבוליזם לא שלם בגוף האדם‪ ,‬מגיעים דרך שפכים ביתיים‪,‬‬
‫חקלאיים ותעשייתיים למכון טיהור שפכים‪ .‬מחקרים שנעשו בארץ ובעולם הוכיחו כי שיטות הטיפול‬
‫בשפכים הנפוצות כיום‪ ,‬כגון בוצה משופעלת‪ ,‬לא מתוכננות להרחקה מלאה של מזהמים אורגניים‬
‫אלו‪.‬‬
‫תהליכי חמצון מתקדם מבוססי אוזון ידועים כבעלי פוטנציאל גבוה מאוד בפירוק ובסילוק מזהמים‬
‫רבים‪ .‬התהליכים כוללים אוזון בלבד וכן שילוב של אוזון ומי חמצן (‪ )H2O2/O3‬או קרינה אולטרה‪-‬‬
‫סגולה (‪ .)UV/O3‬מערכת ‪ (AOP - Advanced Oxidation Processes) AOP‬המשלבת מספר‬
‫מחמצנים וקרינת אור נחשבת כבעלת פוטנציאל רב לפירוק מזהמים ואף לקטילת מיקרואורגניזמים‬
‫במים )‪.(Mamane et al., 2007‬נמצא כי תהליך הכולל אוזון בלבד הינו יעיל בפירוק מיקרו‪-‬מזהמים‬
‫רבים‪ ,‬ואילו מזהמים העמידים לאוזון ניתנים לפירוק על ידי הוספת מי חמצן לתהליך (‪.)H2O2/O3‬‬
‫מורכבות התהליך בטיפול בקולחים גוברת (לעומת טיפול במי שתייה)‪ ,‬כאשר בנוסף לפרוק הישיר‬
‫והעקיף של המזהם‪ ,‬חומרים אורגנים אחרים בקולחים מותקפים גם הם על ידי רדיקלי ההידרוקסיל‬
‫ויכולים ליצור שרשרת תגובות מעגלית אשר תאיץ או תעכב את יצירת הרדיקלים‪ .‬רדיקל ההידרוקסיל‬
‫מיוצר בתהליך החמצון המתקדם‪ ,‬מאופיין ביכולת חמצון מהגבוהות ביותר לפירוק של מיקרו‪-‬מזהמים‬
‫אורגנים‪ .‬התגובה הלא סלקטיבית של הרדיקלים מאפשרת את התאמת התהליך לטיפול במיקרו‪-‬‬
‫מזהמים רבים ומגוונים‪ .‬יצירת הרדיקלים מוצגת בעבודתינו בעזרת טיפולים המשלבים מי חמצן‪,‬‬
‫קרינת ‪ UV‬ואוזון‪ ,‬וכן תהליך פוטוקטליזה העושה שימוש באור השמש ובקטליסט המבוסס על‬
‫תחמוצת ביסמוט‪.‬‬
‫דוח זה מסכם את עיקר תוצאותיו של מחקר בן שלוש שנים שנערך באוניברסיטת תל אביב ומדגים‬
‫את יעילות שיטות הטיפול שנבדקו עבור תרופות מסוגים שונים ועבור סוגי מים שונים‪.‬‬
‫בפרק הראשון ניתן לראות את הירידה בריכוז האנטיביוטיקות ‪ Ciprofloxacin‬ו‪Trimethoprim-‬‬
‫והתרופה הכימותרפית ‪ Cyclophosphamide‬במים‪ ,‬בעקבות טיפול באוזון בלבד (מנת אוזון ‪ 1‬מג"ל)‬
‫ובאוזון משולב עם מי חמצן (ריכוז מי חמצן ‪ 1‬מג"ל)‪ .‬יישום נוסף אשר נבדק לתהליך ‪ AOP‬הינו‬
‫הפחתת רמת הפחמן האורגני (‪ )TOC‬בקולחים שניוניים‪ ,‬כטיפול קדם לפני ממברנות אוסמוזה‬
‫הפוכה‪ .‬הניסויים שנערכו הראו כי תהליכי החמצון המתקדם יכולים להביא למינרליזציה של החומר‬
‫האורגני בקולחים ולירידה משמעותית בריכוז ‪( TOC‬עד ‪ 54%‬ירידה)‪ ,‬כאשר התהליך המשלב אוזון‬
‫ומי חמצן נמצא כיעיל ביותר מבחינה אנרגטית מבין חלופות רבות שנבחנו‪.‬‬
‫‪3‬‬
‫קרינת ‪ UV‬משמשת כיום בעיקר לחיטוי מי שתייה וקולחים‪ ,‬אולם היא נמצאה כבעלת פוטנציאל‬
‫לפרוק מיקרומזהמים כגון שאריות תרופות (בתהליך פוטוליזה)‪ .‬פרמטר חשוב היכול להשפיע על קצב‬
‫הפוטוליזה של חומרים רבים הוא ‪ .pH‬המחקר המוצג בפרק השני מצא כי למזהמים רבים קיים ערך‬
‫‪ pH‬בו קצב הפוטוליזה שלהם מקסימאלי‪ .‬לכן‪ ,‬על ידי אופטימיזציה של ‪ pH‬המים ניתן להגיע לקצב‬
‫פוטוליזה מקסימאלי של המזהם‪ .‬הפוטנציאל של פוטוליזה ב‪ pH-‬אופטימאלי‪ ,‬כשיטת טיפול במים‪,‬‬
‫הודגמה על ידי הסרת תערובת של התרופות סולפאמטוקסזול וטריכלוזן ממי תהום‪ ,‬תוך שינוי ‪pH‬‬
‫המים במהלך הניסוי‪ .‬התהליך המוצע שיפר בצורה דרמטית את יעילות הסרת התערובת (ואת צריכת‬
‫האנרגיה) לעומת טיפול ‪ UV‬ב ‪ pH‬קבוע‪.‬‬
‫פוטוקטליזה הינה שיטת טיפול נוספת מבוססת קרינה שנחקרה ומוצגת בפרק השלישי‪ ,‬כאשר נעשה‬
‫שימוש בקטליסט חדש מבוסס ביסמוט (‪ )BiOCl0.875Br0.125‬ובקרינה סולרית‪ ,‬לפרוק התרופות‪:‬‬
‫‪ .carbamazepine-CBZ, ibuprofen-IBF, bezafibrate-BZF and propranolol PPL‬הקטליסט‬
‫נמצא יעיל בפירוק כל התרופות שנבדקו‪ ,‬כאשר‪ ,‬ב ‪ ,pH = 7‬קצב הפרוק היה על פי סדר יורד‪PPL :‬‬
‫‪ .> BZF > IBF > CBZ‬מספר פרמטרים השפיעו על יעילות הטיפול‪ ,‬כולל ‪ pH‬המים‪ ,‬ריכוז הקטליסט‬
‫והמזהם ותוספת מי חמצן לתהליך‪ .‬לדוגמא‪ ,‬הורדת ה‪ pH -‬מ ‪ 9‬ל‪ 0-‬גרמה לעלייה ברמת הספיחה‬
‫של ‪ CBZ‬לקטליסט‪ ,‬וכתוצאה מכך לעלייה חדה בקצב פרוק התרופה‪ .‬מספר מנגנונים הוצעו‬
‫כאחראים לספיחת המזהמים השונים על הקטליסט‪ ,‬כולל משיכה הידרופובית ל ‪ CBZ‬הניטרלי וחילוף‬
‫יונים ל ‪ BZF‬ו‪ IBF-‬הטעונים שלילית‪.‬‬
‫השפעת חלקיקים על יעילות תהליכי אוזונציה והשפעת אוזון על חלקיקים בקולחים נבדקה אף היא‬
‫במחקר חדשני באוניברסיטת תל אביב ומוצגת בפרק הרביעי‪ .‬חלקיקים בשפכים מיוצגים בדרך כלל‬
‫על‪-‬ידי פרמטר כללי במים כחלקיקים מרחפים (‪ .)TSS‬בעוד מחקרים אחרים מצאו כי ‪ TSS‬משפיע‬
‫מעט על האוזונציה בשפכים‪ ,‬אך פרקציות שונות של ‪ TSS‬עם מאפיינים ספציפיים (למשל גודל‬
‫החלקיקים‪ ,‬ומטען) עשויים להשפיע עדיין על האוזונציה‪ .‬יתר על כן‪ ,‬התגובה של האוזון עם‬
‫החלקיקים עלול להשפיע על מאפייני החלקיקים‪ .‬השפעתם של חלקיקים על האוזונציה באמצעות‬
‫פרמטרים כלליים כמו ‪ TSS‬נבחנה‪ ,‬ואילו אף מחקר לא בדק ההשפעה של פרקציות שונות של ‪TSS‬‬
‫(כלומר גודל חלקיקים ופיזורם)‪ .‬מטרת מחקר זה היה לבדוק את ההשפעה של גודל החלקיקים שונים‬
‫(לאחר פרקצונציה ומדידה ע"י אנליזת חלקיקים) על אוזונציה בשפכים‪ ,‬דרך הפירוק של‬
‫מיקרומזהמים נבחרים ושינויים בפרמטרים הכללים של שפכים (למשל בליעת ‪ .)UV‬בנוסף‪ ,‬נבדקה‬
‫ההשפעה של האוזון על חלקיקים בשפכים ועל פוטנציאל הזטה של החלקיקים‪.‬‬
‫‪0‬‬
CHAPTER I
Ozone Degradation of Cyclophosphamide – Effect of Alkalinity and Key Effluent
Organic Matter Constituents
Introduction
Ozone has potential in degrading and removing trace organic micropollutants in water (e.g.
pharmaceuticals) (e.g. Dodd et al., 2006; Wert et al., 2009). These organic contaminants are
degraded during ozone application through two main pathways: direct oxidation by molecular
ozone and the indirect radical oxidation, where •OH radical is the main contributor. The
degradation kinetics of the contaminant and the contribution of each of the oxidation
processes are influenced by the target contaminant characteristics (i.e. its reaction rate with O3
and •OH) and by the composition of the treated water. Staehelin and Hoigne (1985) showed
that the aqueous ozone is decomposed (and •OH formed) generally through a radical chain
reaction initiated by reaction of O3 with OH-. This chain reaction can be promoted by solutes
which transfer •OH radical into superoxide radical ion (O2•-) (e.g. humic acid), or inhibited by
solutes which do not promote O2•- (e.g. carbonate). Alternatively, Pi et al. (2005) proposed
that the radical chain reaction can be accelerated by different aromatic compounds, where insitu hydrogen peroxide is a main intermediate and chain carrier; while Pocostales et al. (2010)
suggested that •OH generation may additionally occur through ozone adducts to aromatic
compounds, the elimination of singlet oxygen and the formation of phenol. In natural waters,
various researchers have identified natural organic matter (NOM) and alkalinity as the main
promoters and inhibitors of the radical chain reaction in the ozonation process (e.g. Acero and
Von Gunten, 2001). Huber et al. (2003) found that the ozonation efficiency of various
pharmaceuticals, which react slowly with O3 and quickly with •OH, increased with increasing
NOM concentration and decreased with increasing alkalinity in different natural waters.
4
In wastewater effluent, ozonation kinetics differs from drinking waters due to the relatively
high concentration and complexity of the effluent organic matter (EfOM). Effluent organic
matter (EfOM) is mainly composed of NOM, soluble microbial products (SMPs) and nonbiodegradable organic materials. SMPs (e.g. polysaccharides and proteins) mostly originate
from the biological processes within the wastewater treatment plant (WWTP); while NOM,
which includes a vast variety of organic molecules and functional groups, typically originates
from the source water (i.e. drinking water) (Shon et al., 2006). The relative contribution of the
different EfOM constitutes may vary with place, season and treatment.
Numerous studies have demonstrated the differences in ozonation kinetics of pharmaceuticals
between natural water and wastewater effluents (e.g. Benitez et al., 2009). Zimmerman et al.
(2011) investigated the degradation of different micropollutants in a full-scale reactor,
treating secondary wastewater effluent, for ozone doses in the range of 0.21 to 1.24 g O 3: g
DOC. They found that compounds reacting fast with ozone (kP;O3 > 104 1/M s) were
eliminated at almost all O3 doses, while substances with lower ozone reactivity (kP;O3 < 104
1/M s) were only fully eliminated at the high ozone doses. Buffle et al. (2006) concluded that
the decomposition of ozone in wastewater and the formation of •OH radicals are initiated and
controlled by direct reaction of O3 with reactive moieties in the EfOM, rather than through the
radical chain reaction. Their study showed the presence of remarkably high •OH
concentration during the first seconds of the O3 process (higher than in most advanced
oxidation processes (AOPs) in natural waters).
In wastewater effluents, it is difficult to model ozone degradation of micro-pollutants as a
result of the indistinct effects of different EfOM compounds. The objective of the present
study is to determine the degradation of the anticancer drug cyclophosphamide (CPD) by
ozonation at, either low or high alkalinity values, in the presence of different EfOM’s
constituents. Cyclophosphamide (CPD) degradation is of particular interest due to its known
mutagenic properties and its resistance to most conventional and advanced wastewater
treatments (Kim et al., 2008). Three model compounds were used to simulate main
5
constituents in the EfOM, separately and in a mixture: (I) Alginate, an acidic polysaccharide,
extensively studied in the field of biofouling in water and wastewater systems, and a potential
contributor to SMPs in wastewater effluent (Lee et al., 2006), (II) Peptone from casein, used
as a source of amino acids and peptides and, (III) Suwannee River NOM, used as NOM
contributor.
Material and methods
Standards and Reagents
Cyclophosphamide (CPD) standard (>99% purity) was obtained from Sigma-Aldrich, LC-MS
grade
acetonitrile,
methanol
and
water
from Bio-Lab
Ltd.
(Jerusalem,
Israel).
Cyclophosphamide (CPD) stock solution was prepared by dissolving the compound in
deionized (DI) water (Direct-Q3 UV system, Millipore-France; resistivity > 18 mΩ cm) at a
concentration of 100 mg/L. The probe compound p-chlorobenzoic acid (pCBA) was used to
determine •OH radical reaction rate constants with the EfOM model compounds (Cat. No.
13,558-5, Sigma Aldrich, Germany). Alginic acid (alginate), extracted from brown algae, was
purchased from Sigma-Aldrich (CAS: 9005-32-7). Peptone from casein was obtained from
Fluka (CAS: 91079-40-2). Suwannee River NOM was obtained from the International Humic
Substances Society-IHSS (St. Paul, MN). Alkalinity of the water was modified using sodium
bicarbonate (Sigma-Aldrich). Stock solutions of alginic acid and NOM (500 mg/L) were
prepared in DI water, adjusted to pH 10 by the addition of NaOH. Peptone (500 mg/L) and
sodium bicarbonate (30 gr/L) stock solutions were prepared in DI water. All stock solutions
were filtered through a 0.45 μm cellulose acetate filter.
Cyclophosphamide ozonation at different EfOM model compounds solutions
Ozone experiments were performed in a thermostated (21 ºC) 1 L glass cylindrical batch
reactor (Ace-glass, Vineland, NJ). Ozone was generated from pure oxygen (>99.9%) using an
ozone generator (2-5 gr/h, OZO-1VTT, Ozomax, Canada). The ozone stream was introduced
into the aqueous solution using a diffuser located at the bottom of the reactor, while another
tube carried off gases from the headspace of the reactor to the ozone destructor. The tested
solution was constantly mixed via magnetic stirring. Ozone in the aqueous solution was
7
continually measured using the WADIS310 dissolved O3 sensor (Walchem Corporation,
Holliston, MA).
Cyclophosphamide (CPD) ozonation experiments were designed to examine the influence
of the following parameters: (a) different EfOM model compounds at dissolved organic
carbon-DOC concentration range of 0-8 mg/L as C and, (b) the influence of alkalinity at
concentrations of 25 and 200 mg/L as CaCO3 (referred in the text as low and high alkalinity).
In all cases, experiments were initiated by diffusing the ozone stream into a solution
composed of phosphate buffer (2.5 mM) at pH 7.6 (within the pH range of wastewater
effluent) and different concentrations of sodium bicarbonate, until approximately 5.3 mg/L
dissolved O3 concentration was achieved. Ozone diffusion was then stopped, followed by the
injection of CPD (initial concentration of  1 mg/L) and different concentrations of the tested
EfOM model compounds. Sampling began after a mixing time of ~10 s, when dissolved O3
concentration reached exactly 5 mg/L. Samples (2 mL) were taken periodically, quenched
immediately with excess of sodium thiosulfate to decompose residual O3 and analyzed
chromatographically for CPD concentration. Although typical mg O3:mg DOC ratio do not
exceed 1, in the present experiments ozone dose was held constant in order to measure the
isolated effect of different DOC and alkalinity concentrations. The pH of the solution
remained steady throughout the experiment duration.
Determination of the EfOM model compounds rate constants with •OH
The second-order rate constants of •OH with the EfOM model compounds (kEfOM,•OH,
L/(molC) s) were determined in a medium pressure (MP) UV/H2O2 collimated beam
apparatus (described in details elsewhere; Lester et al., 2010), using a modified method
adapted from Rosenfeldt and Linden (2004). UV based AOP was used in the present study as
a simple method for obtaining the rate constants with •OH. Irradiation experiments were
performed on buffered water samples (100 mL PBS 2.5 mM at pH 7.6) with added pCBA (1
µM), H2O2 (1.47 mM) and different concentrations of the tested EfOM model compound (0-8
mg/L as C). The UV incident irradiance was obtained using a calibrated spectroradiometer
5
(USB4000, Ocean Optics, Florida, USA) and the UV absorbance of the treated solution was
measured via UV-Vis spectrophotometer (Varian, Cary 100BIO, Victoria, Australia).
Analytical methods
Cyclophosphamide (CPD) was detected and quantified by HPLC-UV Agilent, model 1100
(ACE-RP C18 column 2.5mm×250mm) and an MS detector (Finnigan LCQ). The mass
spectrometer was used in positive electro-spray ionization (ESI) mode and the probe
temperature was set to 220°C. The flow from the HPLC was passed through a split connector
with 60 μL/min of effluent introduced into the MS interface. Ions in the range 200-300 m/z
were registered in the conventional scanning mode. The mobile phase was ammonium
formate 0.05M (A) and methanol (B), at pH 5. The mobile phase eluent gradient started with
50% of eluent A, followed by a 2.5-min linear gradient to 30% of eluent A, 3-min isocratic
elution and a 2 min linear gradient back to 50% of eluent A, maintained for 4 min to
equilibration time. Conditions for p-chlorobenzoic acid (pCBA) quantification are detailed in
Lester et al. (2010). Dissolved organic carbon (DOC) of the EfOM solutions was measured
using a TOC analyzer (Apollo 9000, TekmarDohrmann).
Results and discussion
O3 decomposition
The difference in dissolved ozone decomposition in the presence of the EfOM model
compounds, at low and high alkalinity values, is demonstrated in Figure 1, where Ln
[O3]t/[O3]0 is plotted with time for alginic acid, NOM and peptone at DOC concentration of 1
mg/L (as C), and alkalinity concentrations of 25 and 200 mg/L. In the presence of NOM and
peptone, an initial rapid O3 decomposition phase was observed (0-60 s), followed by a second
slower decomposition phase. Nothe et al. (2009) observed three-phase kinetics for ozone
decay in wastewater effluent, where the first phase occurred within seconds. Therefore, we
suspect that the two phases measured in our study represent the latter two phases of a threephase O3 decay kinetics. In contrast, only one O3 decomposition phase was observed in the
presence of alginic acid. In all cases, ozone decay exhibited an apparent first-order kinetics,
9
and generally, in the order of: peptone > NOM > alginic acid (Figure 1). O 3 first-order decay
rate constants in the presence of the EfOM model compounds, for the rapid and slower
phases, are presented in Table 1 (k, 1/s).
Figure 1: Dissolved O3 decomposition (O3 dose 5 mg/L) in the presence of alginic acid,
peptone and NOM (1 mg/L as C), at alkalinity concentrations of 25 and 200 mg/L and pH 7.6
(in the presence of CPD). The inset is a magnification of 0-60 s.
The difference in ozone decomposition rate between the EfOM model compounds depends on
their structure and, more specifically, on the existence of different moieties reacting directly
with ozone and/or promoting the radical chain reaction, as follows:

