sonochemical degradation of pharmaceuticals and personal care products in water
TRANSCRIPT
Sonochemical Degradation of Pharmaceuticals and Personal Care Products
Dissertation
Presented in Partial Fulfillment of the Requirements for the Degree Doctor of Philosophy
in the Graduate School of The Ohio State University
By
Ruiyang Xiao, B.S., M.S.
Civil Engineering Graduate Program
The Ohio State University
2012
Dissertation Committee:
Linda K. Weavers, Advisor
John J. Lenhart
Harold W. Walker
Yu-ping Chin
Copyright by
Ruiyang Xiao
2012
ii
Abstract
The widespread use of pharmaceuticals and personal care products (PPCPs) has
raised environmental concerns due to their presence in aquatic environments, unknown
chronic low-dose exposure to humans, and recalcitrance to conventional water treatment
technologies. In this dissertation, ultrasound, especially in pulsed wave (PW) mode has
been explored to remove PPCPs. The focus was to fundamentally and mechanistically
understand how ultrasound degrades PPCPs and what controls degradation kinetics.
First, ultrasound was employed to degrade the pharmaceuticals, ciprofloxacin
(CIPRO) and ibuprofen (IBU), in the presence of Suwannee River fulvic acid (SRFA)
and terephthalic acid (TA) to gain an understanding of the effect of environmentally
relevant matrix organics on degradation kinetics. The matrix organics inhibited the
sonolysis of CIPRO and IBU to different extents. Based on the results, SRFA stays in
bulk solution, either quenching •OH and/or associating with the target compounds.
Similar to SRFA, TA, a commonly used •OH scavenger, reacts with •OH in the bulk
region but we also suspect it accumulates on or interacts with cavitation bubbles. The
indication has caused us to reexamine the validity that TA can be used as a bulk •OH
scavenger, because a flawed bulk •OH scavenger not only misestimates the contribution
of •OH in bulk solution to the overall contaminant degradation, but also misrepresents the
nature of the reaction in the aqueous cavitational systems.
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By using PW ultrasound we evaluated the performance of different •OH scavengers
(i.e., formic acid (FA), carbonic acid (CA), terephthalic acid (TA)/terephthalate (TPA),
potassium iodide (KI), methanesulfonate (MS), benzenesulfonate (BS), and acetic acid
(AA)/acetate) to determine which •OH scavengers react only in bulk solution and which
•OH scavengers interact with cavitation bubbles. The degradation of carbamazepine
(CBZ), a probe compound serving as the reference compound occurred primarily at
bubble-water interface in the study. Based on the pulsed enhancement (PE) of CBZ,
acetic acid/acetate appears to scavenge •OH in bulk solution, and not interact with
cavitation bubbles. Methanesulfonate acts as reaction promoter, increasing rather than
inhibiting the degradation of CBZ. For formic acid, carbonic acid, terephthalic
acid/terephthalate, benzenesulfonate, and iodide, the PE was significantly decreased
compared to in the absence of the scavenger. These scavengers not only quench •OH in
bulk solution but also affect the cavity interface.
Degradation rates by PW
ultrasound were compound dependent with degradation either faster for smaller
compounds or slower for larger compounds than that under CW ultrasound.
Smaller PPCP compounds are able to more readily diffuse to
iv
bubble interfaces and are impacted most by pulsing ultrasound.
In order to test the application of ultrasound to wastewater effluent and gain a
mechanistic understanding of the importance of the octanol-water partition coefficient
(Kow) and diffusivity (Diw) on cavitational systems, we investigated the sonochemical
degradation of a series of pharmaceuticals in DI water and municipal wastewater effluent
under CW and PW ultrasound, respectively. The selection of six target pharmaceuticals
was based on Kow, and Diw. In deionized water, pharmaceuticals with the highest Diw (i.e.,
fluorouracil (5-FU)) and Kow (i.e., lovastatin (LOVS)) exhibited the greatest enhancement
in degradation rates in PW mode as compared to CW mode. This result suggests that a
pharmaceutical with either high diffusivity or hydrophobicity more readily populates the
bubble-water interface during the silent cycle of PW ultrasound. However, in municipal
wastewater effluent, the PE for 5-FU and LOVS disappeared. Our results showed that the
presence of matrix inorganics (i.e., bicarbonate and sulfate anions) did not affect the PE,
indicating that the wastewater effluent organic matter (EfOM) is the cause of the
disappearance of PE. Irradiating 5-FU and LOVS in hydrophobic (HPO), transphilic
(TPI), and hydrophilic (HPI) fractions of EfOM showed that the TPI fraction reduced the
pulse enhancement the most, followed by HPI and HPO fractions. The smaller molecular
weight and high aromaticity of TPI suggests that the TPI fraction is able to diffuse to
cavitation bubble surfaces and interact with pharmaceuticals, resulting in less benefit of
pulsing in the presence of EfOM.
v
Dedication
Dedicated to my doctoral advisor Dr. Linda K. Weavers
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Acknowledgments
I could not have come this far without the help and support of many and I would like
to acknowledge them now.
Dr. Linda Weavers, my advisor, offered me admission to her renowned group,
provided freedom for me to do the research I am interested in, encouraged me to always
scrutinize my work critically, and urged me to never lose sight of the magnitude and
scope of the scientific world.
Dr. Christopher Hadad and Dr. John Lenhart are each impressively knowledgeable
and both provided great detail and information in response to my questions in their
courses. The Spring 2010 courses, Computational Chemistry (Dr. Hadad) and Adsorption
of Pollutants in Environmental Systems (Dr. Lenhart) were extremely rewarding and
developed my interests in computational chemistry and adsorption chemistry. I wish to
thank my final oral examination committee: Dr. Weavers, Dr. Lenhart, Dr. Walker, and
Dr. Chin.
I can not accomplish my computational chemistry study without the encouragement
and support from Dr. Richard Spinney in the Department of Chemistry. He is a skillful
computational chemist and also responsible and patient teacher as well.
I thank Cindy Crawford and Nancy Kaser in our department. Their offices were the
first place I went when I needed a pep talk or suggestions.
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A special thank is given to David Dias-Rivera, an undergraduate who conducted
ultrasound experiments with me from 2009 to 2011. His assistance made this dissertation
possible. In addition, my spoken English improved significantly from our collaboration.
I benefited a great deal from collaboration with Zongsu Wei, Matthew Noerpel, Hoi
Ling (Calvin) Luk, Dr. Ziqi He, Dr. Chin-Min Cheng, and Dr. Meiqiang Cai. They were
sources of intelligence, manpower, and fun. Zongsu and I spent a lot of time in the lab
doing either my pharmaceutical research or his novel horn reactor research, from which I
gained an understanding of the difficulty to scale up a bench-scale technology. Matt’s
name has been mentioned in the acknowledgement part of all of my manuscripts because
he edited every manuscript I wrote during my Ph.D. study. Calvin is from the Department
of Chemistry, although I only got a chance to know him in my last year at OSU, he is
always willing to help and contributed a lot to my computational chemistry paper. Drs.
He and Cheng are like my part-time advisors. Dr. He gave me many valuable suggestions
on my work and Dr. Cheng taught me how to edit my storyline logically. They always
have an ear to the research and non-research problems I run into and provide abundant
solutions and suggestions.
My labmates and friends made my days at OSU enjoyable: Chenyi Yuan, Mengling
Stuckman, Qing Ye, Yuan Gao, Paola Bottega, Lauren Corrigan, Rachael Pasini, Michael
Brooker, Brittnee Halpin, Shuai Liu, Yen-Ling Liu, Maggie Pee, Kate Ziegelgruber,
Maryam Ansari, and Dawn Deojay. We spent many group meetings and coffee breaks at
the Knowlton together.
viii
I want to thank my former advisors, Dr. Qingru Zeng in Hunan Agricultural
University and Dr. Chunxia Wang at the Chinese Academy of Sciences. They are the
people who always provide valuable suggestions even when I am very far away.
Last, but not least, I would like to thank my family for their constant support from
the beginning to the present. Que, my wife, makes my life colorful and meaningful. I
thank my parents for all of their hard work over the years.
ix
Vita
1999-2003 Bachelor of Environmental Engineering
Hunan Agricultural University, China
2003-2006 Master of Science in Environmental Science
Chinese Academy of Sciences, China
2006-2008 Master of Science in Environmental Science
The Ohio State University
2008-present Graduate Research Associate
The Ohio State University
Fields of study
Major Field: Civil Engineering Graduate Program
x
Table of Contents
Abstract .......................................................................................................................... ii
Dedication ...................................................................................................................... v
Acknowledgments......................................................................................................... vi
Vita ................................................................................................................................ ix
Table of Contents ........................................................................................................... x
List of Tables ............................................................................................................... xvi
List of Figures ............................................................................................................ xvii
Chapters 1: Introduction and Background
1.1 Pharmaceuticals and personal care products contamination in water bodies .......... 1
1.2 Conventional treatment technologies ....................................................................... 2
1.3 Sonochemistry as an AOP technology ..................................................................... 3
1.4 Sonochemical degradation of organic contaminants in water ................................. 6
1.4.1 Ultrasonic operating conditions ......................................................................... 7
1.4.1.1 Ultrasonic intensity ..................................................................................... 7
1.4.1.2 Ultrasonic frequency ................................................................................. 10
1.4.1.3 Ultrasonic mode ........................................................................................ 13
1.4.2 Physicochemical properties of organic contaminant ...................................... 15
xi
1.4.2.1 Surface excess ........................................................................................... 15
1.4.2.2 Octanol water partition coefficient ........................................................... 17
1.4.2.3 Henry’s law constant ................................................................................. 19
1.4.2.4 Diffusivity ................................................................................................. 21
1.4.2.5 Relative importance of each physicochemical property ........................... 24
1.4.3 Solution chemistry .......................................................................................... 24
1.4.3.1 pH .............................................................................................................. 25
1.4.3.2 Organic and inorganic matrices ................................................................ 26
1.5 Importance of scavenger study in sonochemistry .................................................. 30
1.6 Motivation of this study ......................................................................................... 31
1.7 Organization of the dissertation ............................................................................. 33
Figures.......................................................................................................................... 35
References .................................................................................................................... 37
Chapter 2: Sonochemical Degradation of Ciprofloxacin and Ibuprofen in the
Presence of Matrix Organic Compounds
2.1 Abstract .................................................................................................................. 53
2.2 Introduction ............................................................................................................ 54
2.3 Materials and methods ........................................................................................... 56
2.4 Results and discussion ........................................................................................... 58
2.4.1 Degradation of ciprofloxacin and ibuprofen .................................................... 58
2.4.2 Effects of matrix organics on degradation ....................................................... 62
2.4.2.1 Terephthalic acid ..................................................................................... 62
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2.4.2.2 Suwannee River fulvic acid: ................................................................... 67
2.5 Conclusions ............................................................................................................ 72
Tables and figures ........................................................................................................ 74
References .................................................................................................................... 82
Chapter 3: Using Pulsed Wave Ultrasound to Evaluate the Suitability of
Common Hydroxyl Radical Scavengers in Sonochemical Systems
3.1 Abstract .................................................................................................................. 91
3.2 Introduction ............................................................................................................ 92
3.3 Experimental .......................................................................................................... 94
3.4 Results and discussion ........................................................................................... 96
3.4.1 Degradation of CBZ ......................................................................................... 96
3.4.2 PE of CBZ in the presence of the •OH scavengers .......................................... 97
3.4.2.1 Formic acid ................................................................................................ 98
3.4.2.2 Carbonic acid ........................................................................................... 100
3.4.2.3 Terephthalic acid/terephthalate ................................................................ 100
3.4.2.4 Potassium iodide ...................................................................................... 101
3.4.2.5 Benzenesulfonate ..................................................................................... 103
3.4.2.6 Methanesulfonate ..................................................................................... 103
3.4.2.7 Acetic acid ................................................................................................ 104
3.4.3 Robustness of AA/acetate as bulk solution •OH scavenger ........................... 107
3.5 Conclusions .......................................................................................................... 109
Figures........................................................................................................................ 111
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References .................................................................................................................. 117
Chapter 4: Degradation of Pharmaceuticals and Personal Care Products (PPCPs)
Using Pulsed Wave Ultrasound in Aqueous Solution
4.1 Abstract ................................................................................................................ 124
4.2 Introduction .......................................................................................................... 125
4.3 Experimental Section ........................................................................................... 127
4.4 Results and Discussion ........................................................................................ 129
4.4.1 Sonochemical degradation kinetics................................................................ 130
4.4.2 Pulse enhancement ......................................................................................... 131
4.4.3 PPCPs degradation in bulk solution ............................................................... 136
4.5 Conclusion ........................................................................................................... 138
Tables and figures ...................................................................................................... 140
References .................................................................................................................. 146
Supporting information .............................................................................................. 151
Chapter 5: Probing the Utility of Pulsed Wave Ultrasound for the Degradation
of Pharmaceuticals in Municipal Wastewater
5.1 Abstract ................................................................................................................ 154
5.2 Introduction .......................................................................................................... 155
5.3 Experimental Methods ......................................................................................... 158
5.3.1 Materials ......................................................................................................... 158
5.3.2 Wastewater effluent collection ........................................................................ 158
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5.3.3 Wastewater effluent organic matter isolation .................................................. 159
5.3.4 Wastewater effluent organic matter characterization ...................................... 159
5.3.5 Sonochemical experiments ............................................................................. 160
5.3.6 Chemical analysis ........................................................................................... 161
5.3.7 Computational modeling ................................................................................. 161
5.4 Results and discussion ......................................................................................... 162
5.4.1 Sonochemical degradation in DI water .......................................................... 162
5.4.2 Sonochemical degradation in wastewater effluent......................................... 165
5.4.3 Effect of inorganic and organic matrix on PE of pharmaceuticals ................ 166
5.4.3.1 Effect of inorganics on PE of pharmaceuticals ........................................ 167
5.4.3.2 Effect of organics on PE of pharmaceuticals ........................................... 169
5.5 Environmental application ................................................................................... 172
Figures........................................................................................................................ 174
References .................................................................................................................. 179
Supporting information .............................................................................................. 189
Chapter 6: Conclusions and Future Work
6.1 Conclusions .......................................................................................................... 196
6.2 Future work .......................................................................................................... 198
References .................................................................................................................. 202
Appendix A: Thermodynamic and Kinetic Study of Ibuprofen with Hydroxyl
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Radical: a Density Functional Theory Approach
A.1 abstract ................................................................................................................ 222
A.2 Introduction ......................................................................................................... 223
A.3 Methodology ....................................................................................................... 226
A.4 Result and discussion .......................................................................................... 229
A.4.1 Thermodynamics ........................................................................................... 229
A.4.1.1 •OH Addition ........................................................................................... 230
A.4.1.2 H Abstraction........................................................................................... 231
A.4.1.3 Reaction at the Carboxylic acid .............................................................. 232
A.4.2 Comparison to experiment ............................................................................. 232
A.4.3 Kinetics........................................................................................................... 233
A.5 Conclusion .......................................................................................................... 235
Tables and figures ...................................................................................................... 237
References .................................................................................................................. 244
xvi
List of Tables
Table 2.1: Physicochemical properties of CIPRO and IBU ........................................ 74
Table 2.2: Degradation kinetics of CIPRO and IBU under various conditions .......... 75
Table 2.3: CIPRO and IBU association with Suwannee River fulvic acid ................. 76
Table 4.1: Selected physicochemical properties of CBZ, IBU, ATP, SFT, CIPRO,
PG, and DP at 25°C. ................................................................................................... 140
Table 5.1: Selected physicochemical properties of 5-FU, IBU, CLND, ESTO,
NIFE, and LOVS at 25°C. ......................................................................................... 190
Table 5.2: Summary of main wastewater effluent characteristics ............................. 191
Table 5.3: Summary of main characteristics for different fractions of EfOM .......... 192
Table A.1: Physiochemical properties of ibuprofen .................................................. 237
Table A.2: R
ΔG andR
ΔH (kcal/mol) for the products involved in the reactions
of ibuprofen with •OH in aqueous solution ............................................................... 238
Table A.3: ΔG‡ (kcal/mol) and imaginary frequency for the transition state
species involved in the reactions of ibuprofen with •OH in aqueous solution. ......... 239
xvii
List of Figures
Figure 1.1: Spatial distribution of temperature and different species in the
three reaction zones in cavitation bubble ..................................................................... 35
Figure 1.2: Schematic diagram of continuous wave (CW) and pulsed wave
(PW) ultrasound ........................................................................................................... 36
Figure 2.1: Sonolysis of CIPRO and IBU at 620 kHz (pH 8.5, 20 °C, and
sonication power density at 400 W/L) ......................................................................... 77
Figure 2.2: Sonochemical degradation rates at different concentrations of IBU
(a) and CIPRO (b) with 2 mM terephthalic acid (TA) (pH 8.5, 20 °C and
sonication power density at 400 W/L) ......................................................................... 78
Figure 2.3: Predicted and observed initial degradation rates with 2 mM
terephthalic acid (TA) at the frequency of 20 kHz (a) and 620 kHz (b) (pH 8.5,
20 °C and sonication power density at 400 W/L) ........................................................ 79
Figure 2.4: The sonochemical degradation rates of CIPRO and IBU with
different concentrations of SRFA at 20 kHz (a) and 620 kHz (b) (pH 8.5, 20 °C,
and sonication power density at 400 W/L). ................................................................. 80
Figure 2.5: Predicted and observed initial degradation rates of CIPRO and IBU
with SRFA at 20 kHz (a. 1 M and b. 10 M) and 620 kHz (c. 1 M and d. 10
M) (pH 8.5, 20 °C, and sonication power density at 400 W/L). ................................ 81
xviii
Figure 3.1: Initial degradation rates of 10 M CBZ in the absence and presence
of 1 mM •OH scavenger candidates under PW and CW ultrasound at pH 3.5 ......... 111
Figure 3.2: Inhibitory effects of each bulk •OH scavenger candidate on sonolysis
of 10 M CBZ under CW and PW ultrasound at pH 3.5 ........................................... 112
Figure 3.3: PE of 10 M CBZ in the absence and presence of the various bulk
•OH scavenger candidates at pH 3.5 (TPA was tested at pH 7.4). ............................. 113
Figure 3.4: Gibbs surface tension ( ) and surface excess ( ) for acetic acid (AA)
and benzenesulfonate (BS) as a function of concentration at room temperature. ..... 114
Figure 3.5: Observed and predicted initial degradation rates of 10 M CBZ at
pH 3.5 under CW and PW ultrasound as a function of acetic acid concentrations
ranging from 0.5 mM to 100 mM. ............................................................................. 115
Figure 3.6: Pulse enhancement of 10 M CBZ in the presence of 1 mM bulk
phase •OH scavenger AA/acetate as a function of pH. .............................................. 116
Figure 4.1: Initial degradation rates of pharmaceuticals (CBZ, SFT, IBU, ATP,
and CIPRO) and personal care products (PG and DP) at initial concentration of
10 µM and pH 3.5 sonicated under CW and PW ultrasound. .................................... 141
Figure 4.2: Pulsed enhancements of 10 µM individual PPCP degradation by
PW ultrasound with a pulse length of 100 ms and pulse interval of 100 ms at
pH 3.5. ........................................................................................................................ 142
Figure 4.3: PE was plotted against the molar volumes (a) and molecular weight
(b) of the PPCPs. ........................................................................................................ 143
Figure 4.4: Initial degradation rates of 10 µM CBZ, IBU, ATP, SFT, CIPRO, PG,
and DP in the absence and presence of 1 mM •OH scavenger AA under CW
xix
ultrasound at pH 3.5. .................................................................................................. 144
Figure 4.5: The portion of degradation that occurred in the bulk solution under
CW ultrasound is plotted against molar volume (a) and molecular weight (b)
of each PPCP molecule. ............................................................................................. 145
Figure 4.6: Degradation rates of the PPCPs under CW and PW ultrasound were
plotted against molecular weight (a), molar volume (b), log vapor pressure (c),
log Kow (d), log Henry’s law constant (e) and log •OH rate constant (f).. ................. 152
Figure 4.7: PE of the PPCPs were plotted against log vapor pressure (a),
log Kow (b), log Henry’s law constant (c) and log •OH rate constant (d). ................. 153
Figure 5.1: The initial sonochemical degradation rates of 5-FU, IBU, CLND,
ESTO, NIFE, and LOVS in DI water (a) and wastewater effluent (b) (pH 7.7,
20 °C, [pharmaceutical]0 = 10 μM, and sonication power density at 45 W/L) .......... 174
Figure 5.2: The PE of the six pharmaceuticals in DI water and wastewater
effluent (a). PE in DI water was plotted as a function of molar volume (b)
(pH 7.7, 20 °C, [pharmaceutical]0 = 10 μM, and sonication power density
at 45 W/L). ................................................................................................................. 175
Figure 5.3: The PE of the six pharmaceuticals in DI water (pH 7.7,
20 °C, [pharmaceutical]0 = 10 μM, and sonication power density at 45 W/L ).