NOM structure is highly complex, consisting of a wide variety of organic molecules
(e.g. humic acids) and functional groups (e.g. phenolic moieties). Numerous studies
have already confirmed the reactivity of NOM and its influence on ozone
decomposition (e.g. Westerhoff et al., 1999). For example, Mvula and Von sonntag
(2003) showed that phenol may enhance O3 decomposition both by direct reaction
and by promoting the radical chain reaction.
15

Peptone contains a mixture of small proteins, peptides and amino acids. Buffle and
Von Gunten (2006) have demonstrated the high reactivity of deprotonated amino
compounds in general, and of different amino acids in particular towered ozone.
Moreover, Hoigne and Bader (1983b) found extremely high rate constants for the
reaction of ozone with different amino acids which contain thio groups (e.g.
cysteine). The peptide linkage connecting the α-amino group of the amino acids in
polypeptides and proteins has a very low reactivity toward ozone (Pryor et al., 1984).
Therefore, the high decomposition rate of ozone in the present study is probably due
to its direct reaction with free amino acids and reactive side-chain groups in the
peptides and proteins.

Alginic acid is a linear polysaccharide composed of mannuronic and guluronic acid
subunits. In general, saturated compounds such as polysaccharides react slowly with
molecular ozone (e.g. reaction rate constant of ozone with glucose = 0.45 1/M s;
Hoigne and Bader, 1983a), thus the radical chain reaction will most likely
predominate O3 decay (hence the one-phase kinetic). Akhlaq et al. (1990) concluded
that over 70% of the •OH radicals reacting with alginic acid during ozonation lead to
the formation of O2•-, further reacting with O3 to enhance its decay.
The influence of the EfOM compound’s concentration on ozone decomposition is
demonstrated in figure 2, for NOM at DOC values of 1, 3 and 8 mg/L, and alkalinity
concentrations of 25 and 200 mg/L, and in Table 1. Generally, as expected, ozone decay rate
increases with increase in DOC concentrations for all EfOM model compounds at both low
and high alkalinity values. However, in the presence of alginic acid, the increase in ozone
decay rate initiates only at DOC > 1 mg/L, emphasizing the alginic acid’s low reactivity
toward molecular ozone.
11
Figure 2: Dissolved O3 decomposition (O3 dose 5 mg/L) in the presence of different NOM
concentrations (as DOC), at different alkalinity concentrations of 25 and 200 mg/L and pH
7.6 (in the presence of CPD). Numbers in the legends (1, 3 and 8) refer to DOC concentration
in mg/L.
Carbonate alkalinity is known to inhibit ozone decomposition in water by reacting with the
generated •OH radicals, and forming oxidation products that do not promote the radical chain
reaction. The stabilizing effect of alkalinity on ozone was demonstrated by comparing ozone
decay rate at low and high alkalinity for a specific EfOM model compound concentration
(Table 1). For alginic acid, the stabilizing effect of alkalinity on O3 decay rate was clearly
demonstrated throughout the entire DOC range, where increasing alkalinity at a specific DOC
value always decreased ozone decay rate by more than 10%. In the presence of NOM,
alkalinity inhibition was less pronounced during the first phase of O3 decay, for DOC ≥ 3
mg/L. For peptone, the effect of alkalinity was minor for DOC ≥ 3 mg/L, during both first and
second O3 decay phases.
The different effects of alkalinity on O3 decay emphasize the diverse mechanisms
responsible for O3 decay in the presence of the different EfOM model compounds. Using
alginic acid, ozone decay occurred most likely through the radical chain reaction path. This
12
reaction path is relatively slow and highly sensitive to the presence of •OH scavengers (i.e.
alkalinity). In solutions containing the highly reactive NOM and peptone, decomposition of
ozone followed both the direct and radical chain reaction paths. Direct reaction of ozone with
different moieties in the organic matter may be more rapid and independent of alkalinity
concentration (Buffle et al., 2006). The relative contribution of each path differs in the first
and second phase of O3 decay and, depends on the EfOM model compound’s concentration
and its reactivity. For example, direct reaction of O3 with the organic matter is probably the
dominant mechanism for O3 decay in the presence of NOM, during the first O3 decay phase,
at DOC ≥ 3 mg/L and, in the presence of peptone, at both the first and second O 3 decay
phases, at DOC ≥ 3 mg/L.
In an attempt to better simulate “real” wastewater effluent, O3 decomposition experiments
were conducted using a mixture of the model compounds. Since the fraction distribution of
EfOM may vary substantially depending on the wastewater origin and the type of treatment
(Imai et al., 2002; Jarusutthirak et al., 2002), an exemplary ratio of 2:1:2 (for alginic
acid:NOM:peptone) was chosen. The experimental O3 decomposition rate (termed mixture)
was compared to the calculated value (termed sum) (i.e. weighted sum of the individual rate
constants) (Table 1). The experimental rate constants were higher than the calculated values
in almost all cases. Rosario-Ortiz et al. (2008) found that the reaction rate constants of •OH
with different non-isolated EfOM were 3-5 times higher than the rate constants of •OH with
fractionated EfOM. Thus, a system containing various organic solutes is not a simple mixture
with respect to its reactivity toward ozone and •OH. Possibly, interactions between the
examined EfOM model compounds contributed to this phenomenon.
13
Table 1: First-order ozone decay rate constant (in the presence of CPD) k*(ksec**), s -1
TOC mg L-1
0
0.5
Alkalinity
25 mg L-1
Alkalinity
200 mg L-1
1
3
5
8
Alginic acid
0.0026
0.0027
0.0026
0.0039
0.0065
0.0081
NOM
0.0028
Peptone
0.0025
Mixture
0.0025
Sum***
0.0026
Alginic acid
0.0016
0.0036
(0.0029)
0.0076
(0.0047)
0.0059
(0.0043)
0.0046
(0.0033)
0.0015
0.0050
(0.0034)
0.0125
(0.0068)
0.0089
(0.0053)
0.0070
(0.0044)
0.0014
0.0135
(0.0115)
0.0288
(0.0173)
0.0198
(0.0172)
0.0158
(0.0108)
0.0017
0.0210
(0.0161)
0.0357
(0.0190)
0.0280
0.0218)
0.0211
(0.0139)
0.0024
0.0380
(0.0205)
0.0407
(0.0204)
0.0373
(0.236)
0.0271
(0.0155)
0.0024
NOM
0.0014
Peptone
0.0014
Mixture
0.0014
Sum***
0.0015
0.0025
(0.0016)
0.0049
(0.0018)
0.0037
(0.0018)
0.0031
(0.0016)
0.0037
(0.0017)
0.0103
(0.0049)
0.0058
(0.0021)
0.0054
(0.0029)
0.0125
(0.0062)
0.0282
(0.0161)
0.0155
(0.0111)
0.0145
(0.0084)
0.0193
(0.0109)
0.0351
(0.0173)
0.0254
(0.0180)
0.0187
(0.0101)
0.0362
(0.0172)
0.0391
(0.0189)
0.0344
(0.0194)
0.0234
(0.0120)
*
ksec– First-order rate constant for the first (rapid) phase of O3 decay
ksec– First-order rate constant for the second (slower) phase O3decay
***
Sum = kalginic acidx 0.4 + kNOM x 0.2 + kPeptone x 0.4
**
Cyclophosphamide (CPD) removal
The removal of CPD was recorded until all applied ozone was consumed (i.e. 5 mg/L);
however, under different experimental conditions, CPD concentration decreased below the
HPLC-MS limit of detection (100 ng/L). Therefore, CPD removal is presented herein for an
applied ozone dose of only 3 mg/L (i.e. until dissolved ozone reached 2 mg/L). Figure 3
presents the CPD removal as a function of DOC concentration, for the different EfOM model
compounds and alkalinity concentrations, at an applied ozone dose of 3 mg/L.
Cyclophosphamide (CPD) removal in a buffered water (without organic matter), for low
and high alkalinity levels was approximately 92% and 58% respectively. Generally, CPD
removal decreased with increase in DOC concentration (with an exception of alginic acid at
10
high alkalinity and DOC 0.5 mg/L), and in some cases it reached a plateau at higher DOC
values. The influence of peptone on CPD removal was most pronounced, where almost no
CPD removal can be seen at DOC ≥ 5 mg/L; while in the presence of alginic acid CPD
removal decreased only moderately with increase in DOC concentration.
Figure 3: Cyclophosphamide (CPD) removal as a function of DOC concentration for alginic
acid, NOM and peptone, at alkalinity concentrations of 25 and 200 mg/L. Ozone dose 3 mg/L.
Due to the low reaction rate of CPD with molecular ozone (kO3,CPD = 2.8 1/M s; Lester et al.,
2011), its removal during the ozonation process occurs mainly through oxidation by •OH
radicals (k•OH,CPD = 1.3 x 109 1/M s; Lester et al., 2011). Thus, the differences in the model
compounds’ impact on the CPD removal can be explained by differences in the •OH
production yield and different scavenging effects on •OH radicals by the EfOM model
compounds (evaluated in the following section).
In ozonation of pure water, the yield for •OH formation through the radical chain reaction
was found to be ~55% (Staehelin and Hoigne, 1982). This relatively high yield may partially
14
explain the high CPD removal rate at low EfOM concentrations (where the radical chain
reaction is the dominant •OH formation mechanism). The high removal rate of CPD at high
concentrations of alginic acid may be due to the alginic acid’s promotion effect on •OH, as
described earlier. Different •OH production yields are expected when direct reaction of ozone
with the EfOM is the dominant mechanism (e.g. at high NOM and peptone concentrations).
For example, Notch et al. (2009) found that the •OH yield during the ozonation process of
wastewater effluent was ∼13%, while Mvula and von Sonntag (2003) calculated •OH yield
for direct reaction of ozone with phenol to be ~22%.
Higher CPD removal is obtained at low alkalinity values compared to high alkalinity values
for all EfOM model compounds at DOC concentration < 5 mg/L, due to the scavenging effect
of alkalinity on •OH radicals. At higher DOC levels (≥5 mg/L), a comparable CPD removal
rate can be seen at both alkalinity values, indicating that the scavenging effect of the EfOM
model compounds is dominant. In natural waters, alkalinity is considered a main •OH
scavenger; while in water with high concentration of organic matter (i.e. wastewater effluent),
its contribution to the water scavenging is less significant. Elovitz et al. (2000) estimated the
•OH scavenging of alkalinity in lake water to be approximately 50% of DOC scavenging;
while Nothe et al. (2009) have calculated this value to be ∼10% in wastewater effluent.
Rate constant for the reaction of •OH with the EfOM model compounds
UV/H2O2 degradation of pCBA, in the presence of different concentrations of the EfOM
model compounds, was used to calculate k•OH,EfOM. Degradation of pCBA involves direct UV
photolysis and indirect photo-oxidation by •OH radicals, as described in the following
equations (Rosenfeldt and Linden, 2004):
 ln
[ pCBA]t
/ t  k obs  k 'k pCBA,OH  OH ss
[ pCBA]0
(1)
k '   pCBA  k s , pCBA
(2)
15
k s , pCBA 


10 3   0p   pCBA   1  10  a  z
200300
a z

(3)
where, [pCBA]0 and [pCBA]t are initial pCBA concentration (M) and its concentration after
exposure time t (s), kobs and k’ are the observed (total) and direct-photolysis time-based
pseudo-first-order degradation rate constants of pCBA respectively (1/s). [•OH]ss is the
steady-state •OH radical concentration (M) and kpCBA,•OH is the second-order rate constant of
pCBA reaction with •OH, reported to be 5x109 1/M s (Buxton et al., 1988). ФpCBA is the
quantum yield for pCBA removal (0.0182 mole/Einstein; Lester et al., 2010), ks,pCBA is the
specific rate of light absorption by pCBA (Einstein /mole s). E 0p(λ) is the incident photon
irradiance (Einstein/cm2 s), εpCBA(λ) is the molar (decadic) absorption coefficient of pCBA
(1/M cm), a(λ) is the solution absorption coefficient (1/cm) and z is the depth of solution
(cm).
The steady-state concentration of •OH radical was calculated as the ratio of the formation of
•OH radicals to the destruction of the radicals (Rosenfeldt and Linden, 2004):
 OH ss
RForm
OH

k pCBA,OH [ pCBA]  k H 2O2 ,OH [ H 2 O2 ]  k EfOM ,OH [ EfOM ]
(4)
where, kH2O2,•OH (1/M s) and kEfOM,•OH (L/(molC) s) are the second-order rate constants of •OH
with H2O2 and the EfOM model compound respectively. RForm
OH is the rate of •OH formation
(M/s) and is calculated using equation 5, taking into account the UV light absorbance of the
irradiated solution.
RForm
OH  k s , H 2O2   H 2O2  [ H 2 O2 ]
(5)
17
where, фH2O2 is the quantum yield for •OH formation (QY = 1; Baxendale and Wilson, 1957),
and ks,H2O2 is the specific rate of light absorption by H2O2, calculated using equation 3
(modified for H2O2).
Substituting equation 4 into 1, inverting both sides and rearranging results in eq. 6.
k H O ,OH [ H 2 O2 ] k EfOM ,OH [ EfOM ]
RForm
OH
 [ pCBA]  2 2

k obs  k '
k pCBA,OH
k pCBA,OH
(6)
Figure 4 is a plot of RForm
OH /kobs-k’ vs. the EfOM model compounds’ concentration.
Multiplying the slope of the linear lines by the known value of kpCBA,•OH resulted in reaction
rate constants of •OH with alginic acid, NOM and peptone of 0.92, 0.95 and 1.30 x 10 8
L/(molC) s respectively. The calculated values are lower than most values presented in the
literature for EfOM constitutes. Moreover, the similar values obtained for alginic acid and
NOM are unexpected, in view of the differences in their characteristics (aliphatic vs.
aromatic). The Suwannee River NOM used in our study contains however relatively high
proportion of polar aliphatic substances in addition to humic and fulvic substances (Serkiz
and Perdue, 1990), which may partially explain its relatively low reaction rate constants with
•OH.
Westerhoff et al. (1999), using ozone as an •OH source, found •OH reaction rate constants
with Suwannee River humic and fulvic acids to be 8.1 and 3.7 x 108 L/(molC) s respectively.
A latter study by Westerhoff et al. (2007), using electron pulse radiolysis, presented values of
1.6 x 108 L/(molC) s for the reaction of •OH with Suwannee River fulvic acids, and an
average value of 2.23 x 108 L/(molC) s for seven DOM isolates from different sources. Myint
et al. (1987) measured the rate constant for the reaction of •OH with hyaluronic acid (a
carbohydrate polymer with some resemblance to alginic acid) to be 7x10 8 1/M s, expressed in
terms of the disaccharide repeating sub-unit (equivalent to approximately 3x108 L/(molC) s),
using pulse radiolysis.
15
Figure 4: RForm
OH /kobs–k’ as a function of DOC concentration for NOM, alginic acid and
peptone. Initial hydrogen peroxide concentration of 50 mg/L.
The reaction rate constants for peptone and alginic acid with •OH obtained in the present
study correlate well with the different CPD removal behaviors (Figure 3). Increasing the
concentration of the highly •OH reactive peptone results in a sharp decrease in CPD removal
rate, while addition of the less-reactive alginic acid only moderately decreases the CPD
removal rate. For NOM, the low reaction rate constant with •OH may provide an explanation
only if considering that a high portion of the •OH radicals reacting with NOM are scavenged
(i.e. do not promote the chain reaction by producing O2•-).
Conclusions
Influence of the EfOM model compounds on ozone decay

Peptone and NOM were highly reactive toward molecular ozone; therefore at high
peptone and NOM concentration ozone decay was fast and controlled by direct reaction
with the model compounds. Alginic acid was least reactive toward molecular ozone, thus,
ozone decay was relatively slow and controlled by a radical chain reaction.
19
Influence of the EfOM model compounds on CPD removal

Cyclophosphamide (CPD) removal occurred mainly through its reaction with •OH
radicals and decreased with increasing model compounds concentration (most
pronounced for peptone and least pronounced for alginic acid).

The sharp decrease in CPD removal at high peptone concentration likely results from the
low yield of •OH formation in the direct reaction of ozone with peptone and its
scavenging of •OH.

The moderate decrease in CPD removal with increasing alginic acid concentration is due
to the relatively high yield of •OH formation from ozone decay through the radical chain
reaction and, the role of alginic acid as an •OH promoter.
Influence of alkalinity

Increasing alkalinity decreased CPD removal rate at DOC concentration < 5 mg/L, due to
its scavenging of •OH. At higher DOC levels (≥5 mg/L), alkalinity had only a minor
effect on CPD removal, indicating that the scavenging effect of the Efom model
compounds was the dominant mechanism.
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23
CHAPTER II
Enhanced removal of PPCPs using pH modification coupled with UV photolysis
Introduction
Ultraviolet (UV) treatment is being increasingly used for disinfection of wastewater and
drinking water in North America, Europe, and numerous other countries around the world.
UV radiation can also be used to chemically degrade or break down organic micropollutants
via photolysis (photodegradation), a process in which a chemical species undergoes a
chemical change as the result of the absorption of photons (Legrini et al., 1993). If a molecule
absorbs a photon, it is then in an excited state and can more readily transform.
For most chemicals, direct UV photolysis alone is not a practical process for
degradation, however, numerous chemical contaminants of concern absorb UV at
wavelengths below 300 nm; hence can potentially undergo direct photolysis (Stefan and
Bolton, 2002). For example, UV was well established as the technology of choice to remove
N-nitrosodimethylamine (NDMA) from drinking water (Stefan and Bolton, 2002; Sharpless
and Linden, 2003). NDMA is a water contaminant of emerging concern in North American
and a potential human carcinogen. Wu et al. (2007) investigated the photodegradation of a
widely used herbicide, metolachlor, applying monochromatic (254 nm) UV light.
Approximately half of the metolachlor was degraded at UV fluence of 1000 mJ/cm2 (at pH
7.5) which is 25 times higher than typical UV dose at water treatment plants (WTPs) required
for disinfection.
Various studies found that pH of the treated solution affected direct photolysis of
different organic pollutants. Shemer et al. (2005) showed that UV photolysis rate of 3,5,6trichloro-2-pyridinol (a degradation product of the insecticide chlorpyrifos) increased with
increasing the solution's pH up to a constant maximum value of 6.40×10−3 cm2/mJ at pH 5,
thus was highly pH dependent within the pH range 2.5–5. Other researchers showed the
20
influence of pH on the photodegradation kinetics of the antibiotics tetracycline (e.g. Werner et
al., 2006), sulfadimethoxine (Lester et al., 2008) and the pesticides atrazine and bensulfuron
methyl (Lam et al., 2003).
Obviously, pH impacts the degradation kinetics of many micro-pollutants. However,
none of the previous studies have examined the use of direct UV photolysis together with an
artificial pH modification of the water to treat water polluted with PPCPs. Therefore, the main
goal herein was to examine the combination of direct UV photolysis and artificial pH
modification of the treated water (during the treatment itself), as a potential treatment
technology for water remediation. The specific goals were to (a) determine the
photodegradation kinetics of different PPCPs separately, in buffered water, using direct UV
light and pH adjustment and (b) demonstrate the potential of pH induced polychromatic UV
treatment to remove a mixture of PPCPs from groundwater.
Material and Methods
The examined compounds
Twelve organic compounds were selected for the study based on their environmental
relevance (Table 2.1). All compounds were purchased from Sigma-Aldrich. Stock solutions
(100 mg/L) were prepared in water or methanol according to the compound’s solubility.
Other high-performance liquid chromatography (HPLC)-grade solvents and chemicals
(acetonitrile, methanol, formic acid, sodium hydroxide, and water) were purchased from BioLab Ltd. (Jerusalem, Israel).
24
Table 2.1 – Selected target contaminants
Class
Compounds
Antibiotics
Ciprofloxacin (CIP), Enrofloxacin (ENR), norfloxacin
(NOR), trimethoprim (TMP), tetracycline (TC),
oxytetracycline (OTC), amoxicillin (AMX) and
sulfamethoxazole (SMX)
Anti-epileptic drug
Carbamazepine (CBZ)
Anti-inflammatory drugs
Ketorolac (KTR) and diclofenac (DCF)
Anti-microbial agent
Triclosan (TCS)
Experimental setup
All irradiation experiments were conducted using polychromatic UV light. Specifically,
photolysis was carried out in a bench-scale collimated beam apparatus, using a 0.45 kW
polychromatic medium-pressure (MP) Hg vapor lamp (Ace-Hanovia lamp, Ace Glass Inc.,
Vineland, NJ, USA). The treated solution (100 mL) was spiked with a known concentration
of the tested compound and then irradiated with constant gentle stirring. Samples were
withdrawn at appropriate intervals for chromatographic analysis (using HPLC-DAD).
UV dose calculation
UV dose (fluence) was determined by multiplying the average irradiance of the UV lamp
between 200 and 300 nm by exposure time. The average irradiance was calculated using a
calibrated spectroradiometer (spectral range 200-390 nm; USB4000, Ocean Optics, Dunedin
FL, USA) placed in the same position as the irradiated solution, using a procedure adapted
from Bolton and Linden (2003). The UV absorption coefficient of the treated solutions at
different pHs was measured in a UV-Vis spectrophotometer (Varian, Cary 100BIO, Victoria,
Australia) and the molar absorption coefficients for the target compounds were determined.
Experimental procedure
25
To achieve the goals of the research, work was carried out in two stages: (I) Determination of
optimal pH values for the photodegradation of each target contaminant, and (II)
Demonstration of the potential of pH enhanced photolysis to remove a mixture of compounds
from groundwater. Each of these stages is discussed in detail below.
I)
Optimal pH determination for the photodegradation of each compound
The optimal pH values for the photodegradation of each target contaminant were determined
by testing the photodegradation rates of each contaminant separately, at different pH values.
Experiments were conducted in phosphate buffered saline (PBS, 5 mM), and pH was adjusted
by adding formic acid or sodium hydroxide at the appropriate volume and concentration.
Once the pH of the buffer containing the contaminant (at an initial concentration of 1 mg/L)
was adjusted, the sample was subjected to irradiation to determine the pH value that leads to
the maximal photodegradation rate. It should be noted that the initial concentration of the
pharmaceuticals was higher than their concentration in aquatic environments; however, it was
low enough to minimize the compound's contribution to the total solution absorbance, such
that first-order photodegradation kinetics were obtained for all compounds.
II) Removal of a mixture of compounds from groundwater
The main goal of the second part of this study was to determine the potential of UV
photodegradation combined with artificial pH modification for the removal of a mixture of
two pharmaceuticals, i.e., sulfamethoxazole (SMX) and triclosan (TCS), from groundwater,
using their predetermined optimal pH values. The removal of SMX and TCS from
groundwater is of particular interest. Triclosan (TCS) has been identified by the EU water
framework directive priority substance list (together with diclofenac, ibuprofen, and clofibric
acid) as a future emerging priority candidate of particular environmental concern, due to its
wide usage (as cited by Ellis, 2006). Sulfamethoxazole (SMX) is one of the most commonly
detected antibiotics worldwide (e.g. Sacher et al., 2001; e.g. Barnes et al., 2008).
Groundwater samples were taken from three different aquifer basins in Israel: GW-1
and GW-3 were from a local carbonate aquifer in the upper Galilee (eastern slopes) and the
major carbonate aquifer in the eastern slopes of the Judea Mountain Ridge in central Israel,
27
respectively, both representative of the Judea group. GW-2 was taken from a local aquifer in
the Golan Heights, which is composed of interbeded carbonate and basalt layers. The
groundwater samples were spiked with the two target contaminant (initial concentration of 0.5
mg/L for each compound), adjusted to their optimal pH value, and then subjected to UV
irradiation. The general chemical composition (major ions) of the groundwater samples used
in this study is presented in Table 2.2
Table 2.2 - General chemical composition of the three groundwater types
GW-1
GW-2
GW-3
pH
Cl -
HCO3-
Ca ++
7.5
7.8
7.9
19
24
80
260
198
236
70
26
62
Mg ++
mg/L
16
18
31
NO3 -
Na +
SO4 -
17
15
33
9
32
46
7
6
28
Analytical methods
The target compounds were detected and quantified by HPLC (HPLC-Agilent 1100 series;
ACE-RP C18 column 2.5 mm×250 mm) equipped with a UV diode array detector and a mass
spectrometer (MS, Finnigan LCQ). The HPLC consisted of a microvacuum degasser, a binary
pump and a thermostatic column compartment. The flow rate was 0.5 mL/min and the
injected volume was 100 μL. The flow from the HPLC was passed through a split connector
at 60 μL/min of effluent introduced into the MS interface. The MS was used in positive
electrospray ionization (ESI) mode, where ions in the range of 70-500 m/z were registered in
the conventional scanning mode.
Data analysis
The degradation kinetics were expressed as a natural logarithm of the ratio of the
concentration (CH) remaining following a UV dose H (mJ/cm) to initial concentration (C0)
(Equation 2.1).
ln
CH
 k H
C0
(2.1)
25
The data were fitted using a linear regression approach resulting in pseudo-first-order reaction
kinetics which reflected the difference in photodegradation between samples. The UV dosebased degradation rate constant k (cm2/mJ) was calculated as the negative slope obtained
when the degradation was plotted logarithmically. The UV dose required for 90% degradation
of the target contaminants was used as a comparative parameter between compounds and
between pH values, to emphasize the improved energy consumption of the degradation
process due to pH modification. The UV dose required for 90% degradation was calculated
using the obtained value for k, by inserting the appropriate numbers:
ln
10%  2.3  k
100%
H
(2.2)
Results and discussion
Optimal pH determination
To better understand the effect of pH variations on photodegradation rate of PPCPs and other
organic micro-pollutants, several photodegradation tests were conducted on the selected
compounds at different pH values. The pH values were in the common environmental range
of 5 to 8. Degradation rates at the different pH values are exemplified in Figure 2.1, which
shows the photodegradation of TCS (in PBS) plotted against UV dose.
29
Figure 2.1- UV photodegradation of TCS (1 mg/L) in PBS (5 mM) as a function of UV dose,
at different pH values (5, 6, 7 and 8), obtained by MP collimated beam apparatus. The
dashed line represents 90% degradation.
pH strongly affected the rate of direct TCS photodegradation: lowering the pH of the UVirradiated solution from 8 to 5 resulted in a significant decrease in the pollutant’s degradation
rate (Figure 2.1); the UV dose required for 90% removal of TCS was above 400 mJ cm -2 at
pH 5, and 162 mJ cm-2 at pH 8. The acid dissociation constant (pKa) of TCS is approximately
7.9 (Equation 2.3) (Wong-Wah-Chung et al., 2007), and there may therefore be a relation
between the acid-base speciation of the compound and its photodegradation rate at different
pH values.
Cl
Cl
(2.3
O
O
)
pKa 7.9
Cl
HO
Cl
Cl
[TCS-] + [H+] [TCS0]
35
-
O
Cl
Key parameters in evaluating the rate of a photochemical reaction are the molar absorption
coefficient of the compound and its overlap with the emission spectra of the UV lamp
(together with the compound’s quantum yield). These parameters quantify the light
absorption of a compound at each wavelength. pH dependence of the selected compounds’
molar absorption coefficients was examined. Figure 2.2 illustrates the molar absorption
coefficients of TCS at different pH values in the range of 5 to 8, as well as the emission
spectra of the UV lamp. TCS absorbed light in the UV range at all examined pH values.
Slight differences in absorbance peaks at around 278 nm could be seen at pH values of 5 to 7,
whereas at pH 8, its absorbance peak was significantly shifted to the right, at 288 nm.
Moreover, the peak at pH 8 was higher (4410 1/cm M) than those at pH 5, 6 and 7 (4050,
3480 and 3500 1/cm M, respectively). The apparent shift in the molar absorption coefficients
of TSC between pH values (with a clear relation to its pKa) and consequently, the change in
the overlap of the compound's absorption coefficients with the emission spectra of the UV
lamp, could potentially explain the differences in the photodegradation kinetics of this
compound at different pH values.
Figure 2.2- Molar absorption coefficients of TCS at various pH values, demonstrated with the
emission spectra of the polychromatic UV lamp.
31
Similarly, Wong-Wah-Chung et al. (2007) found that TCS is more photodegradable at high
pH (i.e. 11.8) than at low pH (i.e. 5.6), using UV light at different wavelengths. Moreover, the
molar absorption coefficient of TCS at pH 5.5 peaked at 280 nm with a value of
approximately 4200 1/cm M, while at pH 11.8, a maximum value of 8300 1/cm M appeared at
291 nm. Other studies also confirm the pH dependence of the TCS molar absorption
coefficient and photostability (e.g. Mezcua et al., 2004).
The UV doses required for 90% removal, and the molar absorption coefficient peaks,
for the studied compounds (in PBS) at different pH values are presented in Table 2.3. UV
dose was calculated for wavelengths between 200 and 300 nm, as these are commonly used in
MP UV treatments (e.g. Bolton and Linden, 2003). It should be noted however, that various
compounds absorb light at wavelengths above 300 nm (e.g. ENR), and these wavelengths
may influence their photolysis rate.
For most of the examined compounds, the UV dose required to achieve 90% removal was
strongly dependent on the pH of the treated solution (with the exception of NDMA and
NDEA), and could often be related to the compound's pKa value. The extent of the pH
dependency and its correlation (i.e. positive or negative) with removal rate by irradiation
varied between compounds. For example, increasing the pH of the treated solution from 5 to 8
significantly reduced the UV dose required for 90% removal of NOR and ENR, from 3286
and 2091 mJ/cm2 to 319 and 852 mJ/cm2, respectively. For SMX, the opposite trend was
observed: changing the pH of the buffer from 5 to 8 increased the UV dose required to
achieve 90% removal (from 173 to 676 mJ/cm2). Thus, by optimizing the pH of the treated
solution, markedly lower UV doses can be exploited to achieve the same removal of a target
compound.
In addition, changes in molar absorption coefficient, due to pH modifications, were
observed for most compounds (Table 2.3), and may provide a partial explanation for the
changes in their degradation rates at different pH values. However, various compounds
showed significant changes in their photodegradability due to pH modification with no
32
associated changes in their molar absorption coefficient (e.g. CBZ). Therefore, various
mechanisms may be involved in the observed changes in the kinetics of the UV
photodegradation of each compound by pH modification (e.g. changes in quantum yield).
Table 2.3 - UV dose required for 90% removal, and the molar absorption coefficient peaks
for the studied compounds (in PBS), at different pH values
compound
Relevant
pKa
UV dose for 90% removal, mJ/cm2
(molar absorption coefficients peak, 1/cm M; wavelength, nm)
pH 5
pH 6
pH 7
pH 8
78
(4410; 288)
TCS
7.9
210
(4050; 279)
163
(3480; 280)
138
(3500; 282)
CIP
5.5, 7.7
2091
(20725; 277)
(6292; 315)
742
(19115; 274)
(6452; 319)
404
(17676; 271)
(6876; 324)
NOR
6.3, 8.7
4600
(40445; 277)
2091
(38079; 275 )
697
(36771; 272)
328
(38265; 272)
ENR
5.94
3286
(38194; 277)
(11512; 316)
2300
(36532; 276)
(11361; 316)
1045
(32844; 271)
(12615; 323)
1264
(32702; 271)
(13041; 323)
SMT
6.08
767
(21698; 268)
958
(23508; 268)
1045
(25990; 268)
885
(25538; 268)
SMX
5.7
112
(17234; 264)
226
(19259; 259)
411
(19615; 257)
451
(16324; 256)
TMP
6.6
6420
(6240; 272)
6570
(5622; 276)
1533
(5312; 278)
742
(6265; 287)
KTR
3.5
1150
(7031; 249)
(20284; 322)
1000
(6519; 249)
(20073; 323)
657
(6678; 249)
(20177; 323)
315
(6324; 246)
(19230; 322)
CBZ
13.9
10000
(11227; 285)
10000
(11334; 285)
5750
(11755; 285)
3594
(11939; 285)
ACR
5.5
23000
(53497; 255)
11500
(435420; 249)
7667
(438824; 249)
5750
(474354; 249)
AMX
6.9
1075
(9531; 229)
(1414; 272)
1095
(9805; 228)
(1125; 273)
710
(9720; 227)
(1241; 273)
708
(9819; 227)
(1696; 273)
TC
4.5, 7.3
3833
(14525; 275)
(14450; 358)
1045
(14898; 276)
(14918; 357)
575
(14854; 275)
(15131; 360)
343
(14848; 271)
(15879; 367)
OTC
3.3, 7.7
5750
(8187; 218)
(12671; 275)
2556
(8477; 218)
(12851; 276)
1643
(9320; 217)
(12940; 276)
33
DCF
4.0
120
(9620; 275)
113
(10342; 276)
112
(10164; 276)
97
(10328; 276)
It is important to note that even though the photolysis rates of several of the examined
compounds were strongly pH-dependent, direct photolysis was still ineffective for their
degradation: UV doses above 6765 and 10,474 mJ cm-2 were required for 90% removal of
CBZ and ACR, respectively (at their optimal pH value).
Groundwater experiments
The main goal of the second part of the study was to determine the potential of UV photodegradation in combination with artificial pH modification, for the removal of a mixture of
two pharmaceuticals (i.e. SMX and TCS) from groundwater, using the optimal pH values
determined earlier. Figure 2.3 demonstrates the UV dose-based photo-degradation rate
constants of the two pharmaceuticals, in PBS at varied pH values. SMX and TCS photolysis
rates show high pH dependency with opposite directions, where, increasing pH of the solution
from 5 to 8 leads to higher TCS degradation rate and lower SMX degradation rate.
Accordingly, the optimal pH values to achieve the highest photo-degradation rate for TCS
and SMX are 8 and 5, respectively.
A treatment of groundwater containing TCS and SMX in a mixture was suggested.
Due to the noticeable differences in their optimal pH values when measured separately, the
treatment was conducted under the following conditions: (a) Irradiation of the mixed solution
at the groundwater original pH (i.e. pH 7.5-7.9), for a span of approximately 1.5 min (UV
dose of 45 mJ/cm2), in order to maximize the TCS degradation; and (b) Modification of the
groundwater pH to a value of 5, followed by an additional irradiation time of approximately
4.5 min (UV dose of 135 mJ/cm2), to accelerate the degradation of SMX. A removal of the
mixture by 90% was considered only when the slowest degraded compound in the mixture
was removed by 90%.
30
0.04
TCS
SMX
k ( cm2 / mJ )
0.03
0.02
0.01
0
4
5
6
7
8
9
pH
Figure 2.3- UV dose-based photo-degradation rate constant for TCS and SMX, measured
separately in PBS (5 mM), as function of pH
The pH modification point was determined using linear programming (LP) which allows
optimizing the treatment, to achieve maximum removal of the mixture at minimum UV dose.
The linear programming goal was to minimize the objective function as follows:
Minimize:
H1  H 2
(2.4)
Subject to constraints:
 10% 
 k1  H 1  k 2  H 2  Ln