In DI water, the concentration of bicarbonate is 2.5 ×10-4
M, assuming water
is equilibrated with the atmosphere ........................................................................... 176
Figure 5.4: The initial sonochemical degradation rates of 5-FU and LOVS in
DI water and different fractions of effluent organic matter (pH 7.7, 20 °C,
[pharmaceutical]0 = 10 μM, [DOC]= 3 mgC/L, and sonication power density
xx
at 45 W/L). ................................................................................................................. 176
Figure 5.5: The PE of 5-FU and LOVS in different fractions of effluent organic
matter (pH 7.7, 20 °C, [pharmaceutical]0 = 10 μM, [DOC]= 3 mgC/L, and
sonication power density at 45 W/L). ........................................................................ 178
Figure 5.6: The molar weight of 5-FU, IBU, CLND, ESTO, NIFE, and
LOVS was plotted against their apparent log Kow at pH 7.7.. ................................... 193
Figure 5.7: Degradation rates of the pharmaceuticals under CW and PW
ultrasound in DI water were plotted against log Henry’s law constant (a),
apparent log Kow (b), and molecular weight (c) and molar volume (d) ..................... 194
Figure 5.8: The molar volume of 5-FU, IBU, CLND, ESTO, NIFE, and
LOVS was plotted against their molecular weight. (Molar volume was
calculated at Hartree Fock (HF) level of theory with 6-31+G* basis set and
PCM solvation method) ............................................................................................. 195
Figure A.1: Optimized geometry of ibuprofen at B3LYP/6-31+G** level of
theory with IEFPCM water model. The “syn” face is on the bottom of this
diagram and the “anti” face the top............................................................................ 240
Figure A.2: Possible pathways for the reactions of ibuprofen with •OH based on
(a) H abstraction, (b) •OH addition, and (c) nucleophilic attack. .............................. 241
Figure A.3: Profile of the potential energy surface (a. •OH addition pathway;
b. H abstraction and nucleophilic attack pathways) for the reaction between
ibuprofen and •OH at the B3LYP/6-31+G** level of theory. .................................... 242
Figure A.4: Product of “syn” addition to C4 with hydrogen bonding. ..................... 243
1
Chapter 1: Introduction and Background
1.1 Pharmaceuticals and personal care products contamination in water bodies
Pharmaceuticals and personal care products (PPCPs) are defined as “any products
used by individuals for personal health or cosmetic reasons or used by agribusiness to
enhance growth or health of livestock” (EPA, 2010). These compounds have greatly
improved our quality of living. In spite of this, PPCPs have received a significant amount
of attention as potential environmental contaminants. Although the information of exact
amounts of PPCPs produced and consumed is hard to obtain, the sale of pharmaceuticals
was as high as $290.1 billion in 2007 (Khetan and Collins, 2007). Depending on the
medication, up to 90% of these products pass through a human body unchanged
(Daughton and Ternes, 1999), resulting in a substantial amount of anthropogenic
compounds being directly or indirectly released into the aquatic environment.
Numerous studies investigated the occurrence of PPCPs in aquatic environments
around the world (Barcelo and Petrovic, 2007; Boyd et al., 2004; Nakada et al., 2007;
Snyder et al., 2003; Xu et al., 2009). The U.S. Geological Survey (USGS) provided the
first overview of the occurrence of PPCPs in water resources across the United States
during 1999-2000 and concluded that these compounds were prevalent at 139 sampling
sites, found in 80% of the streams sampled (Kolpin et al., 2002). In Ohio, eight sampling
sites have been monitored, a median of seven anthropogenic organic contaminants were
2
found at these sites (Kolpin et al., 2002). Miege et al., (2009) reported that acidic drugs,
such as bezabrate, ibuprofen, indometacine, and naproxen, were ubiquitous in German
rivers and streams with concentration in the ng/L range. Both the Lake Erie Millennium
Plan and U.S. Environmental Protection Agency (EPA) and Environment Canada have
reported that PPCPs have been detected in Lake Erie and the other Great Lakes (LAMP,
2006; U.S.EPA, 2007). Peck at el., (2006) reported galaxolide, a synthetic musk, in Lake
Erie sediment cores from a site near Cleveland, OH. They found a strong correlation
between the concentration of galaxolide and the consumption of fragrances in the United
States.
Toxicological studies have explicitly shown that exposure of fish and other aquatic
organisms to PPCPs cause adverse reproductive effects, such as reduced viability of eggs,
endocrine disruption, and changes in sperm density (Crane et al., 2006; Daughton and
Ternes, 1999; Gu et al., 2002; Porte et al., 2009). A full understanding of the hazards of
PPCPs, especially at environmental concentrations over long exposure times, and
exposure to mixtures of PPCPs, to human health is limited, causing relatively incomplete
environmental regulations and laws. However, as knowledge of risk of exposure to
PPCPs via water, regulations are likely to change. Therefore, technologies to treat these
contaminants need to be developed.
1.2 Conventional treatment technologies
Avoiding consumption of PPCPs is unlikely, because they are medically necessary
and improve our quality of life. Thus, considerable effort has been spent on
understanding removal efficiencies (i.e., fast degradation kinetics) of PPCPs in drinking
3
water resources and wastewater discharges (Batt et al., 2007; Halden and Heidler, 2008;
Kim and Tanaka, 2009; Westerhoff et al., 2005; Xu et al., 2009). Researchers evaluated
the removal efficiency of various PPCPs by different conventional water treatment and
advanced oxidation processes (AOPs) (Westerhoff et al., 2005). Technologies were tested
using bench-scale models and included coagulation, lime softening, powder-activated
carbon (PAC), chlorination, and ozonation. The removal efficiency of PPCPs was
technology dependent, ranging from approximately 20% removal using coagulation to
90% using PAC. In addition, they reported that the removal efficiency of a PPCP
depended on its physicochemical properties, including molecular weight, octanol-water
partition coefficient (Kow), aromatic carbon content, and functional group composition.
It is noted that each technology has both advantages and disadvantages. In the study
above (Westerhoff et al., 2005), after a 4 hour contact time with PAC, about 90%
removal of the initial concentration of PPCPs was achieved. However, a 4 hour contact
time of PAC with PPCPs is too long to be applied to large water treatment facilities
which process millions of gallons of water a day. Another example is chlorination. The
removal efficiency of PPCPs by chlorination was also 90% (Westerhoff et al., 2005), but
chlorine reacts with natural organic matter (NOM), generating toxic disinfection
byproducts (DBPs). Therefore, researchers are searching for other high efficiency and
green water treatment technologies.
1.3 Sonochemistry as an AOP technology
Ultrasound is one example of an advanced oxidation process (AOP) that transforms
organic contaminants. Ultrasound contains unique advantages compared to other
4
technologies, including no addition of chemicals, ease of use, and short contact time
(Adewuyi, 2005a; b; Hoffmann et al., 1996). Ultrasonic irradiation in water induces
collapsing bubbles, known as cavitation bubbles. When water is exposed to ultrasound,
acoustic pressure waves are produced. The acoustic pressure waves consist of
compression and rarefaction cycles. In the rarefaction cycle of the acoustic pressure wave,
it leads to the formation of bubbles from the gas nuclei, dissolved gas that preexists in
water. In the compression cycle of the acoustic wave, the bubble volume decreases due to
increasing pressure in the surrounding water. Thus, bubbles grow and shrink in response
to acoustic pressure. The motion for the gas bubble in a sonicated solution can be defined
by the Rayleigh-Plesset (RP) equation (Leighton, 1994):
/R}R4μR
2σ/R)(Rp(t)]p{[p
ρ
1R
2
3RR
3η
0g0v
2 (1.1)
where R and R are the first and second order derivatives of the bubble radius in terms
of time; ρ is the density of liquid; pv is the vapor pressure; P is the ambient pressure; pg0
is gas pressure inside the bubble; R and R0 are the instantaneous and equilibrium radii of
the bubble, respectively; η is polytropic coefficient (under isothermal conditions, η is 1;
under adiabatic conditions, η is the heat capacities, γ, of the gas in liquid); σ is the surface
tension coefficient in N/m or J/m2; and μ is the viscosity of the liquid. The first term in
the curly bracket on the right side of reaction indicates the discrepancy of the applied
pressure in the liquid and the vapor pressure in gas pocket. It is the driving force that
accounts for the growth of the bubble. This term is associated with the entire life of the
5
bubble, including shrinking, growing, collapsing, and oscillating. The second, third, and
fourth terms in the curly bracket are the contribution of non condensable gas, the
contribution of surface tension, and the contribution of dynamic viscosity, respectively
(Leighton, 1994).
When the sound intensity is greater than the cavitation threshold, within several
cycles of growing and shrinking, bubbles exponentially and eventually collapse. However,
the RP equation fails to accurately describe the collapse, in which interesting chemistry
happens (Suslick and Flannigan, 2008). The collapse of bubbles causes extremely high
temperatures and pressures (Suslick et al., 1986) within a microenvironment in the liquid,
known as hot spots. The average temperatures within the bubble is estimated as high as
5000 K (Flint and Suslick, 1991), leading to the breakdown of gaseous water molecules
in the bubbles to hydroxyl radicals (•OH).
H2O→ •OH + •H (1.2)
The hot spot theory (Suslick et al., 1986) is frequently used to interpret how a
contaminant is degraded by ultrasound. The schematic diagram of hot spot is illustrated
in Figure 1.1. In the center of collapsing bubble (i.e., gas region), the temperature and
pressure are about 5000 K and 500 atm (Flint and Suslick, 1991), respectively, so that
water molecules are thermolyzed to •OH and •H. The temperature in the interfacial
region surrounding the hot core is estimated to be 1900 K, and the •OH concentration is
estimated to be up to 4 mM (Gutierrez et al., 1991). The thickness of the interfacial
region is estimated to be 200 nm (Mason et al., 1990). The temperature of bulk region of
6
the cavitation bubble is ambient. •OH forms in the gaseous bubble core and diffuses to
bulk solution to this region (Adewuyi, 2005a). In terms of distribution of concentration of
reactants and products, the concentrations of solutes decrease from bulk solution to
interfacial region of cavitation bubbles; the maximum concentration of products is at the
interfacial region.
In a sonicated solution, organic contaminants undergo degradation by two different
pathways: decomposition by heat in the gas and interfacial regions of the cavitation
bubbles and •OH oxidation in the gas, interfacial, and bulk region of the cavitation
bubbles (Adewuyi, 2001; Hoffmann et al., 1996; Mendez-Arriaga et al., 2008; Riesz et al.,
1985; Weavers et al., 2005). The proportion of each pathway is dependent on the
physicochemical properties of organic contaminants (Adewuyi, 2001; Hoffmann et al.,
1996). For example, the major degradation location for volatile contaminants is in the gas
region; hydrophobic pollutants in the interfacial region; and hydrophilic contaminants in
bulk region.
1.4 Sonochemical degradation of organic contaminants in water
In the past two decades intensive efforts have been exerted to understand what
determines the treatability of ultrasound in removing organic contaminants from water.
However, more and more studies revealed that the sonication system is so complex that
any change in ultrasonic operating conditions (Adewuyi, 2005a; b; Francony and Petrier,
1996; Mendez-Arriaga et al., 2008; Petrier et al., 1996), physicochemical properties of
organic compounds (Colussi et al., 1999; Mizukoshi et al., 1999; Nanzai et al., 2008), and
solution chemistry (Bolong et al., 2009; Brotchie et al., 2009; Jiang et al., 2002) has an
7
influence on the cavitational effects, thus affecting the effectiveness of ultrasound in
wastewater and drinking water treatment. This section of the dissertation will discuss the
effect of ultrasonic operating conditions, physicochemical properties of organic
compounds, and solution chemistry, respectively.
1.4.1 Ultrasonic operating conditions
The ultrasonic operating conditions typically include ultrasonic intensity, ultrasonic
frequency, and ultrasonic mode (i.e. continuous wave and pulsed wave mode).
1.4.1.1 Ultrasonic intensity
The impact of ultrasonic intensity on chemical reactivity in different sonochemical
systems has undergone considerable examination (Hua and Hoffmann, 1997; Kuijpers et
al., 2002; Mendez-Arriaga et al., 2008; Sunartio et al., 2005; Suri et al., 2007). The
ultrasonic intensity imported to the system (I) is positively correlated to the square of the
acoustic amplitude of the ultrasonic wave, PA (eqn. 1.3) (Leighton, 1994):
2c ρ
PI
2
A (1.3)
where c is the speed of sound in water (1482 m/s in water at 20 °C) and ρ is the density of
water (998 g/L at 20 °C). In addition, the ultrasonic intensity also affects the bubble
collapse time, τ, as indicated by eqn. 1.4.
8
)P
P(1)
P
ρ(0.915Rτ
A
vapor1/2
A
max (1.4)
where Rmax is the maximum radius reached by a bubble during expansion; ρ is the density
of the liquid; and Pvapor is the vapor pressure in the bubble. Therefore, with higher
amplitude ultrasonic wave, the cavitation bubbles collapse within a shorter time.
The acoustic intensity has been measured using different chemical dosimeters,
including 5, 5-dimethylpyrroline-N-oxide (DMPO) (Riesz et al., 1985; Sostaric and Riesz,
2001), terephthalic acid (Fang et al., 1996; Villeneuve et al., 2009), potassium iodide
(Entezari and Kruus, 1994; Weissler et al., 1950), and luminol (Price et al., 2010; Tuziuti
et al., 2004). Every chemical used in the dosimetry has advantages and disadvantages.
For example, DMPO has been frequently used in dosimetry. Although Electron spin
resonance (ESR) spectroscopy sensitively traps the spin-adduct signal in cavitational
systems, the ESR spectroscopy is not commonly used due to its high cost and difficulty in
quantification. The terephthalate dosimeter measures •OH by detecting the fluorescence
signal of the •OH-trap adduct. Although a fluorescence spectrophotometer is not
expensive, terephthalic acid has low solubility in an acidic solution (15 mg/L at 25˚C
(Yalkowsky and Banerjee, 1992)). In addition, all these methods depend on •OH reacting
with chemical dosimeters. These dosimeters may not sufficiently differentiate the
locations of •OH reaction, because the •OH-trap adduct forms proportionally based on the
concentration of •OH and the •OH trap in the cavitation bubble, at bubble-water interface
and in bulk solution.
9
The correlation between sonochemical reactivity and ultrasonic intensity has been
studied by previous researchers (Hua and Hoffmann, 1997; Kanthale et al., 2008;
Merouani et al., 2010; Price and Lenz, 1993; Weissler et al., 1950). Merouani et al. (2010)
monitored the formation rate of triiodide in 0.1 M KI aqueous solution at the frequency of
300 kHz and observed that the rate increased as a function of power input, ranging from
45.6 to 85.3 W/L. Weissler et al., (1950) also observed the linear formation of triiodide
under sonication at the frequency of 1 MHz as the power input increased from 0 to 600
W, which is a wider range than that in Merouani et al. study. They both attributed the
strong correlation between sonochemical reactivity (i.e., triiodide formation in this case)
and power input to a greater number of cavitation bubbles and a more complete bubble
implosion.
However, increasing power input does not necessarily cause a faster degradation rate
for all contaminants. Hua et al. examined the dependence of the first-order degradation
rate constant of para-nitrophenol (p-NP) on the power input (Hua and Hoffmann, 1997).
The input power was varied from 0.3 to 1.5 W/cm2 and an optimum rate was observed.
They found the rate constant first increased with an increase of intensity, but it started to
decrease when the power intensity reached at 1.2 W/cm2. A similar trend has been
observed by other studies. Henglein and Gutierrez (1988) degraded polyacrylamide with
1 MHz continuous wave ultrasound in aqueous solution. They observed that the first-
order degradation rate constant reached its maximum at 100 W and gradually decreased
as the power input increased. The decrease in the degradation rate of the target
compounds was attributed to incomplete bubble collapse, a greater extent of coalescence
10
of multi-bubbles, and formation of a bubble shroud at the surface of the transducer,
resulting in less penetration of the acoustic wave (Hua and Hoffmann, 1997).
1.4.1.2 Ultrasonic frequency
Many studies have shown that ultrasonic frequency determines the size of cavitation
bubbles (Beckett and Hua, 2001; Brotchie et al., 2009; Hung and Hoffmann, 1999;
Leighton, 1994; Price et al., 2010). As shown in eqn. 1.5 (Leighton, 1994), the maximum
radius reached by a bubble during expansion, Rmax, is dependent on ultrasonic frequency,
f, the hydrostatic pressure, P0 , and acoustic pressure, PA, and the density of the solution,
ρ. f is inversely proportional to the Rmax, indicating that at higher frequency the size of
the collapsing bubble is smaller than at low frequency.
1/3
0A
0
1/2
A
0Amax))P(P
3P
2(1
ρP
2)P(P
f6
4R (1.5)
In addition, Petrier et al., (1996) investigated the sonochemical degradation of
2,2,6,6-tetramethyl-4-piperidinon with and without argon at 20 kHz and 514 kHz,
respectively. By measuring the formation of triiodide in KI solution under different
conditions, they concluded that, as frequency decreases, large sized cavitation bubble
forms and it collapses more violently, resulting in less •OH to be rejected out of bubbles.
Brotchie et al., (2010) reported that the frequency also affects the lifetime of cavitation
bubbles, which is defined as the time of the cavitation bubbles between their nucleation
and collapse. Cavitation bubbles at a high frequency have more acoustic cycles per unit
11
time with a shorter lifetime than those at low frequencies. For example, at 20 kHz, the
bubble lifetime is 0.26 ms as compared to 0.22 ms at 355 kHz (Brotchie et al., 2010). The
acoustic cycle at 20 kHz is 5 ( 5
1020
1
1026.0
3
3
) as compared 75 ( 75
10355
1
1022.0
3
3
) at 355
kHz. However, the dependence of radius, acoustic cycles, and lifetime of the cavitation
bubble on frequency does not demonstrate the influence of frequency on chemical
reactivity.
Therefore, a number of techniques were employed to elucidate the dependence of
chemical reactivity on frequency (Beckett and Hua, 2001; Brotchie et al., 2009; Kang et
al., 1999; Petrier et al., 1994; Yang et al., 2008). Yang et al., (2008) applied a
terephthalate dosimeter to measure •OH by detecting the fluorescence of the •OH-trap
adduct (i.e., hydroxyterephthalate) at different frequencies. They found that the order of
•OH yield was 354 kHz > 620 kHz > 803 kHz > 206 kHz > 1062 kHz. Beckett and Hua
(2001) monitored the production of H2O2 at four frequencies (205, 358, 618, and 1071
kHz). H2O2 is a product that is generated from the recombination of •OH, as illustrated in
eqn. 1.6
•OH + •OH →H2O2 (1.6)
The order of production of H2O2 was 358 kHz > 205 kHz > 618 kHz > 1071 kHz. Kang et
al., (1999) also investigated the formation of H2O2 at 205 kHz, 358 kHz, 618 kHz, and
1078 kHz and the order of production of H2O2 was 358 kHz > 618 kHz 205 kHz > 1078
kHz. The studies above observed an optimum at 200-600 kHz. The increased •OH yield
12
at high ultrasonic frequency is explained by the fact that cavitation at a high frequency
rejects more •OH, resulting in high •OH yield. But when the ultrasound frequency
increases into megahertz range, high frequency does not produce more •OH, because
rarefaction and compression cycles decrease, resulting in a shorter time for nucleation
process of cavitation bubbles to expand to bubbles. In addition, the collapsing
temperature at higher ultrasound frequencies reduce the •OH yield due to a less adiabatic
collapse (Yang et al., 2008).
In terms of the effects of frequency on contaminant removal, many studies have
investigated the influence of ultrasonic frequency on contaminant degradation (Beckett
and Hua, 2001; Hung and Hoffmann, 1999; Kang et al., 1999; Petrier et al., 1994; Yang
et al., 2008). These studies revealed a similar trend to that of the •OH yield discussed
above. For instance, Hung and Hoffmann (1999) compared the rates of sonolysis of CCl4,
a precursor to refrigerants and a cleaning agent, at different frequencies. The results
showed that the degradation rate increased from 20 to 618 kHz and then decreased
gradually as frequency increased to 1078 kHz. They attributed the high degradation rate
observed at 618 kHz to the long lifetimes and the high surface area to volume ratios of
the cavitation bubbles. Yang et al., (2008) also observed that the initial degradation rates
of octylbenzene sulfonic acid (OBS), a surface active contaminant, as a function of
frequency, were in the order: 620 kHz > 803 kHz = 354 kHz > 1062 kHz > 206 kHz.
They also attributed the higher initial degradation of OBS at 354 and 803 kHz to the
higher extent of adsorption of the surfactant to bubble interfaces due to prolonged bubble
lifetimes and bubble-water interface of cavitation bubbles.
13
In summary, as we discussed above, the extent that frequency affects the removal of
contaminants depends on •OH yield, mass transfer of •OH out of cavitation bubbles, and
accumulation of contaminants on cavitation bubbles. The dependence suggests that, when
ultrasound is applied to water treatment, •OH production rate and physicochemical
properties of target contaminants should be taken into consideration.
1.4.1.3 Ultrasonic mode
The ultrasound is transmitted to the solution in continuous wave (CW) and pulsed
wave (PW) modes. PW ultrasound is widely used as a tool in medical diagnosis due to
less potential biohazard as compared to CW ultrasound. In the PW ultrasound, the
continuous ultrasonic signal train is separated by gaps of no signal (Leighton, 1994).
Figure 1.2 describes the similarity and difference between CW and PW ultrasound in our
study. Both CW and PW ultrasonic signal trains have same amplitude. In order to expose
the solution to the same amount of acoustic energy, a longer total sonication time is
applied under pulsed conditions. The total sonication time is obtained from eqn. 1.7.
R
11
tt ltota
sonication (1.7)
where tsonication is the actual sonication time; ttotoal is the total time for an experimental run;
and R is the signal on/off ratio (Ton:Toff). In our study, Ton=Toff=100 ms was used for all
the experiments.
14
Many studies have shown that at proper settings the degradation kinetics of a
contaminant under pulsed wave (PW) ultrasound is more effective than that under CW
ultrasound (Flynn and Church, 1984; Henglein, 1995; Henglein et al., 1989; Neppolian et
al., 2009; Yang et al., 2007). For instance, with ultrasound 100 ms on and 100 ms off, the
degradation rate of 0.1 mM surface-active contaminant, OBS, was approximately 30 %
faster than that of CW (Yang et al., 2005). Xiao et al., observed that the removal
efficiency of 10 μM carbamazepine (CBZ) under PW ultrasound was greater than CW by
about 6 %. Flynn and Church (1984) investigated the iodine release at Ton= 60 s but
various Toff time intervals, with the acoustic intensity of 30 W/cm2, the maxima in iodine
release was at Toff=6 ms. However, with the acoustic intensity of 20 W/cm2, the maxima
in iodine release was at Toff=60 ms. Surprisingly, with the acoustic intensity of 10 W/cm2,
there is no maxima in iodine release. Neppolian et al., (2009) found a remarkable
enhancement (1.3 fold) of oxidation of arsenic (III) to arsenic (V) under PW ultrasound at
ca. 35 W/L and duty cycle 1:1, as compared to CW ultrasound.
However, the mechanisms for the enhanced sonochemical reactivity under PW
ultrasound remain unclear. The enhanced degradation kinetics in Yang et al. (2005) was
attributed to the accumulation of the surfactant on cavitation bubble surfaces, resulting in
reduced surface tension and lowering the cavitation threshold. In addition, surfactants
inhibit bubble dissolution during the off cycle, resulting in a larger active cavitation
bubble population and increased absorption sites. Deojay et al. (2011) attributed the
enhanced degradation kinetics to the accumulation of OBS. During the time interval
between two successive pulses, surface active OBS molecules accumulate on the surface
of the bubble. In the subsequent pulse more OBS molecules react compared to CW
15
ultrasound. Flynn and Church (1984) attributed the increased formation of iodine in PW
mode to the survival of unstabilized nuclei during the off-time, which survived as stable
cavities from one pulse and caused more violent collapse in subsequent pulse. Neppolian
et al., (2009) attributed the increased sonochemical degradation of arsenic (III) by PW
ultrasound to the “silent” oxidation reactions and the increased number of active
cavitation bubbles. The controversy on the enhancement by PW ultrasound reveals that
there are many uncertainties and fundamental mechanisms need to be answered.
1.4.2 Physicochemical properties of organic solute
The physicochemical properties of an organic contaminant itself impact contaminant
removal efficiency. In a sonicated solution, thermodynamic parameters, such as surface
excess (Γ), octanol-water partition coefficient (Kow), and Henry’s law constant (KH), have
been related to the ability of contaminants to populate cavitation bubbles; kinetic
parameters, such as diffusivity (Diw), control how fast contaminants are able to populate
cavitation bubbles.
1.4.2.1 Surface excess
Surface excess is defined as “the difference between the amount of a component
actually present in the system, and that which would be present (in a reference system) if
the bulk concentration in the adjoining phases were maintained up to a chosen
geometrical dividing surface” (McNaught et al., 2000). The surface excess is calculated
from surface tension (γ) (eqn. 1.8)
16
dln[S]
dγ
RT
1Γ (1.8)
where [S] is the solute concentration, R is the universal gas constant and T is temperature.
The equilibrium surface excess is positively correlated to the degree of accumulation
of organics on the bubble-water interface, where thermolysis occurs. A compound with
higher surface excess accumulates to a larger degree on bubble surfaces, quenching
sonoluminescence (SL), a phenomenon arising from vibrationally excited states of
molecules produced as a result of the high temperatures and pressures at bubble collapse
(Didenko et al., 2000). The quenching effect results from organics partitioning into the
cavitation bubbles and reacting with excited-state molecules, ultimately lowering the
collapse temperature (Ashokkumar and Grieser, 2000). Therefore, the decrease of SL is
considered a sign for enrichment of a substance at the bubble-water interface. A high
degree of quenching indicates a high extent of accumulation of organics on the interfacial
region. Several studies examined the quenching effects of surface active agents, such as
aliphatic alcohols and surfactants, on SL (Ashokkumar et al., 2000; Ashokkumar et al.,
1997). Ashokkumar et al., (1997) measured the SL and surface tension of C1-C5 aliphatic
alcohols and the surfactants sodium dodecyl sulfate, dodecyltrimethylammonium
chloride, and N-dodecyl-N,N-dimethyl-3-ammonio-1-propanesulfonate at a frequency of
515 kHz. They found that the SL signal decreased as surface tension increased,
suggesting a substance with high surface activity tends to accumulate at the interfacial
region. However, using ESR spectroscopy and the spin-trap, 3,5-dibromo-4-
nitrosobenzenesulfonate to trap primary and secondary radicals, Sostaric and Riesz (2001)
17
did not observe a correlation between equilibrium surface excess and the extent of
ultrasonic decomposition of the surfactants, n-alkanesulfonates, n-alkyl sulfates, n-
alkylammoniopropanesulfonates, and poly(oxyethylenes) at 47 kHz. Their results showed
that the surface excess did not sufficiently describe the behavior of the surfactants at the
bubble-water interface. Sonolysis of the surfactants depended on their chemical structure
and their ability to diffuse from the bulk solution to the interfacial region of cavitation
bubbles.
1.4.2.2 Octanol water partition coefficient
Kow is “a measure of the way in which a compound will partition itself between the
octanol and water phases in the two-phase octanol-water system” (Lide, 2005). Octanol is
a typical organic solvent to describe the partitioning behavior of an organic compound
between natural organic phases and water due to its amphiphilic character (favorably
interacting with both bipolar and monopolar solutes). Kow is a physicochemical property
that is a measure of the hydrophobicity of a compound.
Many studies have evaluated the effect of Kow on the sonolytic degradation of
organic contaminants (Emery et al., 2005; Fu et al., 2007; Nanzai et al., 2008; Park et al.,
2011; Wu and Ondruschka, 2006). Nanzai et al., (2008) sonolyzed a variety of
monocyclic aromatic organic compounds, including nitrobenzene, aniline, phenol,
benzoic acid, salicylic acid, 2-chlorophenol, 4-chlorophenol, styrene, chlorobenzene,
toluene, ethylbenzene and n-propylbenzene at the frequency of 200 kHz and power
density of 3333.3 W/L. They found the degradation rate was positively correlated to log
Kow for these compounds and concluded that Kow is the most important factor for
18
sonolysis of aromatic compounds. Park et al., (2011) examined the sonolysis of phenol,
4-chlorophenol, 2,6-dichlorophenol, 2,4,6-trichlorophenol, 2,3,4,6-tetrachlorophenol, and
pentachlorophenol under three different frequencies, 28 kHz, 580 kHz, and 1000 kHz,
respectively. They reported a strong positive relationship between log Kow and first-order
degradation rate, indicating that a hydrophobic compound undergoes faster sonochemical
degradation. The reported correlations in previous studies may be attributed to the fact
that the investigated compounds belong to monocyclic aromatics with similar structure
but with varied properties, thereby simplifying the comparison.
However, the dependence of degradation kinetics on log Kow failed when the target
compounds covered a wide range of physicochemical properties and structures. Wu and
Ondruschka (2006) compared the sonochemical degradation kinetics of diethyl sulfide,
diallyl sulfide, dipropyl sulfide, dibutyl sulfide, diethyl disulfide, and dipropyl disulfide
at 850 kHz. In their study, the hydrophobicity failed to predict the observed rate; diethyl
disulfide did not fit the trend. Although authors did not provide a reason for lack of
correlation, it may be due to the two double bonds at the α position, resulting in the H on
the secondary carbon being easily abstracted. In addition, Fu et al., (2007) investigated
the sonolysis of nine different estrogenic compounds and evaluated the relationship
between estrogen degradation rate and Kow value. However, no correlation whose R2 was
greater than 0.8 was observed between the two variables. They explained the lack of
dependence to the inaccuracy of the documented Kow values.
Although their explanations may be valid in their cases, the sonochemical process
involves thermolysis and •OH oxidation of contaminants and formation of byproducts in
gas, interfacial, and bulk regions of cavitation bubbles. Therefore, in this heterogeneous
19
cavitational system, a single physiochemical property of the parent PPCP may have
limited capability to accurately govern the complex kinetics, especially when the
investigated compounds cover a wide range of physicochemical properties and diversity
of structures. In addition, bubbles in the ultrasound field are subjected to high velocity
oscillations and translations (Leighton, 1994). These phenomena significantly affect the
fluid dynamics in the reactor, resulting in a more complex factor to take into
consideration when it comes to prediction the degradation kinetic using a single
physiochemical property of a PPCP.
1.4.2.3 Henry’s law constant
KH is a measure of the distribution of a compound between the gas and water phase
defined by eqn. 1.9
i
i
HC
pK (1.9)
where pi is the partial pressure of compound and Ci is the concentration of compound in
aqueous solution. KH is in unit of Pa-m3/mol in this equation. Similar to any
thermodynamic properties, KH is dependent on temperature, as indicated by eqn. 1.10.
))T
1
T
1(
R
HΔexp(kk soln
HH
(1.10)
20
where ΔHsoln is the enthalpy of solution;
Hk and Tº are the Henry’s law constant at the
reference temperature and the reference temperature, respectively.
The Henry’s law is applicable to an ideal dilute solution and is relevant in
cavitational systems, since the gas region of cavitation bubbles is the source of chemical
reactivity (i.e., high temperature and high concentration of [•OH]).
Previous studies show that KH influences sonochemical degradation kinetics
(Ayyildiz et al., 2007; Colussi et al., 1999; De Visscher, 2003; Nanzai et al., 2008; Petrier
et al., 1998; Petrier et al., 2010). Colussi et al., (1999) reported a dependence of first-
order rate constant on Henry’s law constants of various chlorinated hydrocarbons such as
CCl4, CHCl3, C2Cl6, or CH2Cl2 at the frequencies of 205, 358, 618, and 1078 kHz. A
compound exhibiting a greater tendency to enter the gas region (i.e., a high KH) degrades
faster than a compound with a low KH in a sonicated solution.
However, degradation does not always correlate with Henry’s law constant. Nanzai
et al., (2008) investigated the relationship between the KH of twelve aromatic organics
and the first-order rate constant of degradation at 200 kHz CW ultrasound. The first-order
rate constant remained unchanged as KH increased from 10-9
to 10-4
atm m3 mol
-1, and
then became positively correlated to KH when it increased from 10-4
to 10-1
atm m3 mol
-1,
suggesting there is a threshold effect of KH on the sonolysis of organics. De Visscher et
al., (2003) evaluated the relationship between the degradation rate of ethylbenzene,
benzene, o-chlorotoluene, and styrene and their Henry’s law constants. No apparent
relationship between degradation rate and Henry’s law constant was observed. They
attributed the independence to the fact that diffusion limitations governed the partitioning
of the aromatic compounds in water to the cavitation bubbles.
21
Similar to Kow, degradation rates are correlated to KH in some cases but not others.
The effect of KH on sonolysis may be more pronounced when the investigated
compounds have similar structures or physicochemical properties, thereby simplifying
the comparison.
1.4.2.4 Diffusivity
The degradation of organic contaminants in a sonicated solution may be controlled
not only by thermodynamics as discussed above, but also kinetics. In aqueous solution,
Fick’s first law, defined in eqn.1.11, dictates that the ability of a substance to travel
through a mixture by Brownian motion (Fick, 1995).
dz
dCDJ i
iwi (1.11)
where Ji is the mass flux of chemical i in direction of concentration gradient with the unit
of mg/m2·s; Diw is the diffusion coefficient of chemical i in water with the unit of m
2/s; Ci
is the concentration of i with the unit of mg/L; and z is distance in direction of
concentration gradient with the unit of m (i.e., thickness of boundary layer).
It is clear that the mass flux depends on the concentration gradient and diffusivity
coefficient. A greater concentration gradient before and after the boundary layer induces
a faster mass flux. Furthermore, the diffusivity coefficient is a function of molar volume
of an organic compound in water (Hayduk and Laudie, 1974) (eqn. 1.12).
22
0.589
i
1.14
w
5
iw
Vμ
1013.26D (1.12)
where Vi is the molar volume in the unit of mL/mol; and w is the viscosity of water. In
some cases, Vi is approximated by molecular weight, since the database for Vi is not large
resulting in Vi being difficult to obtain. The viscosity of water is constant at a certain
temperature; eqn. 1.12 indicates that a smaller molecule has a higher degree of mobility.
There are several studies focusing on the impact of diffusivity of an organic
contaminant on sonolysis (Ayyildiz et al., 2007; De Visscher, 2003; DeVisscher et al.,
1996; Dewulf et al., 2001; Fu et al., 2007; Yang et al., 2005). Fu et al., (2007) reported
the rate of sonolysis of the estrogen family of compounds (e.g. 17β-estradiol, 17α-
estradiol, ethinyl-estradiol, estrone, equilin, gestodene, levonorgestrel, and norgestrel) at
the frequency of 20 kHz. Their results showed a dependence of first-order degradation
rate on molecular weight. The estrogenic compound with smaller molecular weight
degraded faster than those with larger molecular weight. Fu et al., (2007) attributed this
to the small size of the molecule allowing them to diffuse to the gas or interfacial region
of cavitation bubbles, where the reaction occurs. Yang et al., (2005) applied PW
ultrasound to degrade surfactants OBS and sodium dodecylbenzenesulfonate (DBS) at the
frequency of 354 kHz with a pulsed time on and off ratio of 1:1. The pulsed enhancement
(PE), which is a measure of the effect of pulsing on the rate of sonochemical degradation
as indicated by eqn. 1.13, for OBS (31%) was observed to be higher than that of DBS
(7%). They attributed the difference in PE to the different diffusion coefficients:
23
％ 100
dt
Cd
dt
Cd
dt
Cd
(%) PE
CW
CWPW (1.13)
where CW
dt
Cd and
PWdt
Cdare initial degradation rate of the target compounds under
CW and PW ultrasound, respectively. The diffusion coefficient ratio between OBS and
DBS, DOBS/DDBS, was calculated to be 1.17. Thus, OBS diffuses to the bubble surface
faster than that of DBS, resulting in more accumulation of surfactant at the bubble-water
interface.
Diffusivity plays an important role in cavitational systems. First, previous studies
showed that the lifetime of bubble is not long enough for a solute to reach equilibrium
partitioning (Price et al., 2004; Sostaric and Riesz, 2002; 2001; Yang et al., 2006). Thus,
the diffusivity, rather than the equilibrium properties, of a solute controls its ability to
accumulate on the bubble-water interface, which is the source of reactivity. Second, the
concept of PE is based on a comparative method (eqn.1.13). The observed degradation
rate of the target compounds (i.e., PPCPs in this study) by sonolysis involves both high
temperature decomposition and •OH oxidation in the vicinity of cavitation bubbles and
the bulk solution, as indicated by eqn. 1.14.
bubbleinterfacebulkobs dt
dC
dt
dC
dt
dC
dt
dC
(1.14)
where dt
dC
represents the degradation rate of a target compound, occurring in bulk
solution (bulk), occurring at the interface (interface), and occurring in the gas region of
24
cavitation bubble (bubble), respectively. The contribution from bulkdt
dC to
obsdt
dC under
CW ultrasound is similar to that under PW ultrasound (Deojay et al., 2011; Yang et al.,
2005), thus there is no contribution of bulk reaction to PE. In addition, the PPCPs
investigated are not likely to degrade inside the cavitation bubble due to their low
Henry’s law constants, suggesting that a negligible contribution of bubble reaction to PE.
Thus, the remaining component,interfacedt
dC, the reaction at bubble-water interfaces, results
in the controlling mechanism, diffusion, becoming more pronounced in our system.
1.4.2.5 Relative importance of each physicochemical property
The target contaminants in this study are PPCPs. Most PPCPs in aquatic systems are
non-volatile, non-surface active, and present at trace levels (nM to µM) (Daughton and
Ternes, 1999; Ellis, 2006; Kolpin et al., 2002). Thus, the effects of Henry’s law constant
and surface excess on sonochemical degradation kinetics are less influential as compared
to the effect of hydrophobicity and diffusivity.
1.4.3 Solution chemistry
The studies of the influence of solution chemistry on sonolysis of contaminants often
focus on the pH and organic and inorganic matrices in solution (Chakinala et al., 2007;
Ince et al., 2009; Jiang et al., 2002; Tauber et al., 2000; Uddin and Hayashi, 2009). These
studies are of importance for understanding the application of ultrasound in wastewater
and drinking water treatment, since drinking water source waters and wastewaters usually
exhibit different pH values and concentrations of organic and inorganic matrices,
25
depending on geological conditions or/and the sources of waters to be treated (i.e.
industrial or domestic source).
1.4.3.1 pH
Previous studies have shown that pH is an important parameter in ultrasound
treatment technology (Chakinala et al., 2007; Ince et al., 2009; Tauber et al., 2000; Uddin
and Hayashi, 2009). It is generally accepted that the effect of solution pH determines the
protonation of organic acids/bases, thus altering the degradation kinetics. Mendez-
Arriaga et al., (2008) stated that ultrasound may be used to eliminate ibuprofen (IBU) in
aqueous solutions and the initial degradation rate of IBU at pH 3 was significantly faster
than those at pH 5 and 11. At pH values lower than its pKa, 4.9, the protonation of the
carboxylic group on IBU increases its hydrophobicity, resulting in a higher portion of
IBU being degraded at the interface of the cavitation bubbles, leading to a faster
degradation rate. Jiang et al., (2002) studied the sonolysis of 4-nitrophenol (pKa =7.08)
and aniline (pKa= 4.6) with pH values ranging from 2 to 9. They observed that the
degradation rate of 4-nitrophenol decreased with an increase of pH, but the degradation
rate of aniline increased with an increase of pH. They attributed the faster degradation
rates of neutral 4-nitrophenol and aniline over their ionic forms to the protonation of
hydroxyl and deprotonation of amino groups, resulting in hydrophobic neutral species
that more easily diffuse to and accumulate at bubble-water interfaces.
Several studies explain the change of solution pH to the alteration in electrical
change on bubble surfaces (Cheng et al., 2008; De Bel et al., 2009; Watmough et al.,
1992), ultimately resulting in increased degradation kinetics. De Bel et al. examined the
26
application of ultrasound for the degradation of ciproflaxin (CIPRO) in water and showed
that pH affected its degradation rate as well (2009). The degradation rate at pH 3 was
almost four times faster than that at pH 7 and 10. They suggested that the electrostatic
attractive force between the CIPRO molecule and the negatively charged liquid-bubble
interface, which increased with decreasing pH, enhanced the CIPRO degradation rate. As
pH decreased, the bubble surfaces became more and more negatively charged. In addition,
Watmough et al., (1992) used a 1 kV potential electrode to record the dye (i.e., methylene
blue and sky blue dye) deposition on paper in a sonicated solution. They claimed that the
ultrasound induced gas bubbles carry a negative electric charge with a field charge of
about 7×105 V/m.
Other researchers have come up with different explanations (Beattie et al., 2009;
Cheng et al., 2008; Winter et al., 2009). Cheng et al., (2008) monitored degradation of
perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) at different pH values
and attributed the fast kinetics at low pH to interactions of protons with the bubble-water
interface. They suggested that the bubble-water interface becomes increasingly positively
charged as pH decreases below 4. Winter et al., (2009) summarized the surface-selective
spectroscopy (photoelectron spectroscopy) results and molecular dynamics simulations
and concluded that the air-water interface is more positively charged than the bulk due to
the presence of hydronium ion in acidic solution. The conflicting studies all confirmed
that the surface charge on bubbles changes as pH decreases, but it is controversial
whether the charge becomes more positive or negative.
1.4.3.2 Organic and inorganic matrices
27
When AOPs are investigated in water treatment, the water matrix containing organic
and inorganic components significantly alters the degradation of the target contaminants
(Cheng et al., 2010; 2008; Lu et al., 2002; Sanchez-Prado et al., 2008; Seymour and
Gupta, 1997; Taylor et al., 1999; Westerhoff et al., 1999). The mechanisms by which the
matrices of organic and inorganic components affect the sonochemical degradation of
contaminants are complex and not yet completely understood.
Several studies investigated the presence of natural organic matter (NOM), such as
Suwannee River fulvic and humic acid (SRFA/SRHA) on sonochemical degradation of
target contaminants (Cheng et al., 2008; Laughrey et al., 2001; Lu and Weavers, 2002;
Taylor et al., 1999). Some studies reported that NOM reduces the sonochemical
degradation of target contaminants, such as 4-chlorobiphenyl (4-CB) (Lu and Weavers,
2002), and polycyclic aromatic hydrocarbons (PAHs) (Taylor et al., 1999). The reduction
in degradation was attributed to two factors: (1) NOM competes for •OH, potentially
hindering target contaminant degradation, and (2) SRFA alters cavitation bubbles via
surface tension changes, reducing the bubble-water interfacial temperatures during
transient cavitation, ultimately decreasing observed sonochemical degradation rates of
PAHs.
However, other studies claimed that the presence of NOM did not affect the
sonochemical degradation of target contaminants, such as methyl tert-butyl ether (MTBE)
(Kang et al., 1999), PFOS, and PFOA (Cheng et al., 2008). Kang et al., (1999)
investigated the effect of NOM (i.e., made from Fluka AG) on MTBE at 358 kHz and did
not observed any change in degradation rate until the NOM concentration increased to 4
mg/L. The minimal interference for MTBE may be due to first, the high volatility of
28
MTBE, resulting in MTBE reacting in the cavitation bubbles; second, the low mobility of
NOM in sonicated solution, thus less interference by NOM. Cheng et al., evaluated the
effect of the presence of NOM (e.g., humic acid and fulvic acid) on sonolysis of PFOS
and PFOA at 612 kHz. With the 15 mg/L NOM, the degradation kinetics of both PFOS
and PFOA were not altered by the presence of NOM. They attributed the unchanged
kinetics to the nonvolatility of humic and fulvic acids and a greater surface activity of
PFOS/PFOA than NOM.
For the inorganic matrix, studies have been mainly focused on the impact of ions,
especially anions, on sonolysis (Cheng et al., 2010; Minero et al., 2008; Petrier et al.,
2010; Seymour and Gupta, 1997; Suri et al., 2010). However, there is no consensus of
opinions on how anions affect the extent of sonochemical degradation, since a wide range
of results have been reported. It is generally accepted that anions alter the sonochemical
degradation of contaminants through several mechanisms: competition for •OH,
alterations in water structure and accumulation of organic compounds at the bubble-water
interface through salting-out effects. Suri et al., (2010) observed a wide range of anionic
effects when investigating the degradation of seven different kinds of estrogenic
compounds using a 20 kHz sonication unit operating at 2 kW. With the increase of HCO3-
from 0 to 10 mM, lack of effect was observed on the degradation of 17α-estradiol, 17β-
estradiol, and 17α-ethinyl estradiol. However, equilin and 17α-dihydroequilin exhibited a
decrease in degradation rates. Surprisingly, the presence of HCO3- enhanced, rather
reduced, the degradation of estradiol and estrone. In addition, Petrier et al., (2010)
investigated the sonochemical degradation of 22 nM bisphenol-A in the presence of
bicarbonate anion at 300 kHz and observed degradation rate was increased by a factor 3.2.