 100% 
(2.5)
 10% 
 k '1 H 1  k ' 2 H 2  Ln

 100% 
(2.6)
where, H1 and H2 are the UV dose applied (mJ/cm2) from time zero to pH modification point
and from the pH modification point to at least 90% removal of the mixture, respectively. k1
and k2 are the dose-based photo-degradation rate constants of SMX at pH 7.5-7.9 and pH 5
respectively, measured during preliminary irradiation tests on groundwater mixture of TCS
and SMX, under constant pH (e.g. Figure 2.4b). k'1 and k'2 are the degradation rate constants
of TCS under the same conditions.
34
UV absorbance of the water is considered when calculating average UV dose
(fluence) in a collimated beam apparatus. The UV dose is then the product of the average
irradiance and the residence time. Thus, ideally the dose-based photo-degradation rateconstants should be independent of the water quality (dissolved constituents), when
considered in dose calculations. Regardless, minor variations were observed in the rate
constants for SMX and TCS when measured in buffer and groundwater.
Figure 2.4a illustrates the removal of TCS and SMX in a mixture from GW-1, using the
irradiation procedure where the water pH was modified from 7.5 to 5 after UV dose of 45
mJ/cm2. While, Figure 2.4b demonstrates the removal of a TCS and SMX mixture from GW1, at constant pH values of 7.5 and 5 (no pH modification). It is clear from Figure 2.4a that
during the first irradiation period (at pH 7.5) the degradation of TCS was highly accelerated,
while SMX degradation was relatively slow. In the second irradiation set, where pH was
adjusted to 5, the opposite direction was observed. The amount of UV dose required for 90%
removal of the mixture was only 170 mJ/cm2, while, when the pH of the irradiated
groundwater was maintained constant (pHs 7.5 and 5; Figure 2.4b), a significant higher
amount of UV doses of 548 and 256 mJ/cm2 were needed for obtaining 90% removal of the
mixture. Therefore, by modifying the treated groundwater pH during the process, a highly
improved removal efficiency of the target mixture is accomplished (in comparison to UV
treatment at constant pH).
35
(a)
0
-0.5
TCS 5
TCS 7.5
SMX 5
SMX 7.5
Ln ( C H/C0 )
-1
-1.5
-2
90% removal
-2.5
-3
0
50
100
150
200
UV dose (mJ/cm2)
(b)
0
Ln ( CH/C0 )
-1
TCS 5
TCS 7.5
SMX 5
SMX 7.5
-2
90% removal
-3
-4
0
100
200
300
400
500
600
700
2
UV dose (mJ/cm )
Figure 2.4- UV dose-based photo-degradation of TCS and SMX, measured in a mixture (0.5
mg/L each compound), in GW-1, using MP collimated beam apparatus, at (a) pH
modification of the treated water from 7.5 to 5 after irradiation dose of 45 mJ/cm2; and, (b)
constant pH values 5 and 7.5. The dashed lines represent 90% degradation.
37
Table 2.4 presents the UV dose required to achieve 90% removal of TCS and SMX (in a
mixture), from three different groundwater types, using the aforementioned irradiation
procedures. Results indicate that modifying the groundwater pH during the photodegradation
process, from their original pH values i.e. pH 7.5-9.7 to pH 5, has vastly decreased the UV
dose required for removal of TCS and SMX in a mixture in relation to constant pH.
Consequently this pH modification during the treatment has improved the treatment
efficiency.
Table 2.4 - UV dose required to achieve 90% removal for TCS and SMX (in a mixture), in
different groundwater, using constant pH and pH modification during irradiation
UV dose for 90% removal (mJ/cm2)
Constant pH
pH modification
during the treatment
pH 7.5-8
pH 5
GW-1
548
256
170
GW-2
511
247
162
GW-3
325
509
179
Practical application
The influence of pH on direct photolysis rate of PPCPs can be applied to actual practices of
UV water treatment and to design of a pH modified UV based photolytic system. The first
step in treating water polluted with multiple pollutants (more than one) using the proposed pH
enhanced photolysis is the determination of the optimal pH values for each pollutant or group
of pollutants. Then, a sequential pH optimization can be used for an optimized photolytic
treatment of the water, which includes several steps as follows: (a) first pH modification of
the water to optimize the degradation of one pollutant type, UV irradiation of the water and
(b) a second pH modification to optimize the degradation of the second pollutant type and a
second exposure to UV radiation. Figure 2.5 is a schematic diagram illustrating a ‘series’ type
exemplary system for treating a mixture of pollutants in water via pH optimized photolysis.
35
Pollutants A + B
in water
Input
1st pH adjustment
for A
1st direct photolytic
UV reactor
2nd pH adjustment
for B
2nd direct photolytic
UV reactor
Output
Fig 2.5- Sequential pH optimization
Photo-degradation intermediates analysis
When suggesting an engineered photodegradation-based treatment to remove PPCPs from
water, it is highly important to evaluate the influence of the photo-generated intermediate
products on the treatment. It is commonly accepted that intermediates formed during process
only marginally influence the treatment’s efficiency, due to their low environmental
39
concentration. However, active degradation intermediates may occasionally be formed, in
particular if the active part of the molecule remains unmodified, thus adversely impacting the
quality of the treated water (e.g. Sisson et al., 1997). The importance of this topic is
demonstrated in Table 2.5, presenting 12 byproducts identified herein during direct UV
photolysis of CIP.
The moieties which are involved in the CIP photodegradation are the piperazine
residue, the carboxylic group and the fluorine atom. The quinolone skeleton however remains
unchanged. The piperazine ring is either subjected to breakdown (intermediates 1-2, 4-5 and
8-12) or is epoxidized at positions 2, 3 (of the piperazine) (intermediate 6). Decarboxylation
occurs at position 3 of the quinolone skeleton (intermediate 7). The fluorine at position 6 is
either substituted by hydroxyl group (intermediate 3) or by hydrogen atom (intermediates
1,2,4,5 and 12). Various researchers have already identified the formation different by
products as a result of CIP photo-degradation (e.g. Vasconcelos et al., 2009).
In light of the fact that the characteristic quinolone is believed to be responsible for
the CIP antimicrobial activity (Dodd et al., 2006), intermediates which merely involve
modification of the piperazine residue may still potentially exhibit antibacterial potencies.
Table 2.5- HPLC/MS analysis of CIP direct UV degradation byproducts
Nº
(m/z)
CIP
332.1
Rt† Proposed product
structures
4.9 HN
N
Nº
(m/z)
Rt† Proposed product
structures
N
O
F
O
1
2.9
(316.1)
H2N
OH
N
7
7.0
(288.1)
O
N
HN
N
OH
N
F
O
O
O
2
6.1
(316.1)
8
3.4
(334.1)
O
NH
N
O
H2N
N
O
N
NH
OH
OH
O
O
O
05
F
3
2.8
(330.1)
9
6.8
(334.1)
HN
N
N
O
NH
N
O
NH
HO
O
O
OH
O
OH
F
4
7
(330.1)
10
4.0
(306.1)
O
NH
N
O
N
H2N
OH
H3C
N
O
NH
OH
O
O
F
5
5.5
(344.1)
O
NH
N
O
N
11
7.5
(263.1)
H2N
N
OH
O
F
O
O
O
6
4.4
(346.1)
O
12
3.2
(288.1)
HN
N
N
H2N
OH
N
O
O
F
NH
O
†
OH
O
OH
MS retention time (min)
01
CHAPTER III
A highly efficient bismuth-based photocatalyst for the degradation of PPCPs
under solar light irradiation
Introduction
The use of semiconductors in combination with sunlight irradiation (i.e. photocatalysis) for
the treatment of water and wastewater has attracted growing attention and intense research
interest over the last decade. TiO2 is the most widely used photocatalyst, mainly due to its
high efficiency, photochemical stability, nontoxic nature and low cost (Hoffmann et al.,
1995).
TiO2 has been widely applied to remove trace micropollutants (e.g. PPCPs) from
water (Doll and Frimmel, 2005a; Reyes et al., 2006; Yang et al., 2008) and wastewater
effluent (e.g. Hapeshi et al., 2010; Miranda-Garcia et al., 2011), and as a water disinfectant
(Sjogren and Sierka, 1994; Rincon and Pulgarin, 2004). The main limitation of TiO2 is a
relatively wide band gap that results in about 5% spectral overlap between its absorbance and
sunlight emission (λ<~390 nm). Different attempts to overcome this drawback have used
physical and chemical means such as morphological modifications to increase active surface
area (e.g. Wong et al., 2011), doping TiO2 with different elements, such as nitrogen (e.g.
Cheng et al., 2011) and use of other semi-conductors such as WO3 and ZnO, used with TiO2
or as independent catalysts (Lin et al., 2008; Martinez et al., 2011).
Bismuth-based photocatalysts showed effective solar-light activity, due to their
potential absorbance in the UVA and visible range and their non-toxic nature (Rohr, 2002).
Fu et al. (2005) found that Bi2WO6 was more efficient than TiO2 (Degussa P25) in the
degradation of the dye molecule Rhodamine B under visible light, and Wang et al. (2010)
showed its potential to remove the endocrine disrupting compound Bisphenol A under solar
simulated irradiation. Other bismuth-based catalysts such as CaBi2O4 (Tang et al., 2004) and
02
NaBiO3 (Kako et al., 2007) also exhibited reactivity under visible light, degrading dyes and
other organic pollutants.
A recently new development of bismuth-based catalyst is the bismuth-oxyhalides
(common formula BiOX, X=F, Cl, Br, I). These photocatalysts structure comprise of a layer
of [Bi2O2]+2 slabs interleaved by double slabs of halogen atoms. Zhang et al. (2006)
demonstrated the advantage of BiOCl over TiO2 (Degussa P25) in the decomposition of the
dye molecule methyl orange under UVA irradiation. Other studies established the potential of
different mono-halogen compounds such as BiOF (Su et al., 2010), BiOI (Li et al., 2011) and
BiOBr (Feng et al., 2011) to degrade various pollutants in water under UV and visible light
irradiation. Enhanced photoreactivity was reported with composite materials such as xBiOI(1-x)BiOCl (Wang et al., 2007) and xBiOBr-(1-x)BiOI (Wang et al., 2008), where the formed
hetero-junction improved the separation of electron-hole pairs and inhibited charge
recombination.
Shenawi-Khalil et al. (2011) were the first to demonstrate the photoreactivity of a
new family of bismuth-mixed oxyhalides catalyst with a general structure of BiOCl 1-xBrx (0 ≤
x ≤ 1). Specificaly, the bismuth catalyst (with x = 0.5) exhibited a higher reactivity than
Degussa P25 (to remove the dye molecule Rhodamine B), using similar catalyst
concentration, under both UV-vis (λ ≥ 385 nm) and visible light (λ ≥ 420 nm). Recently,
Gnayem and Sasson (2012) have optimized the synthesis of the BiOCl1-xBrx catalyst mainly
by using surface active quaternary ammonium salts as bromide and chloride sources, and
performing the synthesis under acidic conditions. These modifications resulted in a far more
effective photocatalyst, which totally decomposed Rhodamine B within 120 seconds (with
x>0.87) under visible light irradiation. However, this newly optimized BiOCl1-xBrx
photocatalyst was never examined to remove water micropollutants, such as PPCPs.
The goal of the present study is to determine the photocatalytic efficiency of this
catalyst to degrade frequently used PPCPs under simulated solar irradiation, and to examine
the influence of different experimental parameters.
03
Material and methods
Materials and reagents
All PPCPs were obtained from Sigma-Aldrich (>99% purity), LC-MS grade methanol and
water were from Bio-Lab Ltd. (Jerusalem, Israel). All chemicals were used as obtained and
stock solutions were prepared by dissolving the compounds (separately) in deionized (DI)
water (Direct-Q3 UV system, Millipore-France) at a concentrations of 50 mg/L. Degussa P25
(21 nm particle size, ~50 m2/g BET area), a commercially available TiO2, was used as a
reference catalyst.
Preparation and characterization of BiOCl0.875Br0.125
Synthesis of the new BiOClxBr1-x catalyst is detailed elsewhere (Gnayem and Sasson, 2012).
In brief, the catalyst was prepared from bismuth nitrate and surface active quaternary
ammonium salts (such as cetyltrimethylammonium halides), in the presence of acetic acid as
a co-solvent. The received powder, composed of hierarchiral flower-like microspheres
(Figure 3.1), was filtered and washed with water and ethanol and dried at ambient conditions.
The ammonium salts served both as bromide and chloride sources as well as structure
directing agents. The acidic conditions allowed the complete dissolution of the bismuth nitrate
and an instant reaction with the halide anions at room temperature, yielding the desired
bismuth oxyhalide product.
00
A Cl-/Br- ratio in the catalyst of 0.875/0.125 produced the highest photocatalytic
activity for the degradation of Rhodamine B under visible light (Gnayem and Sasson, 2012).
Therefore, the optimized formula BiOCl0.875Br0.125 was used in the present study.
Figure 3.1- SEM image of the as-prepared BiOCl0.875Br0.125 . Adated from Gnayem and Sasson
(2012)
Photodegradation experiments
Solar irradiation experiments were carried out in a 150-W xenon arc lamp solar simulator
(Sciencetech Inc, SS150W, Canada). The light beam was filtered with an air mass (AM) 1.5
Global filter that simulates the total (direct and diffuse) solar spectrum equivalent to natural
sunlight at 48.2° latitude at sea level. Irradiance was measured using a calibrated
spectroradiometer (International Light, Model ILT 900R, USA), placed in the same position
as the irradiated solution. Total incident irradiance, integrated between 280-950 nm was 710
W/m2, including UVA irradiance of 21.5 W/m2 and UVB irradiance of 0.9 W/m2.
Experiments were conducted with 30 mL aqueous solution (DI water) containing the
target pharmaceutical (initial concentration of 1 mg/L) and known concentrations of the tested
catalyst in suspension, at an inherent initial pH of ~ 5. For specific experiments borate buffer
(5 mM) was added to control pH variations during photocatalysis, where pH was adjusted
using HCl or NaOH. The examined solution was stirred for 60 min in the dark to ensure
adsorption/desorption equilibrium of the tested pharmaceutical on the catalyst prior to
irradiation. All experiments were performed in triplicate and relative standard deviations were
less than 10%. Water temperature remained approximately constant at 22 (±2) °C during all
04
photocatalytic experiments. Samples of 0.5 mL were taken periodically and the catalyst
suspension was separated from the aqueous phase by centrifugation (15 min, 14,600 rpm).
The supernatant was then analyzed using HPLC to quantify the tested pharmaceutical
concentration.
Chemical analysis
Pharmaceuticals were detected and quantified by HPLC Agilent 1100 (ACE-RP Phenyl
column 2.1mm×250mm) with a diode array detector. The flow rate was set to 0.5 mL/min and
the volume injected was 50 µL. The mobile phase consisted of water (A) and methanol (B), at
pH 3. The mobile phase eluent gradient started with 60% of eluent A, followed by a 4-min
linear gradient to 10% of eluent A, 4-min isocratic elution and a 2 min linear gradient back to
60% of eluent A, maintained for 5 min to equilibration time. The concentration of
formaldehyde was measured using the Nash reagent (2 M ammonium acetate, 0.05 M acetic
acid, 0.02 M acetylacetone) (Nash, 1953). In this assay, the reagent is mixed with equal
volume of the tested sample (after catalyst removal) and heated for 10 min at 50°C in the
dark. Formaldehyde was determined from its absorption at 412 nm (7600 1/M cm).
Concentration of H2O2 was determined by the molybdate-activated iodide method (Klassen et
al., 1994).
Absorbance measurements
True absorbance measurements of BiOCl0.875Br0.125, accounting for scattering by particles of
the catalyst suspended in DI water, were performed with a UV-Vis dual beam
spectrophotometer (Varian, Model Cary 100BIO, Victoria, Australia) equipped with 150 mm
diameter integrating sphere-IS attachment (Labsphere Diffuse Reflectance accessory (DRA)CA-30) and a center mount sample holder used to position the sample inside the IS. The
turbid sample (suspension with catalyst) is placed in a 1 cm path length quartz cuvette with all
four windows optically polished. The cuvette is fixed to a spring loaded holder that hangs in
the center of the sphere, connected to the top sphere cover.
05
Results and discussion
Parametric study
The goal herein was to examine the influence of different experimental parameters on the
BiOCl0.875Br0.125 photocatalytic process, through the degradation of carbamazepine (CBZ), a
widely used antiepileptic drug. Studies have already determined that CBZ resist conventional
water and wastewater treatments, with removal efficiencies by wastewater treatment plants
(WWTPs) below 10% (e.g. Clara et al., 2004). In the classification scheme for pharmaceutical
biodegradation, CBZ has the status of “no-removal” (Joss et al., 2006).
Catalyst reactivity
The degradation of CBZ versus time at a concentrations of 0 (no catalyst) and 500 mg/L
BiOCl0.875Br0.125, is presented in Figure 3.2. Direct photodegradation of CBZ after 5 min
exposure was negligible (0 mg/L catalyst). Adding 500 mg/L BiOCl0.875Br0.125 to the sample
dramatically increased its removal rate, demonstrating the high photo-reactivity of
BiOCl0.875Br0.125 under simulated solar light. In all cases, the photocatalytic degradation of
CBZ followed apparent first-order kinetics.
1
0 mg/L
0.8
CBZ (C/C0)
Bi-500 mg/L
0.6
P25-500 mg/L
0.4
0.2
0
0
1
2
3
4
5
Time (min)
Figure 3.2- Carbamazepine (CBZ) degradation (initial concentration 1 mg/L) under
simulated solar light vs. time with 0 and 500 mg/L BiOCl0.875Br0.125 (and comparison to P25
500 mg/L), in DI water at pH 5.
07
The BiOCl0.875Br0.125 photo-reactivity was compared to Degussa P25, a commercially
available TiO2, using 500 mg/L concentration of either catalyst, at pH 5 (Figure 3.2). The
photocatalytic removal of CBZ after 5 min of irradiation reached ~66% and 80% with P25
and BiOCl0.875Br0.125 respectively. These results were obtained under specific experimental
conditions, and a detailed comparison between BiOCl0.875Br0.125 and P25 is beyond the scope
of this study.
Reactivity of the catalyst was fully maintained for five consecutive irradiation cycles
(data not shown). Where, with each new cycle, 1 mg/L of CBZ was freshly added to the same
catalyst suspension after the former batch was 99% degraded (after 10 min).
Light absorption by the catalyst and effect of irradiation wavelength
The standard method used to measure absorbance relies on transmittance of light captured by
a detector that is placed in line with the sample, using a spectrophotometer. The drawback of
this method is that particles which scatter light at angles outside the reception angle of the
detector will result in significant error in absorbance measurements (e.g. Du and Rabani,
2004).
True
UV
absorbance,
can
be
measured
using
integrating
sphere
(IS)
spectrophotometers, optical devices that integrate the radiant flux of most reflected and
transmitted radiation simultaneously (Mamane et al., 2006).
1.2
120
Ab…
100
Abssorption coefficient
(1/cm)
1
80
0.6
60
0.4
40
0.2
20
0
200
0
250
300
350
400
Wavelength (nm)
Irradiance (uW/cm2)
0.8
450
Figure 3.3- Irradiance of the solar simulator lamp and absorption coefficient spectra of 500
mg/L aqueous suspensions of BiOCl0.875Br0.125 (in DI water, pH 5), measured in the
integrating sphere.
05
Figure 3.3 illustrates the spectra obtained from the solar simulator and the absorption
coefficient spectra of 500 mg/L aqueous suspensions of BiOCl0.875Br0.125, measured using the
integrating sphere (between 200-450 nm). It is evident that there is a general decrease in the
catalyst’s light absorption with increase in wavelength. Where, above approximately 380 nm,
absorption coefficient is negligible for the specific bismuth formulation studied.
In light of these newly acquired absorption data, we estimate that the degradation of
Rhodamine B under visible light, previously reported by Gnayem and Sasson (2012), was
initiated through the absorption of light by the dye, rather than by the catalyst.
Figure 3.4- Photo-catalytic removal of CBZ with 500 mg/L BiOCl0.875Br0.125 (in DI water, pH
5) at different wavelengths, after light exposure of 10 J/cm2.
The effect of the irradiation wavelength on the photocatalytic degradation of CBZ was further
investigated, using two long pass filters which screened all light below 320 and 400 nm,
respectively. Specifically, the examined wavelength ranges were: λ<280 nm (no filter), λ<320
nm and λ<400 nm. Photo-degradation experiments were carried out in the presence of 500
mg/L BiOCl0.875Br0.125 (DI water, pH 5). The exposure time at each wavelength range was
determined in order to achieve an identical light power density of 10 J/cm2, which was
09
calculated as the product of the exposure time (sec) and the measured incident irradiance (for
each tested range), integrated over the entire UV-visible range.
Carbamazepine (CBZ) was only marginally degraded under visible light irradiation
(~5%), while extending the irradiation spectrum to the UVA and UVB range increased the
removal of CBZ to 68% and 85%, respectively (Figure 3.4). These results were expected
since BiOCl0.875Br0.125 does not absorb light at wavelengths longer than ~380 nm (Figure 3.3).
Quantum yield (ф) calculation
Quantum yield (ф) for the degradation of CBZ with 500 mg/L BiOCl0.875Br0.125 (DI water at
pH 5) was calculated as the ratio between its initial degradation rate and the rate of photon
absorbed by the catalyst, as follows (Dalrymple et al. 2010):

r '0
 ( E ( )  (1  10 a( )z ) / z)

(3.1)
0
p
where, r’0 is the initial degradation rate of CBZ (M/s), Ep0 is the measured photon incident
irradiance on the surface of the sample (millieinstein/s cm2), z is the depth of the sample (cm)
and a is the absorption coefficient of the solution (1/cm), measured with the integrating
sphere (Figure 3.3) (assuming light absorption by CBZ is negligible at these wavelength;
Lester et al., 2012). The summation of Equation 3.1 was taken over the wavelength range
between 280–380nm, accounting for the overlap between the catalyst absorbance and the
lamp emission (Figure 3.3).
Quantum yield for CBZ degradation was found to be 0.75% (±0.05). A similar
quantum yield was calculated when a long pass filter was utilized, screening all light below
320 nm, thus integrating wavelengths between 320-380 nm (data not shown). This implies
that the efficiency of an absorbed photon to excite an electron in the BiOCl 0.875Br0.125 catalyst
is similar in the UVA and UVB range. The reason for the higher degradation rate of CBZ at
λ>280 nm (than at λ>320 nm), presented in Figure 3.4, is therefore the increase in light
45
absorption by the catalyst at shorter wavelengths (Figure 3.3), rather than an increase in its
quantum yield.
Little information exists in the literature on quantum yields for the photocatalytic
degradation of organic pollutants. The main reason is that absorbed photons are
experimentally difficult to measure, due to light scattering by the suspended particulates.
Therefore, most studies use the term quantum yield (or photonic efficiency) referenced to
incident photons, which merely represents the lower limit of the actual quantum yields
(Serpone, 1997). For example, incident irradiance was used to calculate the quantum yield for
the degradation of the antibiotic sulfamethoxazole (0.21-0.75%; Xekoukoulotakis et al., 2011)
and the pesticide formetanate (0.5%; Marinas et al., 2001).
Effect of catalyst and CBZ concentrations
The catalyst concentration is an important operating parameter in a suspended photocatalysis
system. Carbamazepine (CBZ) first-order degradation rate constant (k, 1/min) steadily
increased with addition of BiOCl0.875Br0.125 at concentrations up to 500 mg/L (Table 3.1). With
increasing catalyst concentration above 500 mg/L, a slightly reduced rate constant was
observed. Different studies found optimal concentrations for suspended catalysts in the range
of 75-1000 mg/L, depending on the type of catalyst and target pollutant, on the pollutant's
concentration and on the spectral photon flux (Yang et al., 2008; Achilleos et al., 2010;
Hapeshi et al., 2010; Martinez et al., 2011).
Table 3.1- First-order rate constants for CBZ degradation at different catalyst concentrations
R2
Catalyst concentration
k
(mg/L)
(1/min)
0
-
50
0.062 (±0.011)
0.996
100
0.107 (±0.016)
0.998
250
0.206 (±0.016)
0.994
500
0.303 (±0.007)
0.986
750
0.274 (±0.008)
0.989
41
At low catalyst concentrations, increase in the degradation rate is commonly explained by
increase in available active catalysts sites and light absorption. While at high concentration,
the catalyst provides a screening effect that reduces light penetration into the irradiated
solution. Figure 3.5 demonstrates the screening effect of BiOCl0.875Br0.125, where, absorption
coefficient of the catalyst solution (exemplified for λ = 300 nm) increased with its
concentration up to 750 mg/L, when practically all the light other than the backscattered is
absorbed.
The discrepancy between the optimal catalyst concentration for CBZ degradation (i.e.
500 mg/L) and the catalyst concentration with the highest light absorption (i.e. 750 mg/L) can
be explained by other mechanisms such as internal scattering, deactivation of sites through
collision with ground-state catalysts (Neppolian et al., 2002) and agglomeration of the
catalyst at high concentrations (So et al., 2002).
Absorption coeff. at 300 nm (1/cm)
1.6
1.2
0.8
0.4
0
0
100
200
300
400
500
600
700
800
900
1000
BiOCl0.875Br0.125 concentration (mg/L)
Figure 3.5- Absorption coefficient measured at 300 nm with the integrating sphere as a
function of BiOCl0.875Br0.125 concentration. Measurements were taken in DI water at pH 5.
The influence of initial CBZ concentration on its initial degradation rate is demonstrated in
Figure 3.6. Increasing CBZ concentration increased its degradation rate until it approaches a
42
plateau. This behavior is in good agreement with the Langmuir–Hinshelwood (L–H) model,
commonly used to describe photo-catalytic degradation of organic compounds:
r0  (
k  K ads  [CBZ ]0
d CBZ 
) 0  in
dt
1  K ads  [CBZ ]0
(3.2)
where, r0 is the initial degradation rate of CBZ (mM/min), kin is the intrinsic reaction rate
constant (mM/min), Kads is the L-H adsorption constant of CBZ on the catalyst surface
(1/mM) and [CBZ]0 is the initial concentration of CBZ in the aqueous solution.
Figure 3.6- Initial degradation rate of CBZ (r0) as function of its initial concentration (C0).
The inset presents 1/r0 vs. 1/[CBZ]0 for different initial CBZ concentrations
The L-H model is usually plotted as the inverse of the pollutant’s initial degradation rate
(1/r0) vs. the inverse of its initial concentration in the solution (1/C0), as follows:
1
1
1
1