29
They attributed the enhanced degradation kinetics to oxidation of bisphenol-A by
carbonate radical, a product from the reaction between •OH and bicarbonate.
3HCO + •OH •
3HCO + OH
- (1.15)
•3
HCO • 2
3CO + H
+ (1.16)
The carbonate radical, • 2
3CO , with a standard electron potential 1.78 V at pH 7, may
diffuse to bulk solution, resulting in oxidation of bisphenol-A. Cheng et al., (2010)
investigated the effect of the specific anion on sonolysis of PFOS and PFOA at the
frequency of 612 kHz. They observed that the role anions played on the degradation
kinetics of PFOS and PFOA followed the Hofmeister series: ClO4- > NO3
- > Cl
- > SO4
2-.
They speculated that the changing water structure at the bubble-water interface due to the
ions may alter water vapor transported into the bubble, resulting in a decreased collapse
temperature. Seymour and Gupta (1997) observed a 6-fold increase in degradation rate
for chlorobenzene, 7-fold for p-ethylphenol, and 3-fold for phenol in the presence of 1.38
M NaCl at 20 kHz. They explained that since the major portion of degradation of
chlorobenzene, p-ethylphenol, and phenol occurs in the bubble-water interface, the
addition of salt drove the organic compounds to the bubble surface, which is source of
ultrasonic reactivity, hence resulting in a faster degradation rate.
The application of ultrasound in removal of contaminants in water is a complex
process. Any change in a single parameter results in an alteration in removal efficiency.
In the studies discussed above, although an explanation has been successfully applied to
30
each individual case, it often fails to explain other similar experimental observations. The
difficulty in understanding the characteristics of ultrasonic systems is due to limited
knowledge in fluid and bubble dynamics, and combined effects of the thermodynamics
and kinetics of target contaminants.
1.5 Importance of scavenger study in sonochemistry
According to hot spot theory, the gas and interfacial regions of cavitation bubbles are
the source of reactivity (i.e., thermolysis and •OH oxidation) and the reactions in bulk
solution are due to •OH diffusing out from collapsing bubbles. In terms of •OH
distribution, it is generally believed that •OH is concentrated at the bubble-water interface
and the amount of •OH escaping from the hot spot to the bulk region is small. However,
how much •OH escapes to the bulk solution and contributes to the overall degradation of
a contaminant remains unclear, since there is no direct measurement of free •OH in the
bulk region. ESR measures •OH by detecting the spin signal of an •OH-trap adduct
(Riesz et al., 1985), and the terephthalate dosimeter measures •OH by detecting the
fluorescence of the •OH-trap adduct (Rivas et al., 2010; Song et al., 2005). These
methods do not sufficiently differentiate the locations of the •OH reaction, because the
•OH-trap adduct is not selective for •OH at either the interfacial region or in bulk solution.
However, the question is rather important, because it not only helps one to estimate the
contribution of •OH in bulk solution to the contaminant degradation as a whole, but also
to conceptualize the hot spot model and understand the spatial distribution of reaction
sites.
31
Therefore, researchers have used •OH scavenger studies in attempts to answer this
question (Beckett and Hua, 2001; Czechowska-Biskup et al., 2005; Drijvers et al., 1998;
Henglein and Kormann, 1985; Nam et al., 2003; Rivas et al., 2010; Song et al., 2005;
Tiehm and Neis, 2005; Wayment and Casadonte, 2002). Frequently used scavengers
include terephthalic acid/terephthalate (Rivas et al., 2010; Song et al., 2005), iodide
(Wayment and Casadonte, 2002), and t-butanol (Czechowska-Biskup et al., 2005; Song
et al., 2005). Less commonly used scavengers include bicarbonate (Beckett and Hua,
2001), acetic acid/acetate (Tiehm and Neis, 2005), and benzoic acid/benzoate (Drijvers et
al., 1998). These studies implicitly assume that the scavenger only quenches •OH.
1.6 Motivations of this study
Environmentally relevant matrix organics have been shown to have different effects
on sonochemical degradation of target compounds. How the matrix of organic
components affects the sonochemical degradation of contaminants is complex and not yet
completely elucidated. On one hand, the presence of an organic matrix reduces the
sonochemical degradation of target contaminants (Taylor et al., 1999). NOM is reported
to react rapidly with •OH with a second-order rate constant k•OH-SRFA =2.7×104 (mgC/L)
-1
s-1
(Goldstone et al., 2002), potentially hindering target contaminant degradation. On the
other hand, the presence of organic matter did not affect the sonochemical degradation of
methyl tert-butyl ether (MTBE) (Kang et al., 1999), perfluorooctane sulfonate (PFOS),
and perfluorooctanoate (PFOA) (Cheng et al., 2008). In order to understand the role of
matrix organics on pharmaceutical compound degradation, we degraded a hydrophilic
compound, CIPRO, and a hydrophobic compound, IBU, in the presence of terephthalic
32
acid (TA) and SRFA. For TA, we hypothesized that, similar to matrix organics in
wastewater effluent and raw drinking water, it resides and traps •OH in bulk solution.
Therefore, it is used to assess the contribution of bulk •OH in overall degradation of the
target contaminant.
Many studies have shown that at proper settings, the degradation of contaminants
under pulsed wave (PW) ultrasound is more effective than under continuous wave (CW)
ultrasound (Ciaravino et al., 1981; Deojay et al., 2011; Flynn and Church, 1984;
Neppolian et al., 2009; Yang et al., 2006). PW ultrasound allows time for the surface
active compound to diffuse to bubble-water interfaces, the sources of reactivity.
Considering many PPCPs are hydrophobic and surface active in water with a wide range
of diffusivity, PW ultrasound is expected to exhibit the potential to be a better method to
remove PPCPs, as compared to CW ultrasound. Therefore, we linked the importance of
physicochemical properties of the compound to degradation under PW ultrasound.
The impacts of physiochemical properties, such as surface excess (Γ) (Ashokkumar
et al., 2000; Tronson et al., 2003), octanol-water partition coefficient (Kow) (Nanzai et al.,
2008; Park et al., 2011), vapor pressure (p) (Mizukoshi et al., 1999; Nanzai et al., 2008),
Henry’s law constant (KH) (Ayyildiz et al., 2007; Colussi et al., 1999), second-order rate
constant (k) (Henglein and Kormann, 1985; Tauber et al., 2000), and diffusivity (Diw)
(De Visscher, 2003; Drijvers et al., 2000), on sonolysis of pharmaceuticals in aqueous
solution have been examined. These investigated physiochemical properties can be
classified into thermodynamic (e.g., Γ, Kow, p, and KH) and kinetic (e.g., k and Diw)
parameters. The thermodynamic parameters predict the ability of compounds accumulate
in specific regions of the cavitation bubbles. The kinetic parameters predict the rates that
33
the compounds accumulate on bubbles and react with ultrasound-induced reactivity (i.e.,
heat and •OH). Although the dependence of degradation rate on both thermodynamic and
kinetic parameters have been reported, the question regarding which aspect has a more
profound influence on the cavitational systems remain unclear. We used PW ultrasound
to systematically study the roles that thermodynamic (i.e., Kow) and kinetic parameters
(i.e., Diw) play in determining pharmaceutical degradation. Previous studies found that
PW ultrasound, under certain optimal conditions, enhances the degradation of a
compound, because it allows time (i.e., silent cycle) for the surface active compound to
diffuse to bubble-water interfaces, the sources of reactivity. Thus, with the silent cycle in
PW mode, we can compare the relative influence of each parameter in determining
degradation kinetics.
In addition, we investigated the degradation kinetics of these six pharmaceuticals in
DI and in a wastewater effluent to evaluate the effect of environmentally relevant
matrices. Since previous studies reported that under certain optimal conditions enhanced
degradation was observed with PW ultrasound in pure aqueous solutions, PW ultrasound
exhibits practical potential to effectively remove pharmaceuticals in a wastewater effluent.
Moreover, the selected six pharmaceuticals have been detected in wastewater effluent
(Batt et al., 2008; Langford and Thomas, 2009; Martin et al., 2011; Matamoros et al.,
2009; Nakada et al., 2004; Yu et al., 2006). Thus our study investigated the treatability of
a frequently-detected pharmaceutical by ultrasound.
1.7 Organization of the dissertation
34
The dissertation is composed of five chapters. Chapter 1 introduces the background
and motivations of the study. Chapter 2 investigates sonochemical degradation of IBU
and CIPRO in the presence of matrix organics, SRFA and TA. Chapter 3 evaluates the
commonly used •OH scavengers by using PW ultrasound. Chapter 4 studies the impact of
PW vs. CW ultrasound on degradation kinetics of the PPCPs, including CBP, IBU,
CIPRO, ATP, SFT, PG, and DP. Chapter 5 demonstrates the treatability of ultrasound to
remove the PPCPs, including 5-FU, IBU, clonidine (CLND), estriol (ESTO), nifedipine
(NIFE), and LOVS in municipal wastewater. Finally, Chapter 6 summarizes all the
conclusions we drew in this study and describes the future work to be conducted to
complement this study. During my PhD study, my interest in computational chemistry
has developed. Under the supervision of Dr. Richard Spinney in the Chemistry
department, I studied the thermodynamics and kinetics of ibuprofen with hydroxyl radical
in aqueous solution. Considering this work is independent of my own PPCPs project, it is
in the Appendix.
35
Co
nc
en
tra
tio
n
Distance
R 0 R
product productreactantreactant
Distance
•OH
Co
nc
en
tra
tio
n
Distance
R 0 R
product productreactantreactant
Distance
•OH
Tem
pera
ture
(K
)
298
5000
1900
R 0 RDistance
Tem
pera
ture
(K
)
298
5000
1900
R 0 RDistance
gas region
interfacial region
bulk region
Figure 1.1: Spatial distribution of temperature and different species in the three reaction
zones in cavitation bubble.
36
CW ultrasound
PW ultrasound ton/toff=1
time
time
inte
nsit
y
Figure 1.2: Schematic diagram of continuous wave (CW) and pulsed wave (PW)
ultrasound.
37
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53
Chapter 2: Sonochemical Degradation of Ciprofloxacin and Ibuprofen in the Presence of
Matrix Organic Compounds
21. Abstract
Ultrasonic irradiation was employed to degrade ciprofloxacin (CIPRO) and
ibuprofen (IBU), a hydrophilic and a hydrophobic compound, respectively, at the
frequencies of 20 and 620 kHz. Results show that the initial degradation rate depends on
solute concentration, ultrasound frequency, and the presence of matrix organic matter.
The initial degradation rate of IBU was approximately 1 to 4 fold faster than that of
CIPRO under equivalent conditions, indicating that the hydrophobicity of the compound
plays an important role in sonochemical degradation. In the presence of terephthalic acid
(TA), a commonly used •OH scavenger, CIPRO degradation was inhibited by a factor of
40 to 1500 depending on the frequency and initial concentration. However, the
degradation rates of IBU were only reduced between 30% and 80%. Similar to TA, the
presence of Suwannee River Fulvic Acid (SRFA) inhibited CIPRO to a greater extent
than that of IBU but overall inhibition by SRFA was dramatically less than by TA.
Although both TA and SRFA inhibited the degradation of CIPRO and IBU, the
controlling mechanisms are different. TA reacts with •OH in bulk solution but we also
suspect it accumulates on or interacts with cavitation bubbles. On the other hand, SRFA
stays in bulk solution, quenching •OH and/or associating with the target compounds. Our
54
results suggest that CIPRO primarily degrades by •OH oxidation in bulk solution.
However, both interfacial and bulk solution degradation play important roles in the
sonolysis of IBU, and the relative importance of the two pathways depends on the initial
concentration, the presence of matrix organics, and the ultrasonic frequency.
2.2 Introduction
Numerous studies have reported measuring pharmaceuticals at parts-per-trillion to
parts-per-billion concentrations in aquatic systems (Boyd et al., 2003; Ellis, 2006; Kolpin
et al., 2002; Hartmann et al. 1999; (Miege et al., 2009). Since these microconstituents are
present in much lower concentrations than conventional contaminants, it is difficult to
remove them in the presence of organic and inorganic matrices that are a thousand to a
million fold more abundant. Although the toxicity of continual low level exposure to
these compounds by humans is uncertain and these emerging pollutants are unregulated,
studies have shown adverse effects on aquatic microorganisms (Christensen et al., 2006)
and fish (Carlsson et al., 2009). To date, an effective strategy to mitigate these unforeseen
microconstituents from the aquatic environment does not exist.
Advanced oxidation processes (AOPs) have the potential to substantially reduce
contaminants from water and wastewater (Shannon et al., 2008; Snyder et al., 2003;
Westerhoff et al., 2005). Ultrasonic irradiation is an emerging AOP with advantages over
other AOPs due to its ease of use, lack of chemical addition, and unique degradation
mechanisms. Sonochemical techniques involve the use of ultrasonic waves to produce
cavitation bubbles in solution. During the collapse of cavitation bubbles localized hot
55
spots are formed (Gutierrez et al., 1986), reaching average bubble temperatures of
roughly 5000 K and pressures of approximately 500 atm (Suslick, 1990).
Considering these cavitation bubbles as microreactors, three different reaction zones
have been postulated: (i) the gaseous interiors of collapsing cavities resulting in
dissociation of volatile compounds including water through thermolysis; (ii) the
interfacial liquid surrounding cavitation bubbles where high temperature (ca. 1000-2000
K) and high concentrations of •OH (4 mM) exist (Gutierrez et al., 1991); and (iii) bulk
solution (at ambient temperature) where small amounts of •OH diffusing from bubble
interfaces may contribute to contaminant destruction reactions (Suslick, 1990; Suslick et
al., 1986). The degradation pathway of a particular organic compound depends on its
physicochemical properties, surface activity, and hydroxyl radical reactivity. Thus,
sonochemistry is a complex phenomenon with two primary mechanisms: thermolysis and
radical species reactions.
Recently, sonication of aqueous solutions has been shown to be effective for the
destruction of IBU and CIPRO. Mendez-Arriaga et al. (2008) concluded that ultrasound
may be used to eliminate IBU in aqueous solution and demonstrated that applied
sonication power, dissolved gas, pH and initial concentration played roles during the
process. Madhavan et al., (2010) studied the degradation of IBU in the sonolytic,
photocatalytic, and sonophotocatalytic systems, and observed synergistic degradation
during sonophotocatalysis. De Bel et al., observed that pH affected the ultrasonic
degradation rate, biodegradability, and ecotoxicity of aqueous CIPRO solution (De Bel et
al., 2009; De Bel et al., 2011). Our study verifies and builds upon the previous work.
56
Matrix organics have been shown to have different effects on sonochemical
degradation of the target compounds. Matrix organics may inhibit the degradation of
compounds through competing for hydroxyl radicals and/or altering the availability of
compounds to sonolytic activity (Laughrey et al., 2001; Lu and Weavers, 2002).
Terephthalic acid (TA), a typical •OH scavenger and dissolved Suwannee River FA
(SRFA), a model natural organic matter (NOM) were used as a bulk •OH scavenger and
as a model matrix organic compound, respectively. The very low concentrations of
CIPRO and IBU observed in water and wastewater compared to matrix organics indicates
the need to understand the role of matrix organics on pharmaceutical compound
degradation. IBU and CIPRO were chosen because they are, respectively, relatively
hydrophobic (true log Kow =3.97) and hydrophilic (log Kow =0.28) examples within these
classes of compounds.
2.3 Materials and Methods
IBU (99%, Acros) and CIPRO (BioChemika, 98%) were purchased from Aldrich
and used as target compounds without further purification. Table 2.1 lists relevant data
for both pharmaceuticals. Water used to prepare solutions was from a Millipore system
(R = 18.2 M -cm). Terephthalic acid (99 %) was purchased from Fluka. SRFA reference
material was obtained from the International Humic Substances Society and used without
further purification.
Two ultrasonic units, a 20 kHz ultrasonic probe system (Sonic Dismembrator 550,
Fisher Scientific) with an irradiating area of 1.27 cm2 and a 620 kHz transducer system
57
(ELAC Nautik, Inc., Kiel, Germany) with a 23 cm2 transducer were applied to irradiate
CIPRO and IBU in water. In both systems, sonochemical power density input into the
reactor was adjusted and measured by calorimetry to be approximately 400 W/L. Water
jacketed glass reactors containing 50 mL (for 20 kHz) and 125 mL (for 620 kHz) reaction
solutions were controlled at 20 °C by a cooling bath (Fisher Scientific, 1006S). Although
the power densities are the same, the cavitation fields between the two sonication systems
are different due to different geometries and irradiating surface areas.
Solution pH was continuously monitored and kept constant at pH 8.5 by adding 0.01
M NaOH or HNO3 solution during sonication. Initial individual compound
concentrations of 1, 10, and 100 M were used to probe the concentration effect under
different conditions. In select experiments, 2 mM of TA was added to CIPRO and IBU
solution. The effect of SRFA on the degradation of CIPRO and IBU was examined at
three different initial concentrations (0.66, 3.1 and 16.5 mgC/L of SRFA). All the
experiments were carried out in, at least, duplicate.
At specific times during sonolysis, 0.5 mL of sample was taken from the batch
reactors for analysis. The total volume withdrawn during a single run was less than 5% of
the total sonicating volume. Analysis of CIPRO and IBU was performed using a high
performance liquid chromatograph (HPLC) (Hewlett Packard 1100), with an Allure C18
column (5 m, 150 × 3.2 mm, Restek). CIPRO and IBU were detected at wavelengths of
274 nm and 223 nm, respectively. A pH 3.0 phosphate buffer: 7 % acetonitrile mobile
phase was ramped to 78% acetonitrile at a flow rate 0.5 mL/min. An eluent (0.8 mL/min)
of 50 % of acetonitrile: 50 % pH 3.0 phosphate buffer was used for IBU. The initial
SRFA concentration in solution was quantified by a Shimadzu TOC-5000A analyzer.
58
2.4 Results and Discussion
2.4.1 Degradation of Ciprofloxacin and Ibuprofen
The sonochemical degradation of CIPRO and IBU is shown in Figure 2.1.
Consistent with previous studies (De Bel et al., 2009; De Bel et al., 2011; Madhavan et
al., 2010; Mendez-Arriaga et al., 2008), both compounds were degraded by ultrasound.
However, due to different initial concentrations, frequencies and power input, comparing
the degradation rates in their studies and those in ours is not appropriate. In comparing
the sonolysis of CIPRO to IBU, we observed that, with one exception, the degradation
rate of IBU was faster than that of CIPRO (Table 2.2). At 620 kHz the half-life for IBU
was a factor of 3 to 4 times shorter than CIPRO at the same initial concentration. The
difference in half-life between IBU and CIPRO was less pronounced at 20 kHz with IBU
degrading approximately twice as fast as CIPRO at 100 M and similarly at 1 M.
The discrepancy in degradation rates between CIPRO and IBU is attributed to
different physiochemical properties of the compounds. First, as shown in Table 2.1, IBU
is more hydrophobic than CIPRO based on their true and apparent Kow values, indicating
that IBU is more likely to accumulate to a large extent in the cavitation bubble interfacial
region where it is subjected to thermolysis and higher [•OH].
Also, the diffusion coefficient of IBU (DIBU) in dilute solution is greater than that of
CIPRO (DCIPRO) (DIBU/DCIPRO=1.3), suggesting that IBU has a greater chance to
encounter ultrasound-induced reactivity and be degraded. Drijvers, et al., (2000) observed
that sonochemical degradation rates of halobenzenes were determined by their diffusion
59
coefficient (i.e., the compound with a greater diffusion coefficient exhibited a faster
degradation rate). Our results are consistent with previous studies and corroborates that
the degradation rate is dependent on the hydrophobicity (Nanzai et al., 2008; Okuno et al.,
2000) and diffusivity (Ayyildiz et al., 2007; Drijvers et al., 2000) of the compounds.
As shown in Table 2.2 and Figure 2.2, sonolysis at 620 kHz resulted in faster
degradation than 20 kHz for both IBU and CIPRO with rates being approximately 1 to
100 fold faster, depending on the initial concentrations. It is well-known that ultrasonic
frequency significantly affects sonolysis due to both the critical size and the lifetime of
cavitation bubbles, which consequently affect the number of cavitation bubbles, the
violence of bubble collapse, and the amount of hydroxyl radical production (Petrier et al.,
1996; Petrier et al., 1992). Many studies have reported faster degradation of organic
compounds at frequencies ranging from about 200 to 600 kHz, as compared to 20 kHz
(Chen et al., 2004; Francony and Petrier, 1996; Petrier et al., 1996; Weavers et al., 2000).
Our results are consistent with these previous studies.
The concentration effect on sonochemical degradation rates is also depicted in
Figure 2.2. The degradation rate increases as the initial concentration increases. However,
since the degradation follows apparent pseudo first-order kinetics at each concentration,
faster degradation occurs with lower initial concentration of CIPRO and IBU. The kobs
values for CIPRO and IBU at 1 M were approximately 2 to 20 times higher than those
at 10 M and 100 M under equivalent conditions (Table 2.2). Similar results have been
obtained by others (Drijvers et al., 1999; Drijvers et al., 1998; Weavers et al., 2000). A
degradation study of estrogen hormones at levels as low as 1 μg/L (~3 nM) at a frequency
60
of 20 kHz demonstrated that this trend is valid at even lower concentrations (Fu et al.,
2007).
As shown in Figure 2.2 and Table 2.2, the effect of concentration on degradation
rates is more pronounced for IBU than that of CIPRO especially at 620 kHz. For example,
at 20 kHz for IBU, μM 100
dt
dC:
μM 1dt
dC is 6.1, but for CIPRO
μM 100dt
dC:
μM 1dt
dCis 2.8;
at 620 kHz for IBU μM 100
dt
dC:
μM 1dt
dC is 38, but for CIPRO
μM 100dt
dC:
μM 1dt
dCis
33.8.
The small difference in rates at 20 kHz and large difference at 620 kHz with
increasing concentration are attributed to bubbles undergoing more oscillations during
their lifetime at 620 kHz compared to 20 kHz. At high frequency (620 kHz), stable
cavitation bubbles, experiencing more oscillations and a longer lifetime before collapsing,
dominate the whole bubble population, leading to greater adsorption of CIPRO/IBU to
bubble surfaces (Price et al., 2010; Sunartio et al., 2007). Thus, the high adsorption
capacities of stable cavitation bubbles make the concentration effect more pronounced.
At low frequency (20 kHz), transient cavitation bubbles, undergoing fewer
oscillations, are in the majority; hence, low adsorption capacity results in a smaller
concentration effect. A smaller increase in rate with increase in concentration is observed
with Langmuir type adsorption when reactive surfaces are not saturated. Our results are
consistent with Tronson et al., (2002) who reported that surface active aliphatic alcohols
significantly inhibited sonoluminescence (SL) intensity at a frequency of 515 kHz but not
61
at 20 kHz, suggesting that organics are more able to accumulate on bubble surfaces at
515 kHz compared to 20 kHz.
In addition, the effect of concentration is greater for IBU than CIPRO at both
frequencies, which is attributed to the greater hydrophobicity and diffusivity of IBU. A
compound like IBU will have both a thermodynamic and kinetic advantage increasing its
ability to accumulate on the bubble interface, resulting in thermolysis and •OH reactions.
Thus, with the increase of concentration of the target compound, the advantage becomes
more obvious. Although the potential of CIPRO and IBU to volatilize into the gas phase
of the bubble is low due to low Henry’s law constants, the higher hydrophobicity of IBU
compared to CIPRO suggests that, given a long enough bubble lifetime, IBU will
accumulate at the bubble interface. The smaller difference in rate as a function of
concentration at 620 kHz for CIPRO over IBU suggests that with CIPRO reactive bubble
surfaces are farther from saturation. Therefore, the higher hydrophobicity and diffusivity
of IBU compared to CIPRO leads to more IBU accumulation at the bubble interface; this
effect is more pronounced at 620 kHz.
The feature of faster kobs (Table 2.2) with the decrease in solute concentrations is
promising for the application of sonochemistry to eliminate PPCPs in aqueous systems,
especially for low levels of microconstituents. For instance, the t1/2 is significantly
reduced from 103 min to 6.3 min for IBU at 20 kHz as its initial concentration decreases
from 100 M to 1 M. Many PPCPs such as antibiotics, fragrances and estrogens are
more hydrophobic than IBU and CIPRO and the reported levels in water supplies are as
62
low as nM. Therefore, the sonochemical degradation of PPCPs is expected to succeed
under short treatment times.
2.4.2 Effects of Matrix Organics on Degradation
When AOPs are investigated in water treatment, the water matrix containing
inorganic and organic components significantly slows the degradation of the target
contaminants (Cheng et al., 2010;2008; Westerhoff et al., 1999). How the matrix of
organic components affect the sonochemical degradation of contaminants is complex and
not yet completely elucidated. On one hand, the presence of an organic matrix reduces
the sonochemical degradation of target contaminants (Lu and Weavers, 2002; Taylor et
al., 1999). NOM is reported to react rapidly with •OH with a second-order rate constant
k•OH-SRFA =2.7×104 (mgC/L)
-1 s
-1 (Goldstone et al., 2002), potentially hindering target
contaminant degradation. Lu and Weavers (2002) observed that humic acid (HA)
decreased the sonochemical degradation rate of 4-chlorobiphenyl (4-CB) and the effect
was more pronounced with an increase in HA concentration. Additionally, the sonication
of polycyclic aromatic hydrocarbons (PAHs) with fulvic acid (FA) dramatically reduced
the degradation rate of PAHs (Taylor et al., 1999). On the other hand, the presence of
organic matter did not affect the sonochemical degradation of methyl tert-butyl ether
(MTBE) (Kang et al., 1999), perfluorooctane sulfonate (PFOS), and perfluorooctanoate
(PFOA) (Cheng et al., 2008). In order to remove PPCPs from wastewater effluent or raw
drinking water, ultrasound will need to be selective for PPCPs over matrix organics.
2.4.2.1 Terephthalic Acid
63
To gain a sense of the interference of non-target organic compounds, we first
employed TA as a bulk •OH scavenger. TA has been widely used to monitor the yields of
•OH in different oxidation systems (Matthews, 1980; Page et al., 2010; Price and Lenz,
1993). Sonochemical researchers have employed it as a bulk •OH scavenger (Price et al.,
1997; Song et al., 2005; Villeneuve et al., 2009) due to its high second-order •OH rate
constant (3.3 × 109 M
-1 s
-1, (Buxton et al., 1988) and its deprotonated form in neutral and
alkaline solution (pKa values, 3.52 and 4.46), suggesting that it will stay in bulk solution
during sonication (Mason et al., 1994; Song et al., 2005). Consistent with others, we
assume that, similar to matrix organics in wastewater effluent and raw drinking water,
TA resides and traps •OH in bulk solution. Therefore, it is used to assess the contribution
of bulk •OH in overall degradation of the target contaminant.
As shown in Table 2.2 and Figure 2.2, in the presence of 2 mM TA, the degradation
rates at different initial CIPRO and IBU concentrations are significantly reduced
compared to no TA present (p<0.01). The degradation rate of CIPRO was reduced by a
factor of between 40 and 70 at 620 kHz, and between 70 and 1500 at 20 kHz in the
presence of TA as compared to degradation in its absence. The degradation rate of IBU
was decreased by a factor of 2 and 10 at 20 kHz and 2 at 620 kHz in the presence of TA
as compared to that in its absence.
To predict the impact of TA on degradation rates, the bulk •OH quenching ratio (Q)
of TA was calculated (MWH, 2005):
[C]k][Ck
][CkQ
CTA
C
TA (2.1)
64
where kC and kTA are the reported second order rate constants of •OH reacting with the
target compound (i.e., CIPRO or IBU) and TA, respectively. [C] and [TA] are the
concentrations of the target compound and TA, respectively. Assuming bulk reactions of
TA and the target compounds, the predicted impact of TA was determined by multiplying
QTA by the observed degradation rate of IBU or CIPRO in the absence of TA:
TA w/oobs 0,
TA
predicted 0,dt
dCQ
dt
dC (2.2)
As shown in Figure 2.3, the observed values for IBU are significantly higher than the
predicted values at all concentrations and frequencies, indicating that IBU degradation is
reduced less by TA than predicted if both compounds are equally distributed in the
system. The preferential degradation of IBU is consistent with it preferentially
accumulating on cavitation bubble surfaces. However, for CIPRO the observed values are
lower than the predicted values at 100 M under both 20 and 620 kHz frequencies. The
higher predicted rates than the observed rates at 100 M initial concentration indicates
that TA is preferentially degraded over CIPRO.
In addition to •OH reaction in bulk solution, sonochemical degradation of
compounds occurs by high temperature and •OH reaction in and on cavitation bubbles
(Gutierrez et al., 1986). Thus, sonochemical degradation is described by eqn. 2.3:
bubbleinterfacebulkobs dt
dC
dt
dC
dt
dC
dt
dC
(2.3)
where dt
dC
represents the degradation rate of either CIPRO or IBU observed (obs),
occurring in bulk solution (bulk), occurring at the interface (interface), and occurring in
the gas region of cavitation bubble (bubble), respectively. Assuming that TA remains in
65
bulk solution and quenches bulk •OH, its presence only impacts bulkdt
dC
. A predicted
degradation rate with TA similar to or greater than the observed degradation rate suggests
that reaction of CIPRO occurs dominantly in bulk solution
(i.e.,bubbleinterfacebulk dt
dC
dt
dC
dt
dC). Therefore, at 100 M either CIPRO does not partition
to the cavitation bubble interface and gas region, TA does partition to the bubble
interface, or both.
De Bel et al., (2011) modeled CIPRO degradation in sonochemical systems and
concluded that most of the CIPRO degradation takes place at the bubble-liquid interface.
However, an apparent log Kow value of -1.52 at the pH used in our experiments suggests
that CIPRO will not accumulate at bubble surfaces. Our results are consistent with
CIPRO mainly reacting in bulk solution, contrary to De Bel et al., (2011).
Table 2.2 and Figure 2.2 demonstrate that the presence of TA decreased IBU
degradation to different extents. As shown in eqn. 2.3, with the assumption that TA
quenches •OH in bulk solution, the difference between the observed degradation rate and
the predicted degradation rate is attributed to the portion that degraded in non-bulk
solution (i.e., bubbleinterface dt
dC
dt
dC). With low Henry’s law constants for IBU and TA, the
portion reacting in the gas region of cavitation bubbles (i.e., bubbledt
dC) is expected to be
negligible; hence these results suggest that the majority of IBU degradation occurs at
interface (i.e., interfacedt
dC) at 20 and 620 kHz, respectively. The smaller effect of TA at 620
66
kHz and with IBU over CIPRO is consistent with more IBU accumulation on the bubble
surfaces, particularly at 620 kHz.
The difference in cavitation bubble features between the 20 kHz probe and the 620
kHz transducer systems may also result in a larger effect of TA on CIPRO and IBU
degradation at 20 kHz. Hydroxyl radical is a direct product of the collapse of cavitation
bubbles. Sonochemically-induced chemiluminescence has been applied to elucidate the
spatial distribution of ultrasonic cavitation in solution (Chen et al., 2004; Petrier et al.,
1994; Price et al., 2010). Using a 20 kHz probe system, Chen et al (2004) observed
localized cavitation in the vicinity of the probe surface extending approximately 1 cm
into solution. In a study by Petrier et al. (1994), a 487 kHz transducer sonication system
resulted in a dispersive spatial distribution of cavitation bubbles. In our study, a similar
localized bubble cloud is expected to be in the vicinity of the sonication probe tip at 20
kHz, while the cavitation bubbles at 620 kHz system are expected to affect the entire
reactor. The dense bubble cloud at 20 kHz results in TA being more likely to enter the
localized bubble cloud and quench ultrasound-induced reactivity. TA penetrating the
bubble cloud, therefore, reduces CIPRO and IBU degradation more at 20 kHz than at 620
kHz. The more pronounced reduction in degradation for CIPRO over IBU with TA
present does suggest that IBU does accumulate more than CIPRO even at 20 kHz but the
ability or extent of CIPRO and IBU to accumulate on bubbles is reduced compared to
620 kHz.
Predicted CIPRO degradation rates in the presence of TA greater than the
corresponding observed rates indicating preferential degradation of TA over CIPRO has
caused us to reexamine the assumption that TA resides and traps •OH in bulk solution.
67
Although the apparent log Kow of TA (-2.15 at pH 8.5) indicates it will not be favorable
to partition to bubble interfaces as compared to CIPRO (log Kow = -1.52) and IBU (log
Kow =0.24), previous studies have shown that mass transport of target compounds to the
bubble surface is controlled by diffusion (De Visscher, 2003; DeVisscher et al., 1996;
Drijvers et al., 2000; Yang et al., 2005). The calculated diffusion coefficient of TA (DTA)
in dilute solution is greater than that of CIPRO (DCIPRO) and IBU (DIBU) with diffusion
coefficient ratios, DTA/DCIPRO =1.5 and DTA/DIBU =1.1, respectively, suggesting that TA
will diffuse more rapidly in solution and to bubbles. Therefore, the discrepancy between
these results indicating preferential degradation of TA over CIPRO at 100 M (Figure
2.3) and previous studies indicating degradation of CIPRO occurs primarily at bubble
interfaces combined with the higher diffusion coefficient of TA suggests that TA is not
an optimal bulk phase •OH scavenger. We suspect that TA may partition to or otherwise
affect cavitation bubbles.
2.4.2.2 Suwannee River Fulvic Acid
It is well known that the presence of NOM, which is ubiquitous in natural waters,
poses significant challenges in the implementation of AOPs due to the reaction of NOM
with •OH (Snyder et al., 2003; Westerhoff et al., 1999). Therefore, CIPRO and IBU were
sonicated together with SRFA, a representative water soluble NOM, to elucidate the
ability of NOM to reduce target contaminant degradation and to examine the impact of
hydrophobicity of the pharmaceutical on degradation in the presence of SRFA.
As shown in Table 2.2 and Figure 2.4, similar to TA, the presence of SRFA reduced
the initial degradation rate of CIPRO and IBU, but the reduction in the rate by TA was
68
greater than by SRFA. At 620 kHz the degradation rates of CIPRO at 1 and 10 M are
almost completely inhibited by TA but ca. 60-70% reduced by SRFA even at the highest
SRFA concentration. SRFA inhibited the degradation rate of IBU at 20 kHz (Figure 2.4a),
but little reduction in rates was observed at 620 kHz (Figure 2.4b). The presence of
SRFA had a more pronounced effect on reducing the degradation rate of CIPRO at both
20 and 620 kHz, particularly as the SRFA concentration increased. This decrease is more
pronounced at 20 kHz than 620 kHz.
In comparing SRFA to TA, a greater reduction of degradation of CIPRO and IBU by
TA than SRFA was observed. SRFA is rich in carboxyl and hydroxyl functional groups,
which makes SRFA very soluble and hydrophilic (Chiou, 2002). Thus, SRFA is expected
to remain in bulk solution and not expected accumulate on bubble surfaces. The weight-
average molecular weight of SRFA ranges from 1000 to 26000 Dalton (Wagoner et al.,
1997). Molecular weight of a compound is correlated to size of a compound and the size
of a compound is inversely related to diffusivity. Although the diffusivity of SRFA has
not been reported, it is expected to be significant smaller than CIPRO, IBU, and TA
because of the large molecular weight of SRFA (Wagoner et al., 1997) compared to
CIPRO, IBU, and TA. In addition, Goldstone et al., (2002) measured its second order rate
constant with •OH to be 2.7×104 (mgC/L)
-1 s
-1. At the NOM and CIPRO and IBU
concentrations in these experiments, SRFA will compete with CIPRO and IBU for •OH
in bulk solution. Therefore, SRFA is expected to stay in bulk solution, diffuse
comparatively slowly through solution, and quench •OH reactions with CIPRO and IBU
in bulk solution at high [SRFA] and/or low [CIPRO] and [IBU]. TA, on the contrary, is
an effective bulk •OH quencher under all conditions and diffuses quickly in solution.
69
Previous studies have observed varying effects of organic matter on the sonolysis of
contaminants. Using a 20 kHz probe, Taylor et al., (1999) observed that SRFA inhibited
sonochemical degradation of anthracene, phenanthrene, and pyrene. They attributed
inhibition by SRFA to sequestering of the compounds away from the cavitation bubbles
by SRFA, or changing cavitational conditions due to surface tension changes in the
presence of SRFA. Lu and Weavers (2002) observed that Aldrich HA affected 4-CB
degradation at 20 kHz and its presence increased with an increase of HA concentration.
The reduction in rate was related to two factors: 4-CB and HA competing for •OH, and 4-
CB associating with HA sequestering 4-CB from cavitation sites. Kang et al., (1999)
reported no effect of Fluka HA on MTBE degradation at 358 kHz, and attributed the lack
of effect to the high volatility of MTBE, resulting in MTBE reacting in the cavitation
bubbles. Cheng et al., (2008) found that the effect of NOM (i.e., SRHA and SRFA) on
sonochemical degradation rates of PFOS and PFOA was not significant due to little
interference by the NOM on the interfacial region of cavitation bubbles. Based on our
results and those of others described above, organic matter may compete for •OH,
sequester target contaminant from cavitation bubbles, and alter cavitation bubbles.
Contaminant degradation may be inhibited to varying degrees depending on the nature of
the cavitation and the characteristics of the contaminant.
SRFA competes with CIPRO and IBU for •OH, resulting in a reduction in CIPRO
and IBU degradation. Similar to quenching by TA, the quenching ratio (Q) of SRFA in
bulk solution was calculated by eqn. 2.4 (MWH, 2005):
S)-(1[C]k][SRFAk
)S(1][CkQ
CSRFA
C
SRFA (2.4)
70
where kSRFA is the second order rate constant of •OH reacting with SRFA. S is the
fraction of C sequestered by SRFA calculated from Chin et al., (1997). The term 1-S
represents the fraction of CIPRO or IBU that is not interacting with SRFA and thus
available in solution. Similar to the quenching ratio with TA, assuming SRFA and C are
equally distributed in solution, the degradation rate of CIPRO or IBU in the presence of
SRFA is predicted by multiplying the quenching ratio by the initial degradation rate in
the absence of SRFA.
SRFA w/oobs 0,
SRFA
predicted 0,dt
dC)1(Q
dt
dCS (2.5)
where kobs w/o SRFA is the observed first-order degradation rate (min-1
) of IBU or CIPRO in
the absence of SRFA; C0 is the initial concentration of the target compound. As shown in
Figure 2.5, all the reported values are greater than or equal to the predicted values. The
trend suggests that, as compared to SRFA, both CIPRO and IBU have a relatively greater
mobility towards the cavitation bubbles. The difference between observed rates and
predicted rates is attributed to CIPRO and IBU reaction at cavitation bubble interfaces.
The fractions of CIPRO and IBU associating with SRFA at different concentrations
of SRFA are approximately 1%, 5% and 25% at SRFA concentrations of 0.66, 3.1, 16.5
mgC/L, respectively (Table 2.3). The predicted fractions of the CIPRO and IBU
sequestered by SRFA are smaller than the amount of reduction in degradation. For
example, a 95% reduction in IBU degradation at a concentration of 1 µM at 20 kHz was
observed but only 1% of IBU is predicted to be associated with SRFA. Thus,
sequestration of CIPRO and IBU by SRFA binding alone is not sufficient to explain the
reduction in degradation.
71
Although Taylor et al., (1999) attributed reduced PAHs degradation to SRFA
sequestration of PAHs from cavitation bubbles, they also proposed that SRFA may alter
cavitation bubbles via surface tension changes. Yates and von Wandruszka (1999)
showed the surface tension of bulk solution remained unchanged (ca. 72 mN/m) at 500
mg/L SRFA and pH 8.0. In our work, the highest concentration of SRFA used was 16.5
mgC/L (25 mg/L). In addition, NOM is more aggregated in neutral and basic solution
than in acidic solution, resulting in less ability to change surface tension at pH 8.5 (Yates
and von Wandruszka, 1999). Therefore, we do not expect that SRFA directly impacts
cavitation bubbles.
In comparing Figure 2.4a and b, more interference of SRFA is observed at 20 kHz
than 620 kHz with both IBU and CIPRO. The minimal influence of SRFA on sonolysis
of IBU at 620 kHz suggests that at 620 kHz IBU accumulates on the bubbles to a greater
extent compared to 20 kHz. The limited impact of SRFA on IBU degradation at 620 kHz
is attributed to the affinity of IBU to bubble surfaces and the characteristics of the
cavitation bubbles produced at 620 kHz. The features of ultrasound at 620 kHz include
more bubble oscillations within a bubble lifetime, more bubbles, and a more dispersive
bubble population (Petrier et al., 1996; Petrier et al., 1992), all favoring accumulation of
compounds. On the contrary, the bubbles generated at 20 kHz are not favorable for
accumulation due to the localized bubble population (Petrier, et al. 1992).
Similar to the observation that TA impacted IBU degradation less than CIPRO, the
presence of SRFA has a more pronounced effect on the degradation rate of CIPRO than
IBU. IBU is a more hydrophobic compound with a higher diffusivity as compared to
CIPRO (DIBU/DCIPRO=1.3); therefore, IBU is more likely to transport to and accumulate
72
on the bubble surface. On the contrary, SRFA has a sizeable volume (Zhou et al., 2000),
which hinders its propensity and ability to move to cavitation bubbles, the sources of
reactivity. The greater effect of SRFA on CIPRO degradation than IBU degradation also
supports our suggestion that CIPRO undergoes bulk solution degradation to a larger
extent than IBU due to the differences in their hydrophobicities and diffusivities.
2.5 Conclusions
Our results show that in the absence of matrix organics the degradation rates of
CIPRO and IBU depend on their initial concentrations and ultrasonic frequencies: a short
half-life is observed as the initial concentration decreases, and 620 kHz is more effective
as compared to 20 kHz.
The sonochemical degradation of target compounds proceeds at different rates in the
presence of matrix organics, which provides insight to the application of ultrasound in
wastewater and drinking water treatment. Specifically, a more hydrophobic target
compound with higher diffusivity is impacted little by presence of matrix organics. The
diffusivity and concentration of matrix organics determine treatability by ultrasound. A
larger size and lower concentration of matrix organics in water have a smaller impact on
the treatability of target contaminants by ultrasound as compared to a smaller size and
higher concentration of matrix organics. Obviously, it is optimal to remove of matrix
organics (especially small sized organics) before ultrasound treatment to maximize its
efficiency. However, when this is not feasible, ultrasound may be effective for removal of
hydrophobic and surface active contaminants in the presence of matrix organics that
reside in bulk solution, such as SRFA. Our results provide further insight on the minimal
73
influence of FA and HA on sonochemical degradation of MTBE, PFOS, and PFOA. Like
MTBE, PFOS and PFOA, IBU migrates from bulk solution to cavitation bubbles, the
sources of reactivity. Our results indicate that CIPRO does not migrate to cavitation
bubbles resulting in greater competition with NOM for bulk •OH. The cavitation
conditions in 20 kHz ultrasound limit the ability of IBU to accumulate on bubble surfaces
resulting in more competition with NOM for bulk •OH.