r0 k in k in  K ads [CBZ ]0
(3.3)
The linear reletionship between 1/r0 and 1/[CBZ]0 (inset of Figure 3.6) suggests that the L-H
model adequatly describes the kinetics of the examined photocatalytic process. From the
43
slope and intercept of the linear line, kin and Kads were calculated to be 0.25 mM/min and
261.7 1/mM, respectively. A much lower Kads of 28.9 1/mM was found by Martinez et al.
(2011) for the degradation of CBZ using P25 TiO2 and near UV–vis light, suggesting that
adsorption is more important in the degradation of CBZ with BiOCl 0.875Br0.125 than with P25.
This topic is further evaluated in the following chapter.
Effect of pH
Figure 3.7 illustrates the degradation rate of CBZ, in the presence of 500 mg/L
BiOCl0.875Br0.125, at different pH values in the range of 4-9. Obviously, photocatalysis rate
decreases with increasing pH up to pH 9, with negligible photodegradation rate at a pH value
of 9.
0.5
0.4
k (1/min)
0.3
0.2
0.1
0
3
4
5
6
7
8
9
10
pH
Figure 3.7- First-order rate constant for CBZ degradation with BiOCl0.875Br0.125 (500 mg/L)
vs. pH. Experiments were conducted in borate buffer (5 mM).
pH may alter the pollutant’s charge and the catalyst surface charge which can further impact
the adsorption kinetics and degradation mechanism of the target pollutant (e.g. Hapeshi et al.,
2010; Wang et al., 2010). Furthermore, with TiO2, variation in pH can influence the
production rate of hydroxyl radicals (•OH) (e.g. Lin et al., 2008). For example, in the
40
presence of TiO2 and UVA irradiation, paracetamol degradation rate slightly increased with
increasing pH from 3.5 to 9.5, presumably due to enhanced •OH formation (Yang et al.,
2008). On the other hand, the reaction rate significantly decreased at pH 11.0, mainly due to
increase in electrostatic repulsion between the TiO2 surface (negatively charged at pH > 6.3)
and paracetamol (negatively charged at pH above 9.5). Fu et al. (2005) showed that raising
solution pH from 6.5-10, reduced the adsorption degree and correspondingly the degradation
rate of Rhodamine B using BiWO3 photo-catalysts under visible light irradiation.
To understand the influence of pH on the photodegradation of CBZ, the % adsorption
of CBZ at different pH values was determined (Table 3.2). The charge of the BiOCl 0.875Br0.125
surface (zeta potential), taken from Gnayem and Sasson (2012), is also presented. Adsorption
of CBZ (1 mg/L) on BiOCl0.875Br0.125 (500 mg/L) was measured in the dark after 60 min of
equilibration time.
Table 3.2- Influence of pH on CBZ degradation rate constant (k, 1/min), (dark) adsorption of
CBZ and zeta potential of the catalyst
pH
4
k
(1/min)
0.43
Adsorption
(%)
26.5
*Zeta potential
(mV)
-0.01
5
0.32
19.5
-4.64
6
0.23
16.1
-9.13
7
0.15
14.5
-12.47
8
0.03
8.2
-14.55
9
0.01
7.0
-17.36
*Adapted from Gnayem and Sasson (2012)
The zeta potential was negative at all pH values from 5 to 9. Decreasing solution pH from 9
to 4 shifts the charge on the catalyst surface from negative to neutral (point of zero charge –
PZC at pH 4) and is correlated to the increase in CBZ (dark) adsorption. Carbamazepine
(CBZ) is neutrally charged and relatively hydrophobic (log Kow = 2.45, Trenholm et al.,
2006) at the entire tested pH range, while the catalyst is most hydrophobic at pH 4 (PZC).
Therefore, decreasing the solution pH from a value of 9 to 4 will increase the hydrophobic
44
attraction between CBZ and the catalyst, increasing CBZ adsorption and, ultimately,
enhancing its degradation.
Effect of water constituents
It is generally accepted that non-target organic solutes in water may inhibit the
photodegradation of a target pollutant by the combination of light attenuation (for light
absorbing solutes), competition for reactive species and surface deactivation of the catalyst by
adsorption (Doll and Frimmel, 2005b). To evaluate the influence of natural water constituents
on the BiOCl0.875Br0.125 photocatalytic process, degradation of CBZ was examined in the
presence of different inorganic ions and natural organic matter. Experiments in this section
were conducted by adding different concentrations of various model compounds to a solution
of 500 mg/L BiOCl0.875Br0.125 and 1 mg/L CBZ (in borate buffer 5 mM, pH 7).
Influence of natural organic matter
Suwannee River fulvic acid (SRFA) was used to simulate natural organic matter
(International Humic Substances Society-IHSS; St. Paul, MN). While addition of SRFA to the
solution (up to 20 mg/L) clearly inhibited the degradation of CBZ (Figure 3.8), it did not
affect the CBZ adsorption to the catalyst, which remained unchanged even at 20 mg/L SRFA
(results not shown). Dissolved fulvic acid have a yellow-brown color and strong spectral
absorbance in the UV range (Frimmel, 1994), therefore, we assume that the inhibiting effect
of SRFA on the photodegradation of CBZ results mainly from its role as an inner filter for
incoming photons (rather than competition for active sites).
45
1
0.8
CBZ (Ct/C0)
0.6
0 mg/L
2 mg/L
0.4
6 mg/L
12 mg/L
0.2
20 mg/L
0
0
2
4
6
8
10
Time (min)
Figure 3.8- Photodegradation of CBZ with BiOCl0.875Br0.125 (500 mg/L) at different
concentrations of SRFA. Experiments conducted in borate buffer at pH 7.
Influence of main ions
Different inorganic ions were added (separately) to the irradiated solution at typical natural
water concentrations (Table 3.3).
Table 3.3- Examined ions
Examined ions
Compound
added
Ion concentration
range (mg/L)
Ca2+
Ca(OH)2
0-150
Mg2+
Mg(OH) 2
0-75
Na+
NaCl
0-400
Cl-
NaCl
0-600
SO42-
Na2SO4
0-500
NO3-
NaNO3
0-100
Na2HPO4
0-10
NaHCO3
0-370
PO43HCO3
-
47
While addition of Ca2+, Mg2+, Na+, Cl-, SO42- and NO3- did not affect the degradation rate of
CBZ (results not shown), increasing the concentration of PO43- up to 10 mg/L highly reduced
the CBZ removal rate (Figure 3.9). The influence of HCO3- could not be determined due to
high pH variability between samples of different HCO3- concentrations and variations in pH
that occurred during irradiation.
Abdullah et al. (1990) already demonstrated the inhibiting effect of phosphate (PO43-)
on TiO2, where, the photodegradation rate of salicylic acid, aniline, and ethanol was markedly
decreased with the addition of PO43- to the solution (at ~100 mg/L). The effect of phosphate
on TiO2 was mainly attributed to its high adsorption degree to the catalyst.
1.2
1
CBZ (Ct/C0)
0.8
0.6
[PO4] = 0 mg/L
[PO4] = 2 mg/L
0.4
[PO4] = 6 mg/L
0.2
[PO4] = 10 mg/L
0
0
2
4
6
8
10
time (min)
Figure 3.9- Photodegradation of CBZ with BiOCl0.875Br0.125 (500 mg/L) at different
concentrations of PO43-. Experiments conducted in borate buffer at pH 7.
To further examine the influence of PO43- on BiOCl0.875Br0.125 photocatalysis, the extent of
CBZ (dark) adsorption (%) was measured in the presence of different PO43- concentrations
(Figure 3.10). Clearly, increasing the concentration of PO43- reduced the adsorption of CBZ to
the catalyst (by competing over adsorption sites), presumably reducing its degradation rate.
45
The mechanism for PO43- adsorption is further evaluated at section 3.3.2. It should be noted
that PO43- does not absorb light in the UV range, thus, its inhibiting effect cannot be attributed
to light attenuation.
14%
CBZ Adsorption
12%
10%
8%
6%
4%
2%
0%
0
2
4
6
8
10
12
14
16
[PO43-] mg/L
Figure 3.10- Adsorption of CBZ in the dark vs. PO43- concentration. Experiments conducted
in borate buffer at pH 7.
Reactive species
It is well known that photo-excitation of a semiconductor produces electron-hole pairs which
are highly mobile and quickly recombine, in competition to their trapping (localization) at the
semiconductor surface. Oxidation of a target pollutant occurs either through direct reaction
with valance band holes (h+) or through the intermediate of reactive species such as •OH
radicals (e.g. Hoffmann et al., 1995).
The purpose of this section is not to determine the specific reactive species generated
by BiOCl0.875Br0.125 under solar irradiation, but rather to assess their "oxidative power" and
their similarity to those generated by TiO2. Oxidation of methanol (CH3OH) to formaldehyde
(HCHO) (Equations 3.4-3.5) was used as an indicative step for the photogeneration of highly
reactive species (Goldstein et al., 2008). Oxidizing the aliphatic methanol may occur by either
49
holes or radicals, as long as their oxidation potential is high enough (redox potential of
methanol ~ 1.34V; Goldstein et al., 2008). Experiments were carried out in a 30 mL aerated
suspension (DI, pH 5) of BiOCl0.875Br0.125 (500 mg/L) and methanol (2 M).
h+/•OH + CH3OH
•CH2OH + O2
H+/H2O + •CH2OH
•OOCH2OH
HCHO + H+ + O2•-
(3.4)
(3.5)
The results clearly show a linear accumulation of formaldehyde with irradiation time (Figure
3.11). The linear relationship also suggests that formaldehyde does not react with the reactive
species at the high methanol concentration used. A slower rate of formaldehyde formation is
observed when CBZ and methanol are tested together (as opposed to methanol alone), due to
competition between the two target compounds for the reactive species. A near linear
accumulation of formaldehyde was previously demonstrated for different catalysts (Marugan
et al., 2006; Goldstein et al., 2008), indicating the presence of highly reactive oxidizing
species.
In bismuth based photocatalysis, photo-excitation of valence band Bi+3 generates Bi+5
(or Bi+4) holes, with an estimated standard redox potential of ~1.59V (at pH 0) (Fu et al.,
2005). These holes are usually considerd as the main oxidative species in the
photodegradation of organic pollutants (e.g. Shenawi-Khalil et al., 2011), and we estimate
they are also responsible for the degradation of methanol (and CBZ) in the present study.
55
180
with carbamazepine
[Formaldehyde] uM
150
w/o carbamazepine
120
90
60
30
0
0
5
10
15
20
25
Irradiation time (min)
Figure 3.11- Formaldehyde formation vs. irradiation time in a BiOCl0.875Br0.125 (500 mg/L)
suspensions (DI, pH 5), in the presence of 2M methanol, with and without CBZ (1 mg/L).
Since oxidation of the aliphatic methanol (CH3OH) to •CH2OH is known to be relatively slow
compared to other aliphatic and aromatic compounds (as cited by Goldstein et al., 2008), our
results imply that the BiOCl0.875Br0.125 photocatalytic process will not be limited to a small
number of specific organic compounds.
Addition of H2O2
Photocatalytic degradation of micropollutants can be accelerated by the addition of hydrogen
peroxide (H2O2), as demonstrated for TiO2 by several researchers (So et al., 2002; Hapeshi et
al., 2010). Hydrogen peroxide (H2O2) can promote the formation of reactive species through
direct reaction with the photogenerated electrons or indirectly by reacting with superoxide
ions:
e- + O2
H2O2 + O2-•
H2O2 + e-
O2-•
HO2•
(3.6)
•OH + OH- + O2
(3.7)
•OH + OH-
(3.8)
51
Figure 3.12 demonstrates the influence of added H2O2 (in the range of 0.05-2.5 mM) on the
BiOCl0.875Br0.125 (500 mg/L) photodegradation rate of CBZ (DI, pH 5). Results are presented
as the rate constant for CBZ degradation with H2O2 relative to its degradation rate without
H2O2.
Relative CBZ rate constant
(kwith H2O2/kw/o H2O2 )
1
0.8
0.6
0.4
0.2
0
0
0.5
1
1.5
2
2.5
3
Added H2O2 (mM)
Figure 3.12- BiOCl0.875Br0.125 photodegradation of CBZ (presented as the degradation rate
constant with H2O2 relative to the rate constant without H2O2) vs. added H2O2 concentration.
Experiments were conducted in DI at pH 5.
Addition of H2O2 at low concentrations (≤ 0.1 mM) did not significantly affect the
degradation rate of CBZ. Whereas, increasing H2O2 concentration to 0.5 mM sharply
decreased the CBZ degradation rate, followed by a plateau up to 2.5 mM H2O2. Apparently,
the mechanism for enhanced •OH production by H2O2, previously suggested for TiO2, does
not apply for BiOCl0.875Br0.125. The decrease is CBZ degradation rate at H2O2 concentrations ≥
0.5 mM may be explained by H2O2 competition with CBZ over BiOCl0.875Br0.125 adsorption
sites and/or oxidizing species, as demonstrated below (Goldstein et al., 2009):
H2O2 + h+
2H+ + O2•-
(3.9)
52
Indeed, measurements during the experiments showed up to 30% adsorption of H2O2 to the
catalyst after 60 min in the dark (with no clear correlation to H2O2 initial concentration).
Moreover, concentration of H2O2 in the solution was reduced by up to 50% following 5 min
of irradiation.
Degradation of additional PPCPs: Ibuprofen, Bezafibrate and Propranolol
BiOCl0.875Br0.125 was further evaluated for its ability to degrade other pharmaceuticals of
different therapeutic classes, namely: bezafibrate-BZF (lipid regulator), ibuprofen-IBF (antiinflammatory) and propranolol-PPL (beta blocker). To better simulate real water treatment,
experiments were conducted in borate buffer at pH 7 (typical for natural waters) and catalyst
concentration of 200 mg/L (to minimize particles interferences). The target pharmaceuticals
were selected due to their frequent detection in WWTPs and in the aquatic environment (e.g.
Ternes, 1998), and their different charges at the pH examined. At pH 7, BZF (pKa 3.6) and
IBF (pKa 4.4) were negatively charged, and PPL (pKa 9.67) was positively charged. Figure
3.13a presents the first-order degradation rate constant for the selected compounds, (tested
separately at C0 = 1 mg/L), while Figure 3.13b shows the degree of the compounds’
adsorption (%) to the catalyst, measured in the dark prior to irradiation. Results for CBZ are
presented for comparison, and direct photolysis of all compounds was negligible (data not
shown).
53
(a)
(b)
Figure 3.13- (a) First-order rate constant for the degradation of CBZ, PPL, BZF and IBF
with 200 mg/L BiOCl0.875Br0.125 (in borate buffer 5mM, pH 7) and (b) Adsorption degree of the
compounds measured in the dark prior to irradiation.
All examined compounds were efficiently degraded by the BiOCl0.875Br0.125, validating its
potential to treat pharmaceutical contaminated water. Degradation rate of the different
compounds followed a decreasing order of: PPL > BZF > IBF > CBZ (highest removal for
PPL). Interestingly, while BZF and IBF were highly adsorbed to the catalyst, adsorption of
PPL was negligible (Figure 3.13b). Thus, the importance of (dark) adsorption to the
photodegradation process is compound-dependent. Another observation relates to the
different adsorption mechanism of the examined compounds. The negatively charged BZF
and IBF are most likely adsorbed to the positively charged [Bi2O2]+2, suggesting an ion
50
exchange mechanism; while PPL does not adsorb due to its positive charge. The ion exchange
mechanism may also explain the adsorption of PO43- to the catalyst (observed at section
3.3.1.6), and its inhibiting effect on the CBZ degradation rate.
References
 Achilleos, A., E. Hapeshi, N. P. Xekoukoulotakis, D. Mantzavinos, and D. FattaKassinos, "Factors affecting diclofenac decomposition in water by UV-A/TiO(2)
photocatalysis", Chemical Engineering Journal, 161 (1-2): 53-59 (2010).