74
Table 2.1: Physicochemical Properties of CIPRO and IBU
75
Table 2.2: Degradation Kinetics of CIPRO and IBU under Various Conditions.
a substrate (S)
b concentration of substrate (Cs)
c target compound (TC)
d concentration of target compound (CTC)
e errors were not listed because of limited space.
76
Table 2.3: CIPRO and IBU Association with Suwannee River Fulvic Acid
compounds Koc (L/kg) organic carbon
(mgC/L) sequestering ratio, S (%)
a
CIPRO 2.00×10
4 pH=7.3
(Cardoza et al., 2005)
0.66 1.30
3.10 5.83
16.50 24.77
IBU 1.58×10
4 pH=7.0
(Yamamoto et al., 2005)
0.66 1.04
3.10 4.68
16.50 20.73
a sequestering ratio, S, (%) was calculated from Chin et al., (1997)
77
0 2 4 6 8 10 12 140.0
0.2
0.4
0.6
0.8
1.0
1.2
1 M IBU
10 M IBU
1 M CIPRO
10 M CIPRO
C/C
0
Sonication time (min)
Figure 2.1: Sonolysis of CIPRO and IBU at 620 kHz (pH 8.5, 20 °C, and sonication
power density at 400 W/L).
78
1 10 1001E-3
0.01
0.1
1
10
[IBU]0 ( M)
20 kHz
20 kHz with TA
620 kHz
620 kHz with TA
d[I
BU
]/d
t|0 (
M/m
in)
a
1 10 1001E-4
1E-3
0.01
0.1
1
10
b
20 kHz
20 kHz with TA
620 kHz
620 kHz with TA
[CIPRO]0 ( M)
d[C
IPR
O]/
dt|
0 (
M/m
in)
Figure 2.2: Sonochemical degradation rates at different concentrations of IBU (a) and
CIPRO (b) with 2 mM terephthalic acid (TA) (pH 8.5, 20 °C and sonication power
density at 400 W/L).
79
0 20 40 60 80 1000.000
0.005
0.010
0.015
0.020
0.025
0 20 40 60 80 1000.0
0.1
0.2
0.3
0.4
CIPRO predicted
CIPRO observed
[CIPRO]0 or [IBU]
0 ( M)
d[C
IPR
O]/
dt|
0 (
M/m
in)
IBU predicted
IBU observed
d[I
BU
]/d
t|0 (
M/m
in)
a: 20 kHz
0.00
0.05
0.10
0.15
0.20
0.25
0.30
0.35b: 620 kHz
CIPRO predicted
CIPRO observed
0 20 40 60 80 1000
1
2
3
4
5
6
7
8
9
IBU predicted
IBU observed
d[I
BU
]/d
t|0 (
M/m
in)
d[C
IPR
O]/
dt|
0 (
M/m
in)
[CIPRO]0 or [IBU]
0 ( M)
Figure 2.3: Predicted and observed initial degradation rates with 2 mM terephthalic acid
(TA) at the frequency of 20 kHz (a) and 620 kHz (b) (pH 8.5, 20 °C and sonication power
density at 400 W/L).
80
0.01
0.1
1
CIPRO (10 M)
d[C
]/d
t|0 (
M/m
in)
DI
0.66 mgC/L FA
3.1 mg C/L FA
16.5 mg C/L FA
CIPRO (1 M)IBU (10 M)IBU (1 M)
a: 20 kHz
0.01
0.1
1
10
d[C
]/d
t|0 (
M/m
in)
DI
0.66 mgC/L FA
3.1 mg C/L FA
16.5 mg C/L FA
CIPRO (10 M)CIPRO (1 M)IBU (10 M)IBU (1 M)
b: 620 kHz
Figure 2.4: The sonochemical degradation rates of CIPRO and IBU with different
concentrations of SRFA at 20 kHz (a) and 620 kHz (b) (pH 8.5, 20 °C, and sonication
power density at 400 W/L).
81
0 2 4 6 8 10 12 14 160.000
0.025
0.050
0.075
0.100
0.125a: 20 kHz and C
0= 1 M
CIPRO predicted
CIPRO observed
IBU predicted
IBU observed
d[C
]/d
t|0 (
M/m
in)
[SRFA] (mgC/L)
0 2 4 6 8 10 12 14 160.0
0.1
0.2
0.3
0.4
0.5
d[C
]/d
t|0 (
M/m
in)
b: 20 kHz and C0= 10 M
CIPRO predicted
CIPRO observed
IBU predicted
IBU observed
[SRFA] (mgC/L)
0 2 4 6 8 10 12 14 160.0
0.1
0.2
0.3
0.4
0.5
d[C
]/d
t|0 (
M/m
in)
c: 620 kHz and C0= 1 M
CIPRO predicted
CIPRO observed
IBU predicted
IBU observed
[SRFA] (mgC/L)
0 2 4 6 8 10 12 14 16 180.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
d[C
]/d
t|0 (
M/m
in)
d: 620 kHz and C0= 10 M
[SRFA] (mgC/L)
CIPRO predicted
CIPRO observed
IBU predicted
IBU observed
Figure 2.5: Predicted and observed initial degradation rates of CIPRO and IBU with
SRFA at 20 kHz (a. 1 M and b. 10 M) and 620 kHz (c. 1 M and d. 10 M) (pH 8.5,
20 °C, and sonication power density at 400 W/L).
82
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91
Chapter 3: Using Pulsed Wave Ultrasound to Evaluate the Suitability of Hydroxyl
Radical Scavengers in Sonochemical Systems
3.1 Abstract
Hydroxyl radical (•OH) scavengers are commonly used in sonochemistry to probe
the site and nature of reaction in aqueous cavitational systems. By using pulsed wave
(PW) ultrasound we evaluated the performance of several different •OH scavengers (i.e.,
formic acid, carbonic acid, terephthalic acid/terephthalate, iodide, methanesulfonate,
benzenesulfonate, and acetic acid/acetate) in a sonochemical system to determine which
•OH scavengers react only in bulk solution and which •OH scavengers interact with
cavitation bubbles. The ability of each scavenger to interact with cavitation bubbles was
assessed by comparing the pulse enhancement (PE) of 10 M of a probe compound,
carbamazepine (CBZ), in the presence and absence of a scavenger. Based on PE results,
acetic acid/acetate appears to scavenge •OH in bulk solution, and not interact with
cavitation bubbles. Methanesulfonate acts as reaction promoter, increasing rather than
inhibiting the degradation of CBZ. For formic acid, carbonic acid, terephthalic
acid/terephthalate, benzenesulfonate, and iodide, the PE was significantly decreased
compared to in the absence of the scavenger. These scavengers not only quench •OH in
bulk solution but also affect the cavity interface. The robustness of acetic acid/acetate as
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a bulk •OH scavenger was validated for pH values between 3.5 and 8.9 and acetic
acid/acetate concentrations from 0.5 mM to 0.1 M.
3.2 Introduction
The transformation of organic pollutants in aqueous solution using ultrasound has
been studied for decades [1-3]. In general, ultrasonic waves produce cavitation bubbles in
water and the collapse of cavitation bubbles generates localized hot spots [4]. The hot
spots initiate thermolytic and redox reactions with pollutants [4]. Hydroxyl radical (•OH)
is the primary species attributed to redox reactions. •OH forms in the gaseous bubble core
and diffuses to bulk solution to oxidize contaminants [5]. This process is considered an
important degradation pathway for hydrophilic and non-polar pollutants [6-7].
It is generally believed that •OH concentrates at the bubble-water interface and the
amount of •OH escaping from the hot spot and diffusing to bulk solution is small [8].
However, the question regarding how much •OH escapes to bulk solution and contributes
to the overall degradation of a pollutant remains unclear, since there is no direct
measurement of free •OH in bulk solution.
Two methods are commonly used to detect •OH in cavitational systems. Electron
spin resonance (ESR) spectroscopy measures •OH by detecting the spin signal of the
•OH-trap adduct [9]. The terephthalate dosimeter measures •OH by detecting the
fluorescence signal of the •OH-trap adduct [10]. However, these methods do not
sufficiently differentiate the locations of •OH reaction, because the •OH-trap adduct
forms proportionally based on the concentration of •OH and the •OH trap in the
cavitation bubble, at bubble-water interface and in bulk solution.
93
The question of location of reaction is important. It not only helps one to estimate
the contribution of bulk solution •OH to contaminant degradation as a whole, but also to
understand the spatial distribution of reaction sites. Therefore, researchers have used •OH
scavenger studies to answer this question [11-20]. Frequently used scavengers include
terephthalic acid/terephthalate [11, 13], iodide [14-15], and t-butanol [11, 16]. Less
commonly used scavengers include bicarbonate [17], acetic acid/acetate [18], and
benzoic acid/benzoate [19]. These studies implicitly assume that the scavenger only
quenches •OH.
In reality, the •OH scavengers may interact with the bubble-water interface and enter
cavitation bubbles. Volatile scavengers not only quench •OH in gas, interfacial, and bulk
regions of cavitation bubbles, but also decrease the energy available for thermolysis of
H2O, reducing •OH formation [12, 20-21]. For example, Rae et al., demonstrated that
even trace amounts of alcohols (such as n-butanol or t-butanol) evaporated into the
bubble during the expansion phase of bubbles, decreasing the measured bubble
temperature [21]. Therefore, because alcohol scavengers, such as t-butanol, methanol,
and n-hexanol, are already known to interact with cavitation bubbles, they were not
investigated in our study. •OH scavenging compounds with surface activity may interact
with cavitation bubbles, affecting bubble collapse dynamics. As these surface active
compounds preferentially partition to bubble surfaces, they affect assessments of bulk
•OH reactivity as well. As a consequence, without assurance that scavengers do not
interact with cavitation bubbles, attempts to infer mechanisms or sites of reaction, such as
bulk •OH reaction to overall contaminant degradation using these scavengers is
speculative.
94
In this study, pulsed wave (PW) ultrasound was used to systematically study
whether and how •OH scavengers affect cavitation bubbles. Previous studies [22-25]
found that PW ultrasound, under certain optimal conditions, enhances the degradation of
a compound, because it allows time for the compound to diffuse to bubble-water
interfaces, the sources of reactivity. On the contrary, an ideal bulk •OH scavenger stays in
bulk solution and does not accumulate in the interfacial and gas region of cavitation
bubbles between two successive pulses. We used a surface active probe compound, CBZ,
as a reference compound that undergoes faster degradation under PW conditions
compared to CW due to CBZ accumulation at bubble-water interfaces. Thus, in the
presence of a bulk •OH scavenger, faster degradation under PW over CW is expected.
The PW enhancement is due to accumulation at the bubble-water interfaces. Any change
in the enhancement is an indication of interaction of the scavengers with CBZ or the
cavitation bubbles.
3.3 Experimental
CBZ (Sigma-Aldrich, 99%), KI (Acros, 99%), sodium formate (Fluka, 99%),
sodium acetate (Fisher-Scientific, 99%), sodium bicarbonate (Fisher, 99%), sodium
phosphate (monobasic and dibasic) (Fisher-Scientific, 99%), terephthalic acid (TA)
(Sigma-Aldrich, 98%), benzenesulfonate (BS) (Sigma-Aldrich, 98%), methanesulfonate
(MS) (Sigma-Aldrich, 98%), and phosphoric acid (Fisher-Scientific, 85%) were used as
received. Stock solutions of these chemicals were prepared using Millipore purified water
(Millipore, MA) with the resistivity R = 18.2 MΩ cm.
95
A 10 µM initial concentration of CBZ was used for all experiments. The •OH
scavenger concentration was 1 mM, a 100-fold excess relative to CBZ, unless specified.
A 1 mM phosphate buffer was used because of its low •OH reactivity (104 ~ 10
5 M
-1 s
-1)
[26]. The pH of solution was adjusted by adding phosphoric acid or NaOH (Reagent ACS,
Pellets 97%, Fisher Scientific).
Ultrasound at 205 kHz was emitted from a USW 51-52 ultrasonic flat plate
transducer (A = 23.4 cm2) (ELAC Nautik, Inc., Kiel, Germany) into a glass vessel with a
volume of 300 mL. The reactor was equipped with a cooling water jacket to maintain
solutions at 20 °C. A SM-1020 Function/Pulse generator (Signametrics Corp., Seattle,
WA) generated sound waves continuously or by pulsing with an operation mode of 100
ms on and 100 ms off. A linear amplifier (T & C Power Conversion, Inc., Rochester, NY)
magnified the generated electrical signal. The acoustic energy density to the reactor,
determined by calorimetry, was measured to be 45 W/L. Calorimetry was monitored
periodically during experiments to confirm system functionality. During sonication, 0.5
mL samples were taken from the reactor at designated times using a 1 mL glass syringe
(Gastight 1001, Hamilton Corp.) for chemical analysis. The sample volume taken during
the course of sonication did not exceed 1 % of the total volume in order to maintain the
distribution of the acoustic field in the reactor. The experiments were carried out, at least,
in duplicate. Statistical t-test analysis was conducted using SPSS 13.0 (LEAD
Technologies Inc).
A Hewlett-Packard 1100 high performance liquid chromatograph (HPLC) equipped
with a diode array detector (DAD) was used to analyze the concentration following
sonolysis of CBZ. A 5 µm, 150 × 2.1 mm SB-C18 column (883700-922, Agilent
96
Technologies) was employed. An isocratic flow of acetonitrile/water (40/60, v/v) was the
mobile phase for the quantification of CBZ. The injection volume was 50 µL and the UV
wavelength was set at 220 nm.
Gibbs surface tension ( ) was measured by a McVan Analite Surface Tension meter
(2141, McVan Instruments) with a glass Wilhelmy plate. All measurements were carried
out in triplicate at room temperature. Gibbs surface excess ( ) was obtained from the
change of : surface tension,
dln[C]
dγ
RT
1Γ (3.1)
where R is gas constant, T is temperature, and C is concentration of solute. Surface
tension was smoothed using a Fast Fourier Transform (FFT) tool to eliminate instrument
noise (Origin 8.1, OriginLab Inc).
3.4. Results and Discussion
3.4.1. Degradation of CBZ
The sonochemical degradation rate of the probe compound, CBZ, in the absence and
presence of •OH scavengers at pH 3.5 is shown in Figure 3.1. In the absence of •OH
scavengers, CBZ degraded 5.8 % (p < 0.01) faster by PW ultrasound compared to CW
ultrasound. Similar to previous work [22-25], the higher degradation rate under PW
ultrasound indicates that a higher fraction of CBZ is degraded in the interfacial region of
cavitation bubbles.
During the time interval between two successive pulses, CBZ molecules accumulate
on the surface of the bubble. In the subsequent pulse more CBZ molecules react
97
compared to CW ultrasound. Faster degradation by PW ultrasound is consistent with
Naddeo et al. [27]. They attributed the majority of decomposition of CBZ to reaction in
the vicinity of the cavitation bubbles.
Figure 3.1 also shows the initial degradation rates of CBZ in the presence of 1 mM
concentration of each individual •OH scavenging agent. With the exception of MS, the
presence of the •OH scavenger reduced the initial degradation rate of CBZ in both CW
and PW conditions. Using eqn. 3.2, the inhibitory effect of each individual scavenger on
the sonolysis of CBZ was calculated.
％100
dt
Cd
dt
Cd
dt
Cd
% inhibition
scavengerwithout
scavengerwith scavengerwithout
(3.2)
where scavengerwith
dt
d[C] and
scavengerwithout dt
d[C]are the initial degradation rates of CBZ
with or without an •OH scavenger, respectively. Figure 2 indicates that these seven
scavengers have different effects on the sonolysis of CBZ.
The changing level of inhibition of all the scavenger candidates, as shown in Figure
3.2, implies that they exert different influences on the sonochemical system. In addition,
the benefit of pulsing in the presence of the scavenger is only observed with MS and AA.
3.4.2 PE of CBZ in the presence of the •OH scavengers
In order to investigate the effects of scavengers on cavitation bubbles, a pulse
enhancement (PE) was calculated. By comparing the degradation of CBZ between CW
and PW ultrasound through PE, the effect of pulsing was determined:
98
％100
dt
Cd
dt
Cd
dt
Cd
(%) PE
CW
CWPW (3.3)
where CW/PW
dt
d[C] is the degradation rate of CBZ under CW or PW ultrasound. In the
presence of a bulk •OH scavenger, the PE of CBZ is expected to be equal to that in the
absence of the scavenger, indicating that the presence of the scavenger does not have any
pronounced effects on the degradation of CBZ (i.e., competing for adsorption sites at
interface and diffusing to gas region).
In the absence of •OH scavengers, the PE of CBZ is 5.8%. The PE of CBZ in the
presence of various scavengers is shown in Figure 3.3. In the presence of FA, CA, TA,
and KI, the PE of CBZ is negative ranging from -20% to -5%. Again, a similar PE value
with and without the scavenger indicates its influence on cavitation bubbles is negligible.
The results suggest that FA, CA, TA, and KI not only scavenge •OH in the system but
also significantly affect cavitation bubbles. In the presence of MS, BS, and AA, the PE
ranges from 0.92% to 11.3%, suggesting little influence on cavitation bubbles. To
understand the effects of these scavengers on cavitation bubbles, we will discuss each
scavenger separately below.
3.4.2.1. Formic acid
FA has been characterized as an ideal •OH scavenger in radiolysis [28]. It has a fast
reaction rate constant and known reaction pathways with •OH. However, when it comes
to its application in sonochemistry, FA shows a negative pulse enhancement (PE=-6.2%).
99
Partitioning of FA and its ultrasound-induced reaction products to the gaseous
bubble and bubble interface has been reported in the literature. Hart and Henglein
observed CO production during sonication of FA at 300 kHz [29]. Substantial formation
of CO in sonicated aqueous FA solution is attributed to thermal decomposition of FA
(eqn. 3.4), indicating that FA enters the high temperature gas region of cavitation bubbles.
HCOOH CO + H2O (3.4)
In addition, Jolly et al., investigated products formed from •OH reacting with FA during
laser-irradiated photolysis [30]. They proposed that the oxidation of FA by •OH releases
CO2 according to eqns. 3.5 and 3.6:
HCOOH + •OH •COOH + H2O (3.5)
•COOH + •OH CO2 + H2O (3.6)
Gaseous CO2 formed enters cavitation bubbles and decreases the polytropic index, , (κair
=1.405 and κCO2 =1.312) of the gas phase, resulting in a decreased bubble collapse
temperature as shown in eqn. 3.7:
]P
1))(κp(p[TT A0
0max
(3.7)
where Tmax is the maximum bubble collapse temperature; T0 is the ambient temperature
of the liquid, p0 is the hydrostatic pressure; pA is the applied acoustical pressure; and P is
the pressure in the bubble at its maximum size, which is set to the vapor pressure of the
liquid. The lower collapse temperature of the bubble produces comparatively less •OH.
Therefore, we do not recommend the use of FA as bulk •OH scavenger in cavitational
systems.
100
3.4.2.2 Carbonic acid
CA has not been reported as a •OH scavenger. Our interest in CA was primarily to
explore the effect of CO2 on the cavitation bubbles, rather than as an •OH scavenger.
Based on FA results, we suspected that H2CO3 would significantly affect cavitation
bubbles mostly due to its equilibrium with gaseous CO2 at pH 3.5 in aqueous solution.
H2CO3 ⇌ H2O + CO2 (3.8)
Figure 3.3 shows that CA significantly affects cavitation bubbles, resulting in the lowest
pulse enhancement -20.9% among the scavengers tested. Gaseous CO2 is expected to
enter stable microbubbles during the interval between the pulses, reducing the bubble
collapse temperature, resulting in less reaction of CBZ under PW ultrasound. In addition,
Henglein observed that the main product of the sonolysis of CO2 is CO [31], which is
also capable of reducing the maximum bubble collapse temperature, Tmax, (eqn. 3.7) due
to a decrease in the polytropic index (κAir =1.405 and κCO =1.380). Further, CO
formation is attributed to thermal decomposition of CO2 [31].
CO2 CO + •O (3.9)
Thus, CO2 reduces the temperature of collapsing cavitation bubbles, resulting in a
negative impact on •OH formation and contaminant degradation.
3.4.2.3 Terephthalic acid/terephthalate
TA has been widely used as an •OH scavenger in sonochemistry, radiochemistry,
and photochemistry due to its high sensitivity to trap •OH to form the fluorescent product,
hydroxyterephthalate [32-35]. In addition, in neutral and basic solution, the anionic form
of TA, terephthalate (TPA), dominates (pKa,1=3.52 and pKa,2=4.46). Thus, the polar
101
anionic TPA is assumed to reside in bulk solution and not partition to the interface or
vapor region of cavitation bubbles [10-11]. At pH 3.5, the PE of 1 mM TA was -10.0%,
suggesting that TA interacts with bubble-water interface. As TA is not in its anionic form
at pH 3.5, we retested at pH 7.4. However, the PE remained negative (-5.9%). Unlike FA
and CA which enter into microbubbles, we suspect that TA competes for adsorption sites
during the silent cycle, and accumulates at the interface of oscillating and collapsing gas
bubbles, causing reduced degradation of CBZ in PW as compared to CW. As a result, a
negative PE was observed.
This result is in agreement with our previous work investigating the sonolysis of
ibuprofen and ciprofloxacin in the presence of TPA [36]. To understand these results we
calculated the hydrophobic enrichment factor, Cinterface/Caq, following the method from
Tauber et al. [14]. Tauber et al., used this concept to study the enrichment of 4-
nitrophenol at the interfacial region and concluded that 4-nitrophenol, with Cinterface/Caq
≈80, primarily degraded in the gas-liquid boundary layer. Cinterface/Caq for TA was
calculated to be 7.7 at pH 3.5, indicating that TA readily partitions to the bubble-water
interface. Thus, we do not recommend the use of TA or TPA as bulk •OH scavenger in
cavitational systems.
3.4.2.4 Potassium iodide
KI is one of the most popular •OH scavengers [12, 14-15]. Figure 3.3 shows a PE of
-11.5% for CBZ when KI is present. The reduced PE of CBZ with KI compared to in the
absence of KI suggests that KI interacts with the bubble-water interface, resulting in a
reduced degradation rate of CBZ in PW mode compared to CW mode.
102
Although iodide is a non-volatile solute, the sonication of a solution of KI liberates
considerable iodine [37], and, in fact, is the reported pathway for quantifying the reaction
of KI with •OH:
I- + •OH I• + OH
- (3.10)
I• + I• I2 (3.11)
I2 + I- I3
- (3.12)
Kotronarou [37] observed that the rate of I3- formation in an aqueous solution decreased
by 1/3 when the ports of the sonication reactor were left open compared to closed
conditions, suggesting substantial iodine degassing. Although no studies directly report
iodine entering into the gas phase of cavitation bubbles, I2 is volatile with a reported
Henry’s law constant of 0.0245 atm-m3/ mol [38]. For comparison purposes, Drijvers et
al., [19] investigated the sonolysis of chlorobenzene with a Henry’s law constant of
0.00311 atm-m3/mol in aqueous solution. They concluded that chlorobenzene penetrated
the gas/liquid boundary layer, lowering the specific heat ratio of the gas inside the
cavitation bubbles, reducing the bubble collapse temperature, and ultimately reducing
degradation of chlorobenzene by ultrasound. The Henry’s law constant of iodine is nearly
one order of magnitude greater than that of chlorobenzene, suggesting that the
ultrasound-induced byproduct, iodine, partitions to the gas phase of the cavitation bubble.
In addition, Liu et al., [39] revealed that the hydrogen bonding network at the air/water
interface is disturbed by the addition of iodide due to its large size and high polarizability,
suggesting iodide itself perturbs the bubble-water interface. Therefore, I- itself and its
reaction product I2 interact with cavitation bubbles; we do not recommend the use of KI
as bulk •OH scavenger in cavitational systems.
103
3.4.2.5 Benzenesulfonate
Although not a commonly reported •OH scavenger, BS was considered in our study
since it is hydrophilic (apparent log Kow=-2.93 at pH 3.5) [40] due to its sulfonate
functional group, and can sensitively trap •OH (4.7× 109 M
-1 s
-1) [26]. Thus, we expected
BS to effectively scavenge •OH in bulk solution. In Figure 3.3, the PE in the presence of
1 mM BS (0.93%) is statistically less than in its absence (5.8%) (p<0.01). This reduction
in PE with BS suggests that BS affects cavitation bubble collapse. Measurements of the
Gibbs surface tension of BS showed that the surface excess remained unaltered until the
concentration of BS reached 0.1 M (Figure 3.4). The Gibbs surface excess of a substance,
which has been directly correlated to the intensity of sonoluminescence, is a measure of
enrichment of the substance at the bubble-water interface [41-42]. Although Gibbs
surface excess changes were not detected at low concentration (<0.1 M), the hydrophobic
enrichment factor, Cinterface/Caq, was calculated to be 40.7 for BS. Because of the lower PE
of CBZ in the presence of BS and the high calculated enrichment factor, we do not
recommend the use of BS as a bulk •OH scavenger in cavitational systems.
3.4.2.6. Methanesulfonate
MS shows a greater pulse enhancement (PE=11.3%), as compared to that in the
absence of MS (5.8%) (p<0.01). In addition, the presence of MS promoted, rather than
inhibited the degradation of CBZ. Enhanced degradation of CBZ with MS may be
attributed to reactive secondary reaction species [43]. Therefore, because of the enhanced
degradation of CBZ with MS, it is not appropriate for use as an •OH scavenger.
104
3.4.2.7 Acetic acid
The PE of CBZ in the presence of AA was 4.6 %, as compared to 5.8 % in the
absence of the scavenger (Figure 3.3). Unlike the other scavengers tested, the lack of
statistical difference in PE values (p<0.01) suggest that AA and its by-products neither
perturb the adsorption sites of CBZ at the bubble-water interface, nor partition into gas
microbubbles during the silent cycles in PW mode, but rather reside in bulk solution.
To confirm that AA does not partition to the bubble-water interface of cavitation
bubbles, surface tension measurements and corresponding Gibbs surface excess
calculations were performed for the AA-water binary system at room temperature.
Results indicate that AA does not accumulate on gas-water interfaces until the AA
concentration reaches 140 mM (Figure 3.4). This observation is similar to multi-bubble
sonoluminescence of acetate [44]. The reported sonoluminescence intensity of acetate
remained relatively constant until the concentration reached 100 mM, indicating that
acetate does not accumulate on the bubble surface at concentrations lower than 100 mM;
thus, no sonoluminescence quenching was observed. The similar point at which acetic
acid (140 mM) and acetate (100 mM) appear to accumulate on interfaces suggests that
AA and acetate behave similarly in the ultrasonic field.
In addition to surface accumulation, AA has the potential to enter the gas phase of
cavitation bubbles. Henry’s law predicts the distribution of AA between aqueous solution
and the gas at equilibrium. Nanzai et al., investigated the sonolysis of twelve organic
compounds with Henry’s law constant (KH) ranging from 7.34 × 10-9
to 1.05 × 10-2
atm-
m3/mol. Their results showed that the effect of the Henry’s law constants on the
105
degradation rates became pronounced at KH above 2.40 × 10-5
atm-m3/mol and the
correlation between degradation and KH was observed only at high KH values [45]. Chiha
et al., examined sonochemical degradation of phenol (KH=3.33× 10-7
atm-m3/mol), 4-
isopropylphenol (KH=1.09× 10-6
atm-m3/mol), and Rhodamine B (KH=2.20× 10
-21 atm-
m3/mol) in aqueous solutions [46]. They claimed that these compounds are not degraded
inside the cavitation bubbles because of their low Henry’s law constants. Based on these
studies of the effects of Henry’s law constants on sonolysis of contaminants, AA, with a
Henry’s law constant 1.00 × 10-7
atm-m3/mol [38], does not appear likely to diffuse to the
gas phase of cavitation bubbles to a large degree.
Calculations were also conducted to validate our conclusion that AA will not
migrate to bubble interfaces or interiors. Based on the method from Tauber et al., [14] the
hydrophobic enrichment factors, Cinterface/Caq and Caq/Cg, for acetic acid were calculated
to be 0.8 and 8.13×104, respectively. Both factors indicate that AA preferentially stays in
bulk solution.
While our macroscopic observations and calculations do not provide molecular level
information regarding AA on the bubble-water interface, Johnson et al., studied the AA
[47] and acetate [48] molecular orientation and speciation at the air interface by
vibrational sum frequency spectroscopy. They found that the structure of the interface is
disrupted in the presence of 0.3 mole % (0.165 M) AA with AA covering 7% of the
interface [47], whereas acetate was not observed to disrupt the interface [48]. Although
they did not investigate lower AA concentration, their work is consistent with our surface
tension measurement; thus we expect the observed perturbation of the interface will
disappear at concentration below 0.1 M.
106
The kinetics and mechanism of the sonolysis of acetate solution was studied by
Gutierrez et al. [8]. A large concentration (~0.1 M for acetate and ~0.01 M for acetic acid)
was required to suppress the overall formation of 50 % H2O2. However, based on product
formation yields, half of •OH in bulk solution was scavenged at an acetate concentration
of 1.2 M, four to five orders of magnitude less than the bulk acetate concentration.
Based on the low CO to H2 product formation ratio at [acetate]0 = 0.1 M, they stated that
the sonolysis of acetate (≤ 0.1 M) occurs via decomposition of water, attack of acetate by
•OH, and the reaction of the secondary radicals with the oxygen. The trace amount of CO
formed at 0.1 M acetate may be due to sputtering of acetate in bulk solution into the gas
bubbles during sonolysis [12, 49]. They concluded that the amount of •OH in bulk
solution is small, and acetate is inert toward the interfacial and gas region.
Figure 3.2 depicts the effectiveness of AA scavenging for •OH on sonochemical
degradation of CBZ. In the absence of AA, the degradation rates of CBZ were 0.519 and
0.549 µM min-1
under CW and PW, respectively. In the presence of 1 mM AA the
observed rate was 0.362 and 0.379 µM min-1
under CW and PW, respectively,
representing a ~30 % inhibition in the presence of a bulk •OH scavenger in both modes of
ultrasound. The result indicates that •OH in bulk solution is responsible for ~30 % of
CBZ degradation, being consistent with our assumption that the interfacial region of the
cavitation bubbles are the dominant location responsible for CBZ degradation.
AA/acetate has been investigated in sonochemical system previously as a pollutant
and a co-existing pollutant. [12, 18, 50-51]. Findik and Gunduz investigated the
degradation of AA in the presence of NaCl to examine the salting-out effect on sonolysis
[51]. The presence of NaCl exhibited a positive enhancement on AA degradation until its
107
concentrations reached 0.75 M, suggesting that AA favors bulk solution, thus, a high
NaCl concentration is needed to push AA from bulk solution to the bubble interface. For
comparison purposes, Seymour and Gupta examined the salting-out effect on the
sonolysis of chlorobenzene, a surface active pollutant with major degradation occurring
in the gas and interfacial region of cavitation bubbles [52]. In their study, the degradation
efficiency is enhanced by 10% with 0.17 M NaCl as compared to in its absence. Tiehm
and Neis [18] examined sonochemical degradation of chlorophenol in the presence of
acetate and observed that the degradation rate was reduced slightly less than it was in the
presence of glucose. They concluded that hydrophilic acetate or glucose did not interfere
with chlorophenol, which was mainly degraded in the bubble-water interface. These
studies confirm that AA preferentially stays in bulk solution and inertially diffuses to
bubble surfaces.
3.4.3 Robustness of AA/acetate as bulk solution •OH scavenger
In order to determine the range of conditions in which AA/acetate acts as a bulk
•OH scavenger, its effect on CBZ degradation was tested under different conditions. Our
goals were to explore the concentration and pH conditions at which AA/acetate
adequately scavenges bulk •OH but does not affect cavitation bubbles. First, the sonolysis
of the probe compound, CBZ, was conducted at pH 3.5 with an AA concentration
ranging from 0.5 mM to 100 mM.
Figure 3.5 illustrates that with different concentrations of AA, the reported initial
degradation rates of CBZ remain relatively constant in both CW and PW mode
ultrasound. At high concentration of AA/acetate, AA/acetate has been shown to interact
108
with cavitation bubbles, while at very low concentration AA/acetate may not adequately
scavenge •OH in bulk solution. The degradation rate of CBZ in the presence of AA was
predicted by multiplying the degradation rate of the target compound in the absence of
AA by the quenching ratio, Q, the ratio of reaction occurring by •OH between competing
solutes. Q of AA of the degradation of CBZ was predicted from eqn. 3.13:
100%CkCk
CkQ
CBZCBZAAAA
CBZCBZ (3.13)
where kCBZ and kAA are the second order rate constants of •OH reacting with CBZ and
AA, respectively, and CCBZ and CAA are the concentrations of CBZ and AA, respectively.
The predicted initial degradation rates monotonously decrease from 0.35 to 0.0048
µM min-1
under either CW or PW ultrasound. The significantly higher and relatively
stable observed initial rates of CBZ degradation compared to predicted initial rates
suggests that CBZ diffuses to the interfacial region of cavitation bubbles, regions of high
temperature and •OH concentration. The lack of effect of AA within a wide concentration
range indicates that AA does not affect the interfacial and gas region of cavitation
bubbles. Further, because adding more of the •OH scavenger does not alter the initial
degradation rate of CBZ, the lowest concentration of AA appears to adequately scavenge
the bulk •OH. Only a small amount of •OH diffuses to bulk solution during sonication;
thus, the presence of small amount of AA sufficiently outcompetes CBZ, scavenging the
•OH in bulk solution. An increase in the AA concentration has little impact on the
observed initial degradation rates of CBZ for the concentration of CBZ used.
Generalizing to other systems, the lowest concentration of AA may not be appropriate if
a high compound concentration will be used and AA is used as a bulk •OH scavenger.
109
Next, we tested the robustness of AA/acetate within a range of pH values. Sonolysis
of 10 M CBZ was conducted in the presence of 1 mM AA/acetate from pH 3.5 to 8.9.
Figure 3.6 shows that the PE of CBZ remains constant through the entire pH range tested.
Henglein and Kormann [12], suggested that both AA and acetate have similar scavenging
behaviors based on the concentration needed to inhibit formation of 50% H2O2. Our
results are consistent with their work, suggesting that acetate behaves similarly to AA.
3.5 Conclusions
The effects of the •OH scavengers, FA, CA, TA/TPA, KI, MS, BS, and AA/acetate,
on cavitation bubbles have been examined with the sonolysis of the probe compound,
CBZ, under CW and PW ultrasound. PW ultrasound provides a technique to determine
whether scavengers affect the interfacial and gas region of cavitation bubbles. All the
scavengers, except for MS, exhibited inhibitory effects on CBZ sonolysis. Based on the
reduction in PE of CBZ, all the scavengers except AA/acetate affect cavitation bubbles
and are not recommended to use as bulk •OH scavengers in cavitational systems.
We recommend acetic acid/acetate as a bulk •OH scavenger. It efficiently quenches
•OH in bulk solution, resulting in a reduced degradation rate of CBZ. The range of utility
of AA/acetate as a bulk •OH scavenger was tested under different concentrations and pH
values. The observed degradation rates of CBZ within the range of concentrations and pH
values tested were unaltered. Evidence suggests that at concentrations below 0.1 M and
all pH values tested, AA/acetate is an ideal bulk •OH scavenger in sonochemical systems.
Acknowledgments
110
Financial support provided by Ohio Sea Grant College Program is gratefully
acknowledged. R. X. gratefully acknowledges Dr. Chinmin Cheng and Matthew Noerpel
for the valuable comments on this manuscript.
111
None FA CA TA KI BS MS AA0.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
d[C
BZ
]/d
t|0 (
M/m
in)
PW ultrasound
CW ultrasound
Figure 3.1: Initial degradation rates of 10 M CBZ in the absence and presence of 1 mM
•OH scavenger candidates under PW and CW ultrasound at pH 3.5
112
FA CA TA KI BS MS AA
-20
-10
0
10
20
30
40
50
60
70
80
PW ultrasound
CW ultrasound
Inh
ibit
ion
(%
)
Figure 3.2: Inhibitory effects of each bulk •OH scavenger candidate on sonolysis of 10
M CBZ under CW and PW ultrasound at pH 3.5.
113
None FA CA TA TPA KI BS MS AA
-40
-30
-20
-10
0
10
20
30
Pu
lsed
en
han
cem
en
t (%
)
Figure 3.3: PE of 10 M CBZ in the absence and presence of the various bulk •OH
scavenger candidates at pH 3.5 (TPA was tested at pH 7.4).
114
-10 -8 -6 -4 -2 0 2 420
30
40
50
60
70
80 AA
BS
Su
rface t
en
sio
n (
mN
/m)
ln[concentration] (M)
0
2
4
6
8
10
12
AA
BS
Su
rface e
xcess (
mo
lecu
les/n
m2)
Figure 3.4: Gibbs surface tension ( ) and surface excess ( ) for acetic acid (AA) and
benzenesulfonate (BS) as a function of concentration at room temperature.
115
1E-3 0.01 0.10.0
0.1
0.2
0.3
0.4
0.5
d[C
BZ
]/d
t|0 (
M/m
in)
[AA] (M)
CW observed
CW predicted
PW observed
PW predicted
Figure 3.5: Observed and predicted initial degradation rates of 10 M CBZ at pH 3.5
under CW and PW ultrasound as a function of acetic acid concentrations ranging from
0.5 mM to 100 mM.
116
pH 3.5 pH 5.7 pH 7.4 pH 8.9
-10.0
-7.5
-5.0
-2.5
0.0
2.5
5.0
7.5
10.0
12.5
Pu
lse
d e
nh
an
ce
me
nt
(%)
Figure 3.6: Pulse enhancement of 10 M CBZ in the presence of 1 mM bulk phase •OH
scavenger AA/acetate as a function of pH.
117
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Chapter 4: Degradation of Pharmaceuticals and Personal Care Products (PPCPs) in
Aqueous Solution Using Pulsed Wave Ultrasound
4.1 Abstract
In this study, continuous wave (CW) and pulsed wave (PW) ultrasound at 205 kHz
was used to degrade five pharmaceuticals (carbamazepine, ibuprofen, acetaminophen,
sulfamethoxazole and ciprofloxacin), and two personal care products (propyl gallate and
diethyl phthalate). These seven compounds, covering a range of physicochemical
properties and a diversity of structures, are degraded to different degrees by ultrasound.
Acetaminophen is the most recalcitrant compound to ultrasound with degradation rates of
0.27 and 0.44 µM/min under CW and PW ultrasound, respectively, whereas, the
degradation rates of propyl gallate are almost one order of magnitude faster. Degradation
rates by PW ultrasound are compound dependent with degradation faster for smaller
compounds or slower for larger compounds than that under CW ultrasound.
maller PPCP compounds with molar volumes less
than 130 mL/mol are able to more readily diffuse to bubble interfaces and are impacted
most by pulsing ultrasound.
125
4.2 Introduction
Pharmaceuticals and personal care products (PPCPs) are a large and diverse class of
compounds, consisting of prescription and over-the-counter therapeutic drugs, veterinary
drugs, fragrances, and cosmetics. According to the National Center for Health Statistics,
annual pharmaceutical sales in North America was $299.6 billion in the year 2008
(Sebelius et al., 2010). Unavoidably, a large fraction of these PPCPs are disposed of or
discharged into the environment on a continual basis through various pathways, including
domestic sewage, landfills, and wet weather runoff (Daughton and Ternes, 1999),
resulting in both aquatic and terrestrial biota and human exposure to these compounds.
The U.S. Geological Survey provided the first nationwide reconnaissance of the
occurrence of PPCPs in water resources during 1999-2000 and found PPCPs in 80% of
the streams sampled (Kolpin et al., 2002). Moreover, studies have shown that certain
PPCPs, such as estrogenic compounds, persist in the environment and have a very high
bioaccumulation potential (Ternes et al., 2004; Topp et al., 2008).
In addition to changing disposal practices of PPCPs, a variety of water treatment
technologies have been investigated and developed to reduce the concentrations of
PPCPs in water (Oulton et al., 2010). Among these treatment technologies, ultrasound is
an emerging option with advantages that include ease of use, lack of chemical addition,
and ability to degrade contaminants with a wide range of physicochemical properties
(Adewuyi, 2005a; b; Weavers et al., 2000).
Ultrasound decomposes contaminants through thermal degradation and/or radical
oxidation mechanisms. When continuous or pulsed ultrasound waves propagate through
water, cavitation bubbles form. These bubbles oscillate with the rarefaction and
126
compression phases of the ultrasound wave. Once the bubbles reach a critical size, the
bubbles implode resulting in an enormous concentration of energy at localized hot spots
in the solution (Gutierrez et al., 1986; Suslick, 1990). These hot spots not only provide
sufficient heat to dissociate chemical bonds in contaminants (i.e., thermolysis) (Currell et
al., 1963; Song and O'Shea, 2007), they also dissociate water, forming hydroxyl radical
(•OH) and initiating oxidation reactions in water (Kotronarou, 1992; Petrier et al., 1994).
Applying ultrasound to remove PPCPs from drinking water and wastewater has
drawn increasing interest. However, predicting removal efficiencies for PPCPs,
especially pharmaceuticals, is challenging. Nanzai et al., (2008) concluded that a more
hydrophobic monocyclic aromatic compound undergoes a faster degradation due to a
higher degree of accumulation on bubble-water interface. Fu et al., (2007) stated that
smaller compounds diffuse faster in bulk solution to the interfacial/gas region of
cavitation bubble, resulting in a faster degradation rate.
Many studies have shown that at proper settings, the degradation of contaminants
under pulsed wave (PW) ultrasound is more effective than under continuous wave (CW)
ultrasound (Ciaravino et al., 1981; Neppolian et al., 2009). For instance, pulsing
ultrasound at 100 ms on and 100 ms off, the degradation rate of 0.1 mM sodium 4-
octylbenzene sulfonate (OBS), was approximately 30 % faster than that under CW (Yang
et al., 2005). Xiao et al., (2012) observed that the removal efficiency of 10 μM
carbamazepine (CBZ) under PW ultrasound was greater than CW by about 6 %.
The mechanisms for the enhanced sonochemical reactivity under PW ultrasound
remain unclear though. Yang et al., (2005) and Xiao et al., (2012) attributed the enhanced
degradation by PW ultrasound to increased contaminant accumulation on the surface of
127
the bubbles during the intervals between pulses, causing more molecules to react with the
cavitation bubbles. However, Ciaravino et al., (1981) attributed the increased formation
of iodine in PW mode to the survival of unstabilized nuclei during the off-time, resulting
in more violent collapse in the subsequent pulse. Neppolian et al., (2009) attributed the
increased sonochemical oxidation of arsenic (III) by PW ultrasound to the “silent”
oxidation reactions and an increased number of active cavitation bubbles.
In this study, we investigated the effect of pulsing on the degradation of PPCPs. Five
pharmaceuticals (carbamazepine (CBZ), ibuprofen (IBU), acetaminophen (ATP),
sulfamethoxazole (SFT) and ciprofloxacin (CIPRO)), and two personal care products
(propyl gallate (PG) and diethyl phthalate (DP)), were selected on the basis of
environmental relevance (Daughton and Ternes, 1999; Kolpin et al., 2002) and diversity
in physicochemical properties and structures (Table 4.1). Many PPCPs are hydrophobic
and surface active in water; thus, PW ultrasound exhibits potential to be a better method
to remove PPCPs, as compared to CW ultrasound. Further, we linked the importance of
physicochemical properties of a compound to degradation under pulsed ultrasound.