Batt AL, Kim S, Aga DS (2007) Comparison of the occurrence of antibiotics in
four full-scale wastewater treatment plants with varying designs and operations.
Chemosphere, 68, 428-435.

Bolton, J.R.and Mihaela, I.S. (2002). Fundamental photochemical approach to the
concepts of fluence (UV dose) and electrical energy efficiency in photochemical
degradation reactions, Res. Chem. Intermed., 28, 857–870.

Buxton, B. V., Greenstock, C.L., Helman, W.P and Ross, A.B. (1988). Critical
review of rate constants for reactions of hydrated electrons, hydrogen atoms, and
hydroxyl radicals in aqueous solution. J. Phys. Chem. Ref. Data 17, 513–886.

Hapeshi, E.A. Achilleos, M. I. Vasquez, C. Michael, N. P. Xekoukoulotakis, D.
Mantzavinos, and D. Kassinos, (2010) Drugs degrading photocatalytically: Kinetics
and mechanisms of ofloxacin and atenolol removal on titania suspensions", Water
Research, 44 (6): 1737-1746.

Huber, M.M., Canonica, S., Park, G.Y. and Gunten, U.V. (2003). Oxidation of
pharmaceuticals during ozonation and advanced oxidation processes. Environ.
Sci.Technol. 37, 1016–1024.

Mamane, H. Shemer, H. and Linden, K.G. (2007) Inactivation of E.coli, B. subtilis
spores, and MS2, T4, and T7 phage using UV/H2O2 advanced oxidation, Journal of
Hazardous Materials, 146, 479-486.

Martinez, C., M. Canle, M. I. Fernandez, J. A. Santaballa, and J. Faria (2011)
Kinetics and mechanism of aqueous degradation of carbamazepine by
heterogeneous photocatalysis using nanocrystalline TiO(2), ZnO and multi-walled
carbon nanotubes-anatase composites, Applied Catalysis B-Environmental, 102 (34): 563-571
54
CHAPTER IV
The Interaction between Ozonation and Wastewater Particles
Introduction
Ozone was shown to be an effective barrier against emerging trace organic contaminants
(TrOCs), and is increasingly used to remove these compounds from municipal and industrial
wastewater effluent (Ternes et al. 2003; Huber et al. 2005; Bahr et al. 2007; Hollender et al.
2009; Wert et al. 2009; Lester et al. 2013). The extent of TrOCs removal is typically
determined by the applied ozone dose, the (second-order) reaction rate constant between
ozone and the target contaminant (kO3,TrOC, M-1s-1) and the wastewater composition (Huber et
al. 2003; Wert et al. 2009; Nöthe et al. 2009). Among the different wastewater constituents,
dissolved organic matter and nitrite (which readily react with ozone) are considered the main
influencing parameters (Staehelin & Hoigne 1985; Paraskeva & Graham 2002). The influence
of particulate matter on the other hand is less clear.
Interactions of ozone with particles during drinking water application were investigated for
several years (Bourgine et al. 1998; Chandrakanth & Amy 1996; Jasim et al. 2008; Currie et
al. 2003). Several researches have reported that following ozonation, particle count decreases,
particle size increases, zeta potential increases and turbidity removal increases (Jekel 1983;
Georgeson & Karimi 1988; Farvardin & Collins 1989; Chandrakanth et al. 1996; Jasim et al.
2008). Few mechanisms were suggested for those phenomena such as polymerization of
NOM, reduction in adsorbed organics molecular weight and/or lysis of algae followed by
release of coagulating biopolymers (Grosso & Weber 1988; Chandrakanth & Amy 1996).
Other studies found that interaction between ozone and wastewater particles may also take
place (Jekel 1994; Zhu et al. 2008; Genz et al. 2011). For example, reaction of dissolved
ozone with wastewater particles was suggested to contribute to the particles destabilization,
polymerization and subsequent removal by coagulation (Jekel 1994).
55
Ozone-particles interaction mainly depends on the properties of the particles. For example,
biopolymers (proteins and polysaccharides) particles were found to be transformed to smaller
biopolymers fragments following ozonation (Genz et al. 2011). The particle size may also
affect the interaction since ozone can aggregate fine particles and break down large ones (Yan
et al. 2007). Other particles characteristics such as electrostatic interaction, dispersion forces
and hydrophobic bonding were also found to play an important role in the of the particles are
also can be readily affected by ozonation, (Zhu et al. 2008). Other influencing mechanisms
may include ozone absorption by particles, sorption of some compounds to particles and/or
interaction with colloids (Holbrook et al. 2004; Huber et al. 2005).
Very little studies are available on ozone-particles interactions during wastewater ozonation,
and their influence on the removal of TrOCs. It was shown that wastewater suspended
material only marginally affected ozone degradation of TrOCs (Huber et al. 2005); however,
total suspended solids (TSS) bulk parameter was used, with no additional fractionation. The
overall goal of this study was to examine the interaction of ozone with particles of different
sizes, during wastewater ozonation. This was done by evaluating the influence of different
size particles on ozone degradation of selected TrOCs, as well as the changes in wastewater
bulk parameters and the counter-effect of ozone on wastewater particles of different sizes.
Materials and methods
Standards and Reagents
Six trace organic compounds (TrOCs) and were analyzed during experiments and can be
categorized in two groups; fast (kO3 > 104 M-1s-1), and slowly reacting (kO3 < 10 M-1s-1)
compounds. Diclofenac, Sulfamethoxazole, Carbamazepine, and Iohexol (>99% purity) were
obtained from Sigma-Aldrich. Iopromide and Iopamidol (>99% purity) were provided from
Holland-Moran, Israel. Pharmaceutical subclass, CAS numbers, rate constants and analytical
data are summarized in table 1. LC-grade methanol and water were purchased from Bio-Lab
(Jerusalem, Israel). All chemicals were used as obtained and working solutions were prepared
with deionized (DI) water (Direct-Q3 UV system, Millipore).
57
Experimental Procedure
To determine the interaction of ozone with particles of different size, secondary wastewater
effluent from the Shafdan site (the largest wastewater treatment plant in Israel) was filtered
using filters with different pore-size (11 μm, 6 μm, 2.5 μm paper filters and 1.2 μm, 0.45 μm
glass filters, Whatman). Particle analysis showed that approximately 80% of particles above
the pore size were removed by filtration. To ensure detection before and after ozonation, the
TrOCs standards were spiked before starting the experiments at the initial concentration of
100 µg/L. Bench-scale ozonation experiments were conducted in 500 mL stirred glass
vessels. Ozone stock solution (~40 mg/L) was prepared by continuously bubbling ozone gas,
produced from an oxygen fed generator (up to 4 g/h, BMT 802N, Germany), in a glass reactor
filled with chilled deionized water. Ozone dose is described as specific ozone consumption
(Zspec) by normalizing applied ozone doses (mg/L) to the initial dissolved organic carbon
(DOC0) concentration. Effluents with different sizes of particulate content were ozonated
using three different Zspec - 0.17, 0.69, and 0.93 mg O3/mg DOC0. After complete reaction
with ozone, ozonated samples were analyzed for bulk parameters, TrOCs, and particle size
distribution. Experiments were repeated in three different sampling events during the period
of October 12 to June 13.
Analytical Methods
Ozone measurements
Ozone concentration in the bubbling ozone gas was measured by an ozone gas analyzer
(BMT 963, Germany). Ozone concentration in the stock solution was measured by the indigo
method (APHA, method 4500B).
TrOCs sample analysis
The target compounds were detected and quantified by HPLC (Agilent 1100 series; ACE-RP
phenyl column 2.1 mm×250 mm) equipped with a UV diode array detector and a mass
spectrometer (QTof MS, Waters Premier). The column temperature was 40 ºC, the flow rate
was 0.5 mL/min and the injected volume was 100 µL. The HPLC mobile phase consisted of
water (A) and methanol (B), adjusted to pH 3 by the addition of formic acid. The mobile
55
phase eluent gradient started with 10% eluent B for 1 min, followed by an 4 min linear
gradient to 90% B, a 5 min isocratic elution at 90% B and a 2 min linear gradient back to 10%
B, maintained for 4 min for equilibration prior to next run. The flow from the HPLC was
passed through a split connector with 60 µL/min of effluent introduced into the MS interface.
The mass spectrometer was used in ESI positive mode. Removal efficiencies during
ozonation were calculated from the concentrations in spiked secondary and ozonated effluent.
Analysis of wastewater-quality parameters
Total organic carbon (TOC) of the wastewater was measured using a TOC analyzer (Torch,
Teledyne Tekmar, OH, USA). To measure DOC, samples were filtered at 0.45 μm (APHA,
method 5310B). Ultraviolet absorbance at 254 nm (UVA) was measured via UV-Vis
spectrophotometer (Varian, Cary 100 BIO, Victoria, Australia) for 0.45 μm filtered samples.
The turbidity was measured by a conventional 90º side-scatter instrument. Other wastewater
parameters were analyzed using standard methods (APHA et al. 2005).
Particle analysis
Particles suspended in liquid were analyzed by the “Micro Flow Imaging” technology (DPA
4100, ProteinSimple Inc, Ottawa Ontario, Canada). This apparatus employs a digital camera
with an illumination and magnification system to capture in-situ images of suspended
particles in a flowing sample. Basically, a sample fluid is drawn through a flow cell and
sections of the fluid are illuminated with light emitting diode light source at 470 nm
wavelength, magnified and imaged onto a digital camera. These captured images are
automatically analyzed to determine various size and shape parameters that represent the two
dimensional projection of the particles. Analysis was conducted for particle size between
2.25–400 µm.
Table 1. Pharmaceutical subclass, second-order rate constants for reaction with ozone (kO3) at pH 7
and CAS registry No. of spiked target compounds
59
MS
kOH
kO3
[M-1s-1]
CAS
Registry
No.
fragmentation
[109 M-1s-1]
Pharmaceutical
Subclass
296.023
7.5
6.8*105
1530779-6
anti-inflammatory
Diclofenac
254.059
5.5
5.5*105
723-46-6
sulfonamide
antibiotic
Sulfamethoxazole
237.102
8.8
5*105
298-46-4
antiepileptic
Carbamazepine
66108Iohexol
95-0
95551791.877
3.1
0.8
contrast media
Iopromide
09-5
60166777.861
2.8
4.1
Iopamidol
93-0
References: (Sein et al. 2008; Dodd et al. 2006; Huber et al. 2003; Baus et al. 2004)
821.888
5.5
4.1
Results and Discussion
Particle size and distribution in Shafdan secondary effluents
Turbidity is the most common parameter used to monitor particles. However, it does not
provide information on the size, shape, and concentration of particles and thus may not be the
suitable measurement of particle removal efficiency (Mamane et al. 2008). In order to extent
the understanding of particle distribution, a dynamic image analysis of particles was
conducted. The particle size distribution (PSD) was measured during October 2012-June 2013
in 10 different trails. The size of natural particles may be simplified to one parameter by
defining particles as spheres and isolating the Equivalent Circular Diameter (ECD). The PSD
is shown in table 2 in seven different size ranges as well as the total count in the PSD
instrument analysis range (2.25–400 µm). Correlation of turbidity and particle counts has
been widely studied (e.g. (Bourgine et al. 1998)) and a direct relationship between turbidity
and 2-3 um particle concentration was seen here as well (data not shown). It can be seen that
during the Israeli winter (December 2012 to March 2013), the total particle concentration was
higher due to when the biological activated sludge treatment is not in optimum conditions.
However, in February the particle concentration decreased, probably due to dilution of
rainfalls. Similar trend was seen in the different size ranges up to 30 µm. the concentration of
higher particle size was hence the accuracy reduced and comparison was difficult to operate.
75
Opposite phenomenon was found in UK as particle concentration wad higher during the
summer which was attributed to high algae concentration during summer (Bourgine et al.
1998).
Table 2. NFE’s PSD in different dates during Oct 2012-June 2013
3.6.13
2.6.13
6.3.13
12.2.13
24.1.13
17.1.13
26.12.12
29.11.12
5.11.12
15.10.12
5,521
2,077
4,021
912
322
60
0
12,912
5,233
1,871
3,429
888
247
34
0
11,702
28,189
12,100
10,662
5,049
4,185
692
134
61,011
12,471
4,199
14,493
2,071
680
0
0
33,915
29,763
10,362
10,253
5,877
1,790
266
62
58,372
26,909
9,907
6,473
2,391
744
105
28
46,558
28,285
11,786
8,716
2,244
1,028
190
30
52,279
19,315
4,350
6,925
1,957
1,438
308
63
34,356
19,759
3,533
4,872
1,023
654
70
15
29,926
21,563
7,694
5,665
1,463
780
135
45
37,346
Influence of Particles on the Ozone Treatment – Degradation of the Target Pollutants
The degradation of the target TrOCs during ozonation was determined for filtered and
non-filtered effluent (NFE). Carbamazepine, diclofenac and sulfametaxazole, with ozone
second-order rate constants higher than 104 M-1s-1 (fast-reacting), were removed below the
limit of detection (~100 ng/L) at Zspec of 0.69 mg O3/mg DOC0. Other studies also showed
that compounds with kO3>104 M-1s-1 were eliminated to concentrations below the detection
limit (~1-100 ng/L) for an ozone dose of 0.47 mg O3/mg DOC0 (Hollender et al. 2009).
Figure 1 presents the removal of the three slow-reacting iodinated contrast media by ozone, at
0.69 and 0.93 mg O3/mg DOC0, for different pore size filtered effluents. It can be seen that
filtration with decreased pore-size filters improved the ozone removal rate of the iodinated
contrast media with similar trend for both ozone dosages. For example, iopamidol was
eliminated by 71% at NFE compared to 80% at 11 µm and 93% at 0.45 µm filtered effluents
in 0.93 mg O3/mg DOC0. With any filter pore size, increasing Zspec from 0.69 to 0.93 mg
O3/mg DOC0 enhanced iodinated contrast media removal. Furthermore, experiments with
high dosage (0.93 mg O3/mg DOC0) of ozone showed smaller error values in comparison to
low dosage (0.69 mg O3/mg DOC0). From studies that investigated the fate of the selected
TrOCs during activated sludge treatment, it was clear that sorption onto suspended solids is
71
ECD size
(µm)
2-3
3-5
5-10
10-15
15-30
30-50
50+
Total
not a relevant process (Ternes et al. 2004). The improved removal with decreasing pore-size
filtration indicates that particles interfere with ozone treatment, probably by competing with
dissolved matter over ozone. The three iodinated contrast media have negligible rate constants
for the reaction with ozone (table 1) and are mainly oxidized by hydroxyl radicals. Therefore,
it can be concluded that hydroxyl radical exposure increases with filtration, most likely since
ozone reaction with dissolved organic matter generates more hydroxyl radicals (higher
quantum yield) than ozone reaction with particles (on surface or via ozone intra-particle
diffusion). Possibly, scavenging effect of fast-reacting hydroxyl radicals by the particles
through particle-surface reactions may also influence the removal of micropollutants.
72
Figure 1. Removal of iodinated contrast media by 0.69 and 0.93 mg O3/mg DOC0 ozonation in different pore size filtered effluents
73
Influence of Particles on the Ozone Treatment – Water Quality Parameters and Ozone
Depletion
The effect of particles on water quality parameters during ozonation was tested. The effluent
had a pH of ~7.5 and a DOC concentration of ~9 mg/L which was not influenced by filtration
or by ozone treatment. However, DOC changes might not be detected due to low sensitivity
of the TOC analyzer or due to formation of smaller DOC organic fractions (not mineralized).
Ozone depletion in secondary effluents can be described as a reaction in two phases; a very
fast depletion from 0-20 seconds, operationally defined as instantaneous ozone demand
(IOD), followed by first order ozone depletion (Buffle et al. 2006; Nöthe et al. 2009). Figure
2 presents the decrease in UVA following ozonation at 0.93 mg O3/mg DOC0, for the
different pore-size filtration as well as IOD. Decrease in effluent UVA following ozonation is
expected due to ozone attack on conjugated systems (Wert et al. 2009). Direct correlation can
be seen between particle filtration and UVA decrease by ozone; where, increasing particle
fractions removal (by decreasing filter pore-size) increases UVA removal. It can also be seen
from figure 2 that increasing particle fractions removal decreases IOD. This result fits well
with the UVA and TrOCs removal. With decreasing pore size filtration, particle load
decreases, less ozone consumed by particles and ozone reacts more with conjugated system
and TrOCs. This strengthens the conclusion that particles in wastewater indeed compete with
dissolved matter over applied ozone (or react with generated hydroxyl radicals), thus
influencing the ozone treatment efficacy.
Table 2 shows water quality parameters such as filtered biochemical oxygen demand (BODf filtered at 0.45 µm), filtered chemical oxygen demand (CODf -filtered at 0.45 µm), total
suspended solids (TSS), and volatile suspended solids (VSS). BOD increased after ozonation
as well as TSS and VSS. It was shown before that following ozonation, effluent's BOD and
the BOD/CODf ratio increases, which implies of an increase in the effluent's biodegradability,
which is highly desirable if ozonation is followed by a further biological treatment (Lester et
al. 2013). Change in particle count after ozonation as a function of pore size filtration will be
discussed in the next section.
70
Table 3. water quality parameters before and after moderate ozonation, in NFE and 1.2 µm filtered
samples
VSS
TSS
COD
BOD
O3 treatment
0.5
3.3
32
3.0
Before ozonation
1.4
3.5
32
6.1
0.69 mg O3/mg DOC0
0.5
0.3
29
1.5
Before ozonation
1.1
1.4
26
4.5
0.69 mg O3/mg DOC0
NFE
1.2 µm
Filtration
NFE 11u 6u 1.2uNFE 11u 6u 1.2u
0.45u, 6 µm, 0.45u, 1.2 µm,
51.04%
NFE 11u 6u 1.2u 50.72%
0.45u, 11 µm,
NFE 11u 6u 1.2u 48.35%
0.45u, NFE,
45.75%
IOD (mg/L)
% Removal UVA
NFE 11 µm 6 µm
NFE 11 µm 6 µm
NFE 11 µm 6 µm
NFE 11 µm
NFE6 11
µmµm 6 µm
1.2 µm 0.45 µm1.2
, µm 0.45 µm1.2
, µm 0.45 µmNFE
1.2
, µm
1.21.2u
µm 0.45
,
µm ,
11u0.45
6u
1, 7.0623809522, 6.8076190483, 6.612380952
4, 6.464761905
5, µm
5.990952381
0.45u,
0.45
,
56.11%
Figure 2. UVA removal (red bars) and IOD (blue markers) with standard deviations by 0.93 mg O3/mg
DOC0 ozonation in different pore size filtered effluents
74
Influence of Ozone Treatment on the Particles – Size Distribution
Non-filtered secondary effluent were ozonated with a dose of 0.93 mg O3/mg DOC0. Figure 3
shows the PSD before and after ozonation for NFE samples during summer and winter
experiments. The change in concentration (C) for each size range due to ozonation was
calculated. Image analyzer is not capable of analyzing particles smaller than 2 µm (e.g.
colloids, oxides, viruses, and bacteria), but can efficiently analyze larger particles (e.g.
bacteria, bioflocs, organic debris, algae, cysts and silt).
After ozonation, the total particle concentration in all of examined samples decreased. The
most significant particle concentration decrease (~80%) was observed for particles with ECD
range of 5-15 µm (medium size particles). A lower decrease was observed for 15-30 and 2-5
µm particle ranges. Similar trends were observed in summer and winter. Since O3 can act as a
flocculent-aid (Jekel 1994) by destabilizing particles, there may be formation of larger
particles (30–50+ µm) during ozonation process, as observed herein.
The population balance for specific sized particles in a suspension can be described by
integration of few mechanisms. Particles of each size class can be formed by flocculation of
smaller particles and lost by flocculating to larger particles. Moreover, particles of each size
class can be also formed by larger particles breakage and lost by breaking into smaller
particles making them more mineralized and easier to remove or even dissolute into dissolved
constituents (Zhu et al. 2008; Li et al. 2009). Surface dissociation, i.e., departure of ions from
the surface and their transfer to a bulk electrolyte solution (Ofir et al. 2007) was not detected
in DOC measurements possibly as the TOC analyzer is not sensitive to these changes.
This PSD analysis shows that both flocculation and particle-breakage can co-exist
simultaneously and their influence may be determined by the particles distribution among
other factors.
75
Winter BO,
AO, 2.75, 12,471
11,748
Winter BO
Winter AO
Summer BO
Summer AO
Particle Concentraion (#/ml)
Summer BO, 2.75, 5,233
Summer AO, 2.75, 4,615 Winter BO, 7.75, 4,644
Winter BO, 8.75, 3,782
Winter
Winter BO, 6.75,
2,311BO, 9.75, 2,326
Winter
Winter
BO,
BO,
2,100
Winter
AO, 4.75,
4.75,
2,041
Winter
AO, 3.75,
3.75,
1,9962,100
Winter BO, 5.75, 1,429
Winter
5.75,
Summer
BO,AO,
4.75,
9921,077
Summer
BO, 7.75,
884BO, 10.75, 879
Summer
BO, 3.75,
879
Winter
Summer
AO,
4.75,
7996.75,
Summer
AO,
3.75,
754
Winter
AO,
711
Summer
BO,535
8.75,
654 617
Summer
BO,
5.75,
653
Summer
BO,
6.75,
622
Summer
BO, 9.75,
Winter
AO,
7.75,
Winter
BO,
11.75,
478
Summer AO,
5.75,
415
Winter
AO,
8.75,
410
Summer
BO,
10.75,
393
Summer
AO,
8.75,
353
Winter
AO,
9.75,
352
Summer
AO,
7.75,
304
Summer
AO,
9.75,
283
Summer
AO, 6.75,
279
Winter
BO,
12.75,
276
Winter
AO,
10.75,
264
Summer
BO,
11.75,
235
Winter
BO,
13.75,
225
Winter
BO,
14.75,
213
Winter
AO,
11.75,
186
Winter
BO,
176
Winter
BO,
149
Winter
AO,
12.75,
142
Summer
BO,
12.75,
127
Summer
AO,
10.75,
124
Winter
AO,
112
Winter
BO,
100
Winter
AO,
92
BO,
18.75,
87
Winter
AO,
15.75,
87
Summer
AO,
11.75,
8613.75,
BO,
85
BO,
13.75,
74
Winter
AO,
16.75,
71
AO,
65
Summer
AO,
12.75,
5914.75,
BO,
14.75,
58
Summer
AO,
5215.75,
Winter
AO,
49
BO,
15.75,
45
Winter
AO,
4236
40
BO,
16.75,
37
BO,
BO,
19.75,
Summer
AO,
3616.75,
BO,
17.75,
36
35
21.75,
32
Winter
AO,
30
BO,
18.75,
30
Summer
Summer
AO,
AO,
2817.75,
27
Summer
AO,
2619.75,
26
22.75,
24
23
22
24.75,
21
23.75,
20
Summer
AO,
1820.75,
18
Summer
AO,
25.75,
27.75,
17 37
17
26.75,
16
15
28.75,
14
13
29.75,
11
10
30.75,
24.75,
25.75,
26.75,
9
23.75,
27.75,
8
31.75,
32.75,
22.75,
7
33.75,
29.75,
6
38.75,
28.75,
34.75,
35.75,
5
36.75,
39.75,
4
37.75,
40.75,
3
52.75,
Winter
AO,
41.75,
42.75,
2
50.75,
60.75,
80.75,
1
Summer
Summer
BO,
100.75,
101.75,
102.75,
103.75,
104.75,
105.75,
106.75,
107.75,
108.75,
109.75,
110.75,
43.75,
44.75,
45.75,
46.75,
47.75,
48.75,
49.75,
51.75,
53.75,
54.75,
55.75,
56.75,
57.75,
58.75,
59.75,
61.75,
62.75,
63.75,
64.75,
65.75,
66.75,
67.75,
68.75,
69.75,
70.75,
71.75,
72.75,
73.75,
74.75,
75.75,
76.75,
77.75,
78.75,
79.75,
81.75,
82.75,
83.75,
84.75,
85.75,
86.75,
87.75,
88.75,
89.75,
90.75,
91.75,
92.75,
93.75,
94.75,
95.75,
96.75,
97.75,
98.75,
99.75,
0
ECD Size (µm)
Figure 3. Particle concentration distribution (PSD) during winter and summer experiments before ozonation (BO) and after 0.93 mg O3/mg DOC0 ozonation (AO)
77
Influence of Ozone Treatment on the Particles – effect of pre-filtration & ozone dose
Filtered and non-filtered secondary effluent were ozonated with three different Zspec of 0.17,
0.69, and 0.93 mg O3/mg DOC0. The PSD analysis before and after ozonation is shown in
figure 4. When effluent was pre-filtered with 0.45 and 1.2 µm filtration, particle concentration
increased for any ECD size as function of Zspec. However, for 11 µm filtered effluent and
NFE, particle concentration increased only at 0.17 mg O3/mg DOC0. In 0.69 mg O3/mg DOC0
ozonation, a decrease was observed in particle concentration and for 0.93 mg O3/mg DOC0
ozonation, different effects seen in 11 µm filtered effluent and NFE. One possible reason
might be that in effluents with low particle concentration, the lower probability of ozoneparticle interaction results in efficient coagulation effect of dissolved matter while for nonfiltered samples, particle brakeage is the main mechanism. Another reason might be
difference is particle properties such as morphology, hydrophobic / hydrophilic nature and
molecular weight. It was reported before that hydrophobic NOM with intermediate molecular
weight increases at lower ozone dosage, while at higher ozone dosages, NOM becomes more
hydrophilic and its molecular weight becomes smaller, decreasing NOM removal (Yan et al.
2007).
Colloidal stability is often measured by zeta potential of the particles. However, the average
zeta potential was difficult to determine for a highly dispersed system containing an unknown
number of colloidal particles (Farvardin & Collins 1989).
75
Figure 4. PSD analysis before and after 0.17, 0.69 and 0.93 mg O3/mg DOC0 ozonation in different pore size filtered effluents
79
Potential application
Pre-filtration before wastewater ozonation is a possible application for enhancing the removal
of micropollutants. This was further examined by using an on-site deep-bed filtration field
pilot in the Shafdan WWTP. Particle analysis revealed similar filtration properties for both
deep-bed field filtration and 11 µm laboratory filtration. The removal of Carbamazepine
(CBZ, an antiepileptic drug) by ozonation from the Shafdan non-filtered and deep-bed filtered
effluent was examined. CBZ concentration in both effluents, as determined by an SPE-HPLCMS/MS method, was approximately 850 ng/L.
CBZ decomposition by ozonation process is presented in figure 5 for secondary and deep-bed
filtered effluents. CBZ was efficiently removed in both effluents, with improved removal
achieved in the deep-bed filtered effluent. Thus, pre-filtration may promote micropollutant
removal via ozonation, by removing particulate matter.
1
C/Co
0.8
Secondary effluent
Co = 840 ± 200 ng/L
Deep-bed filtered effluent
Co = 850 ± 200 ng/L
0.6
0.4
Not
detected
0.2
0
0
0.3
0.5
Specific Ozone Consumption (mg O3 /
mg DOC0)
0
0.3
0.5
Specific Ozone Consumption (mg O3 /
mg DOC0)
Figure 5. CBZ decomposition by ozonation in different specific ozone consumption for
secondary (left) and deep-bed-filtered effluents
OH-radical exposure was back-calculated using the oxidation of ozone resist TrOCs
(iodinated contrast media). OH radical exposure can be described as a linear function of the
ozone consumption (figure 6). This function is characterized by the slope indicating the
efficiency of OH-radical formation and lag ozone consumption without significant radical
formation, which can be calculated from intercept and slope as reported before (Hübner et al.
2013). It can be seen that the radical exposure for pre-filtered effluents is higher.
55
0.04
ʃ[OH]dt (109 Ms)
500
0.03
OH exposure
y = 0.0031x - 0.0017
R² = 0.8764
ASF
0.02
y = 0.0028x - 0.007
R² = 0.7763
0.01
0
0
2
4 O3 dose6 [mg/L] 8
10
12
Figure 6. OH-radical exposure determination using 3 iodinated contrast media as trace
substances in Shafdan secondary and deep-bed-filtered effluents (ASF)
In figure 7, relative residual concentration of fast and moderate reacting TrOCs is presented
as function of rate constants. in low applied dosages, a difference could be seen between the
different TrOCs relative residual concentration.
C/Co
0.8
Secondary effluents
0.6
0.4
0.2
0.0
0
0.5
30
ASF effluents
0.4
C/Co
10
20
kO3 (105 1/Ms)
0.3
0.2
0.1
0.0
0
10
20
kO3 (105 1/Ms)
51
30
Figure 7. Relative residual concentration of fast and moderate reacting TrOCs as function of
rate constants with ozone in Shafdan secondary and deep-bed-filtered effluents (ASF)
The attempt to calculate ozone exposure from reduction of different TrOCs is illustrated in
figure 8 as a function of second order rate constant with ozone (kO3). Data show a strong
correlation of exposure with rate constants (high data points from benzafibrate are not
shown). These results confirm that ozone exposures cannot be calculated from tracer
substances under the examined conditions. Thus, simple modeling of compound removal
calculated ozone
exposure (10-5 Ms)
from second order rate kinetics is not possible.
5
10 mg/L 500
4
3
8.57 mg/L
500
2
1
0
0
10
20
kO3 (105 1/Ms)
30
Figure 8. Calculation of ozone exposure using fast reacting TrOCs as function of rate
constant with ozone in Shafdan secondary effluents
Conclusions
The present study demonstrates the influence of wastewater effluent particles with different
size distribution on ozonation and the influence of ozone on the particles’ concentration.
Particle analysis showed a decrease in particles' count after ozonation for non-filtered
samples, which may indicate both flocculation and decomposition of organic particles into
smaller particles (not detected by the analyzer) or to dissolved organic matter. Medium size
particles concentration was mainly affected by the ozone treatment. Filtered samples showed
an increase in particle concentration after ozonation in at ECD sizes. UVA absorbance and
micropollutant concentration decreased more substantially after ozonation as function of pore-
52
size filtration, with increased removal for the 0.45 µm filtered effluent. It can be concluded
that the mutuality between particle size distribution and ozone treatment efficiency exist.
53
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