4.3 Experimental Methods
CBZ, IBU, ATP, and PG all 99% from Sigma-Aldrich, SFT (99%), DP (99%) and
CIPRO (98%) from Fluka, and sodium phosphate and sodium acetate both 99% from
Fisher-Scientific, were used as received. Table 4.1 lists physicochemical properties for
these PPCPs. Stock solutions of these chemicals were prepared using Milli-Q purified
water (Millipore, MA) with the resistivity R = 18.2 M Ω cm. A 10 µM initial
concentration was chosen to be close to an environmentally relevant concentration yet
128
concentrated enough to be analyzed. Due to its low •OH reactivity, solutions were
buffered by 1 mM NaH2PO4 at pH 3.5. A pH value of 3.5 was selected because the
apparent and true Kow values for the PPCPs are equivalent at this pH value, recognizing
that hydrophobicity changes dramatically as pH varies due to ion-pair partitioning
(Avdeef et al., 1998). In select experiments, 1 mM acetic acid was added to 10 μM PPCP
solutions to scavenge bulk •OH (Xiao et al., 2012).
Ultrasound at 205 kHz was emitted from a USW 51-52 ultrasonic flat plate
transducer (A = 23.4 cm2) (ELAC Nautik, Inc., Kiel, Germany) into a glass vessel with a
volume of 300 ml. The reactor was equipped with a cooling water jacket (Isotemp 1006S,
Fisher Scientific) to maintain the temperature of solutions at 20 °C. A SM-1020
Function/Pulse generator (Signametrics Corp., Seattle, WA) was used to deliver a pulse
of 100 ms on and 100 ms off. The selection of PW ultrasound settings was based on
previous studies which showed enhanced degradation of surfactants by PW compared to
CW (Yang et al., 2006). A linear amplifier (T&C Power Conversion, Inc., Rochester, NY)
magnified the generated electrical signal to drive the transducer in either CW or PW
ultrasound. A digitizing oscilloscope (54501, Hewlett-Packard) was used to detect the
pulse signal received by the transducer. The acoustic energy density to the reactor,
determined by calorimetry, was measured to be 45 W/L. Throughout the experimental
runs, calorimetry and the pulsed signal were monitored periodically to confirm system
functionality.
0.5 mL samples were taken from the reactor at designated times using a 1 mL glass
syringe (Gastight 1001, Hamilton Corp.). The sample volume taken during the course of
sonication did not exceed 1 % of the total volume in order to keep the distribution of
129
acoustic field in the reactor consistent. All the experiments were carried out in, at least,
duplicate.
A Hewlett-Packard 1100 high performance liquid chromatograph (HPLC) equipped
with a diode array detector (DAD) and a 5 µm, 150 × 2.1 mm SB-C18 column (883700-
922, Agilent Technologies) was used to analyze the concentration following sonolysis of
PPCPs. 0.5 mL/min eluent consisted of 20 mM pH 3 phosphoric acid buffer and
acetonitrile (HPLC grade, Fisher Scientific). Eluent ratios and the injection volumes
varied depending on the compound analyzed. The UV wavelengths were set at 220, 223,
215, 214, 274, 225, and 230 nm for CBZ, IBU, ATP, SFT, CIPRO, PG, and DP,
respectively.
Gibbs surface tension ( ) were measured for each PPCP solution by a McVan
Analite Surface Tension meter (2141, McVan Instruments) with a glass Wilhelmy plate.
The measurements were carried out in triplicate at room temperature.
A non empirical calculation at Hartree Fock (HF) level of theory with 6-31+G*
basis set was applied to determine the molar volume of the molecules (mL/mol) in water,
using the polarizable continuum model solvation method. HF/6-31+G* is a reliable level
of theory to calculate the geometry of large organic molecules (Cummins and Gready,
1989; Schlegel, 1982). All molecules were drawn using GaussView 4.0.8, and
calculations were performed using Gaussian 03 at the Ohio Supercomputer Center (OSC).
Statistical analysis, including the Bivariate Pearson two-tailed correlation test and
the t-test, were performed using SPSS 13.0 (LEAD Technologies, Inc).
4.4 Results and Discussion
130
4.4.1 Sonochemical degradation kinetics
The sonochemical degradation of CBZ, SFT, IBU, ATP, CIPRO, PG, and DP is
shown in Figure 4.1. The initial degradation rates of the seven PPCPs varied. For
example, the initial degradation rates of PG under CW and PW ultrasound are 2.98 and
2.90 µM /min, respectively, approximate one order of magnitude faster than the other
compounds. ATP degrades the slowest by CW ultrasound, 0.27 µM /min.
Strong correlations have been reported between sonochemical degradation rates of
contaminants and their physicochemical properties (Colussi et al., 1999; Nanzai et al.,
2008; Shemer and Narkis, 2005). For instance, Fu et al., (2007) observed a linear
correlation between the rate of sonolysis of eight different estrogen compounds and their
molecular weight. Shemer and Narkis (2005) observed, for a series of trihalomethanes
(THMs), a linear correlation between the vapor pressure and the sonochemical
degradation rate. Similarly, Colussi et al., (1999) found that at various frequencies the
sonolytic degradation rate of chlorinated hydrocarbons, including CCl4, CHCl3, C2Cl6,
and CH2Cl2, was faster for compounds with larger Henry’s law constants. Nanzai et al.,
(2008) reported that the initial degradation rate of twelve monocyclic aromatic
compounds linearly correlated to their log Kow values.
The dependence of degradation rates of the PPCPs under both CW and PW
ultrasound on physicochemical properties, including molecular weight, molar volume,
vapor pressure, log Kow, and Henry’s law constant, was evaluated by non-parametric
correlations (Spearman’s rho) to analyze for any possible relationships. We did not
observe any correlation between degradation rate and any single property (p>0.05) under
either CW or PW ultrasound (Figure 4.6 in the Supporting Information). The reported
131
correlations in previous studies may be attributed to the fact that the compounds
investigated lie within classes of compounds with similar structure, thereby simplifying
the correlations. Previous correlations have been made using equilibrium and kinetic
parameters. Sonochemical degradation involves thermolysis and •OH oxidation of PPCPs
and their byproducts in gas, interfacial, and bulk regions of cavitation bubbles. In this
heterogeneous system, a single physicochemical property of the parent PPCP has limited
capability to accurately govern the complex kinetics, especially when the investigated
PPCPs cover a wide range of physicochemical properties and diversity of structures.
The results may have significant implications for the ultrasonic treatment of PPCPs
in wastewaters and drinking waters. Contaminants, to be removed in municipal water and
wastewater treatment plants, consist of a wide variety of classes of contaminants. Hence,
the physicochemical properties of target contaminants may not be the principal
consideration in determining the degradation of a contaminant by ultrasound. Unlike
other treatment technologies, such as sedimentation or activated carbon adsorption,
whose treatment depends on a specific physicochemical property (Halden and Heidler,
2008), our results suggest that one physicochemical property does not predict treatment
by ultrasound. Moreover ultrasound can remove PPCPs covering a wide range of
physicochemical properties.
4.4.2 Pulse enhancement
Pulse Enhancement (PE), defined as the difference in initial degradation rates
between PW and CW over that of CW (eqn. 4.1), is a measure used to compare the
difference in the rate of degradation between CW and PW ultrasound
132
％ 100
dt
Cd
dt
Cd
dt
Cd
(%) PE
CW
CWPW (4.1)
where CW
dt
Cd and
PWdt
Cdare initial degradation rate of the PPCPs under CW and
PW ultrasound, respectively. Accordingly, the PE is positive, negative or zero. If the PE
is positive, PW ultrasound is more efficient than CW and vice versa.
Figure 4.2 shows the PE for the seven PPCPs. For ATP, IBU, and CBZ, pulsing
caused a statistically significant increase (p<0.01) as compared to CW ultrasound. For
instance, ATP under PW is 1.6 times faster than that under CW. However, for PG, SFT,
DP, and CIPRO, PW ultrasound is slower than that under CW mode by 2.45 %, 5.89 %,
5.87 % and 15.34 %, respectively.
Several studies investigated and compared the degradation of compounds under CW
and PW ultrasound (Clarke and Hill, 1970; Deojay et al., 2011; Neppolian et al., 2009;
Yang et al., 2005). Clarke and Hill (1970) observed a 1.5 fold increase in iodine
production from KI solution under PW ultrasound as compared to CW ultrasound.
Bubble size reduction below the bubble resonant size during the silent cycle was
attributed to the enhancement by pulsing; in the subsequent ultrasonic pulse, this batch of
bubbles and pre-existing nuclei in solution grow to resonant size. A 1.3 fold enhancement
for the oxidation of arsenic (III) to arsenic (V) by PW ultrasound, compared to CW, was
attributed to an increased number of active cavitation bubbles and “silent” oxidation
reactions initiated by •OH (Neppolian et al., 2009). Both studies imply that pulsing will
133
be beneficial for all PPCPs; however, we observed a negative PE for PG, SFT, DP, and
CIPRO.
Yang et al. (2005) investigated the sonolysis of a surfactant, OBS, and non-
surfactant, 4-ethylbenzene sulfonic acid (EBS). PW ultrasound was better than CW
ultrasound with a PE of 94 %; however, the increase in degradation rate of PW over CW
was not statistically significant for EBS. Enhancement was attributed to the accumulation
of the surfactant on cavitation bubble surfaces, resulting in reduced surface tension and
lowering the cavitation threshold (Yang et al., 2005). In addition, surfactants inhibit
bubble dissolution during the off cycle, resulting in a larger active cavitation bubble
population and increased absorption sites (Fyrillas and Szeri, 1995). Their mechanism is
specific to surfactants; the PPCPs in our study did not exhibit any alteration in surface
tension at 10 M (data not shown).
While previous explanations of the enhancement of PW ultrasound are not
applicable to our results, studies conducted by Yang et al., (2005; 2006) and others (De
Visscher, 2003; DeVisscher et al., 1996; Drijvers et al., 2000; Fu et al., 2007), showed
that mass transport of target compounds to the bubble surface is controlled by diffusion.
For instance, Fu et al., (2007) correlated the degradation rate constants of seven estrogen
compounds to MW and stated that small sized compounds diffuse faster to the vicinity of
cavitation bubbles, resulting in a faster degradation rate. Drijvers et al., (2001) found that
the monohalogenated benzene concentration ratios between the vicinity of cavitation
bubbles and the bulk solution strongly correlated to their diffusion coefficients rather than
their Henry’s law constants. Based on their studies, we hypothesized that, during the
silent cycle, different size of PPCPs may diffuse to and accumulate on cavitation bubbles
134
to different extents, resulting in different PE values. Thus, the molecular size would play
a role in this diffusion-controlled process.
To test the hypothesis, a correlation analysis was conducted between PE and
physicochemical properties listed in Table 4.1 (Figure 4.7 in the Supporting Information).
Correlations were not observed. However, diffusion governs the transport of PPCPs
based on the diffusivity equation of Hayduk and Laudie (1974) (eqn. 4.2), diffusion
coefficients (Diw, cm2/s) in water are correlated to the molar volume (Vi, in the unit of
mL/mol) in of a compound.
0.589
i
1.14
w
5
iw
Vμ
1013.26D (4.2)
where w is the viscosity of water at 25 °C. Vi can be approximated by molecular weight
when documented values and computational resource are not available.
Thus, PE was plotted against molar volume (Figure 4.3a) and molecular weight
(Figure 4.3b) of the PPCPs, respectively. Figure 4.3 shows that small sized PPCPs exhibit
a positive PE, while large sized PPCPs exhibit a negative PE. The results suggest that
diffusion to the cavitation bubble surface is an important mechanism, resulting in
enhanced degradation of the PPCPs by pulsing in our study.
Unlike Figure 4.3a and b, the degradation rate of the PPCPs under CW and PW
ultrasound were not correlated to either their molar volumes or molecular weight (Figure
4.6a and b in the Supporting Information). The observed degradation rate of the PPCPs
by sonolysis involves both high temperature decomposition and •OH oxidation in the
vicinity of cavitation bubbles and the bulk solution, as indicated by eqn. 4.3.
135
bubbleinterfacebulkobs dt
dC
dt
dC
dt
dC
dt
dC
(4.3)
where dt
dC
represents the degradation rate of a PPCP, occurring in bulk solution (bulk),
occurring at the interface (interface), and occurring in the gas region of cavitation bubble
(bubble), respectively. Clearly, a single physicochemical property of a PPCP is unlikely
to determine the degradation rate.
However, the concept of PE is based on a comparative method (eqn.4.1). The
contribution from bulkdt
dC to
obsdt
dC under CW ultrasound is similar to that under PW
ultrasound (Deojay et al., 2011; Yang et al., 2005), thus there is no contribution of bulk
reaction to PE. In addition, the PPCPs investigated are not likely to degrade inside the
cavitation bubble due to their low Henry’s law constants (Table 4.1), suggesting that a
negligible contribution of bubble reaction to PE. The remaining component,interfacedt
dC,
reaction at bubble-water interfaces, results in the controlling mechanism, diffusion,
becoming more pronounced in our system.
During the on time, ultrasound induces the formation, growth through many wave
cycles, and subsequent collapse of microbubbles. During the off time, microbubbles may:
(1) coalesce and be removed by buoyancy, (2) dissolve and disappear or (3) shrink by
dissolution to size whereby it will actively cavitate during the next on cycle (Leighton,
1994). Also, during the silent cycle (i.e., 100 ms) of a pulse sequence, PPCP molecules
have more time to diffuse to and accumulate on the bubble surfaces. Although the time
scale during the silence cycle is too short for equilibrium partitioning to be reached
136
(Eastoe and Dalton, 2000; Yang et al., 2005), when sonication resumes, the remaining
bubbles and pre-existing nuclei surface populate with more PPCP molecules than before
the silent cycle, grow to resonance size, and collapse. Hence, more PPCP molecules are
in the vicinity of highest temperature and radical concentrations. In this diffusion-
controlled process, small PPCP molecules more quickly diffuse to and accumulate on the
cavitation bubbles during the off time, resulting in a positive PE. Large molecules take
longer to diffuse to the surface, resulting in less accumulation. The competing process of
bubble dissolution results in an overall negative PE for these larger compounds.
As observed in Figure 4.3a, the relationship between PE and molar volume is not
perfect, especially when the compounds have similar molar volumes. For the five
compounds (IBU, PG, DP, CBZ, and SFT) that assemble around 170 mL/mol (Figure
4.3), the hydrophobicity varies significantly. For example, the highest and lowest log Kow
values among these five compounds are IBU (3.70) and SFT (0.89), respectively.
Correspondingly, the highest and lowest PE values among these five compounds are also
IBU (8.61) and SFT (-5.90), respectively, suggesting that the hydrophobicity affects the
PE to some extent. Similar to Figure 4.3a, the relationship between PE and molecular
weight in Figure 4.3b is not perfect, suggesting that the size of PPCPs may not be
perfectly predict the PE value and hydrophobicity play a role in PE. However, although
not studied explicitly, the impact of hydrophobicity on PE appears to be smaller than the
size of PPCPs.
4.4.3 PPCPs degradation in bulk solution
137
Studies have shown a relationship between the aqueous diffusion coefficient of a
compound and the fraction of a compound degraded in and on cavitation bubbles
(DeVisscher et al., 1996; Drijvers et al., 2000). Specifically, small compound fast
diffusing degrade in and on cavitation bubbles whereas large compounds slowly diffusing
degrade primarily in bulk solution. Our interpretation of the trend of increasing PE with
decreasing molar volume is similar. Small PPCP molecules more readily diffuse to and
accumulate on the bubble surfaces during the silent cycle; upon the subsequent ultrasonic
pulse, more PPCP molecules are in the vicinity of high temperature and high [•OH],
resulting in a positive PE.
To validate the role of PPCPs diffusion during the silent cycle, a bulk solution •OH
trapping agent, acetic acid (AA), was irradiated with each compound to differentiate the
contribution of bulk •OH oxidation in the overall degradation under CW ultrasound.
Acetic acid (log Kow= -0.17) in bulk solution reacts with •OH, rather than migrating to
the interface or gas region of cavitation bubbles (Xiao et al., 2012).
The portion of each PPCP degraded in bulk solution was determined from eqn. 4.4.
%100
dt
Cd
dt
Cd
dt
Cd
solution bulk in n degradatio %
AAwithout
AAwith AAwithout (4.4)
where AAwithout
dt
Cd and
AAwith dt
Cdrepresent the degradation rate for each PPCP
without and with 1 mM of the bulk •OH scavenger, acetic acid. As shown in Figure 4.4,
the degradation rates were significantly reduced in the presence of AA for all the
compounds. Based on eqn. 4.4, •OH in bulk solution is responsible for 6.0 % of
138
degradation of PG, the compound least affected by AA. The relatively small reduction in
the degradation rates with AA presence reveals that the majority of the degradation
occurred in and around cavitation bubbles. CIPRO degradation decreased by 66.2 % in
the presence of AA. Unlike the other PPCPs tested, the high decrease in degradation with
AA indicates that the degradation of CIPRO primarily occurred in the bulk solution.
As shown in Figure 4.5a and b, the amount of degradation in bulk solution trends
with increasing molar volume and molecular weight of a PPCP, respectively. The result
indicates that the smaller and hence more mobile compound is more able to reach the
bubble surfaces, the sources of reactivity, resulting in a greater proportion degraded in
non-bulk solution. However, larger PPCP molecules (e.g., CIPRO) have greater steric
hindrance slowing diffusion in solution. Slower diffusion results in a smaller portion of
the compound reaching the bubble surface ending up with a higher fraction degraded in
bulk solution. Therefore, there is an inherent link between the fraction that degraded in
bulk solution and PE, corroborating that molecular size plays a role in the diffusion-
controlled process in the cavitational system.
4.5 Conclusion
The independence of degradation rates of the PPCPs under both CW and PW
ultrasound on their physicochemical properties indicated that there is no single property
that controls the degradation kinetics. In addition, a comparison between CW and PW
modes of ultrasound demonstrates that the enhancement of degradation of PPCPs under
PW ultrasound was compound dependent. For the enhancement due to pulsing, there
exists a relationship between PE and the molar volume and molecular weight, indicating
139
that small size PPCPs exhibit positive PE values, while large PPCPs yield negative PE
values. This relationship implies that diffusion and transport of the PPCPs to the bubble
surface during the silent cycle is key to pulsing. To explore this phenomena, a bulk
solution •OH scavenger, acetic acid, was added in the sonicated system to differentiate
the contribution of bulk •OH oxidation to the overall degradation. It is observed that the
portion of degradation of the PPCPs in bulk solution is associated with their sizes. Small
size molecules may quickly diffuse to the cavitation bubbles, resulting in a higher portion
of the PPCP in and around cavitation bubbles. Large size molecules slowly diffuse to the
surface due to their large size, resulting in a higher portion of these PPCPs molecules
degraded in bulk solution. This trend in bulk degradation is consistent with the PE results.
Our results suggest that PW ultrasound is a method to enhance degradation for the small
PPCPs, while CW ultrasound is more effective to degrade large PPCPs in aqueous
solution.
Acknowledgments
Financial support provided by Ohio Sea Grant College Program is gratefully
acknowledged. R. X. gratefully acknowledges Dr. Chinmin Cheng and Matthew Noerpel
for the valuable comments on this manuscript.
140
Table 4.1: Selected physicochemical properties of CBZ, IBU, ATP, SFT, CIPRO, PG, and DP at 25°C.
1. Molar volume was calculated at Hartree Fock (HF) level of theory with 6-31+G* basis set and PCM solvation method.
2. Estimated from the group method using EPI SuiteTM
v.4.0 (U.S. EPA).
140
141
CBZ SFT IBU ATP CIPRO PG DP0.0
0.5
2.5
3.0
d[C
]/d
t|0 (
M/m
in)
CW ultrasound
PW ultrasound
Figure 4.1: Initial degradation rates of pharmaceuticals (CBZ, SFT, IBU, ATP, and
CIPRO) and personal care products (PG and DP) at initial concentration of 10 µM and
pH 3.5 sonicated under CW and PW ultrasound.
142
-20
0
20
40
60
80
CIPRO DPSFT
CBZ IBU
PE
(%
)
ATP
PG
Figure 4.2: Pulsed enhancements of 10 µM individual PPCP degradation by PW
ultrasound with a pulse length of 100 ms and pulse interval of 100 ms at pH 3.5.
143
120 140 160 180 200 220 240
-20
0
20
40
60
80
100
(a)
Molar volume (mL/mol)
PE
() ATP
IBU
PG
DP
CBZ
SFT
CIPRO
140 160 180 200 220 240 260 280 300 320 340
-20
0
20
40
60
80
100
(b)
DP
PE
()
Molecular weight (g/mol)
ATP
PG
IBU
CBZ
SFT
CIPRO
Figure 4.3: PE was plotted against the molar volumes (a) and molecular weight (b) of the
PPCPs.
144
CBZ SFT IBU ATP CIPRO PG DP0.0
0.5
2.5
3.0d
[C]/
dt|
0 (
M/m
in)
without AA
with AA
Figure 4.4: Initial degradation rates of 10 µM CBZ, IBU, ATP, SFT, CIPRO, PG, and DP
in the absence and presence of 1 mM •OH scavenger AA under CW ultrasound at pH 3.5.
145
120 140 160 180 200 220 2400
10
20
30
40
50
60
70
80
% d
eg
rad
ati
on
in
bu
lk s
olu
tio
n
Molar volume (mL/mol)
DPPG
CIPRO
ATP
IBU
SFT
CBZ
(a)
120 140 160 180 200 220 240 260 280 300 320 3400
10
20
30
40
50
60
70
80 (b)
% d
eg
rad
ati
on
in
bu
lk s
olu
tio
n
Molecular weight (g/mol)
ATP
PGDP
SFT
IBUCBZ
CIPRO
Figure 4.5: The portion of degradation that occurred in the bulk solution under CW
ultrasound is plotted against molar volume (a) and molecular weight (b) of each PPCP
molecule.
146
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151
Supporting Information
152
140 160 180 200 220 240 260 280 300 320 3400.0
0.5
1.0
1.5
2.0
2.5
3.0
DP
IBU
CIPRO
PG
ATP
SFT
CW
PW
d[C
]/d
t|0 (
M/m
in)
Molecular weight (g/mol)
CBZ
(a)
100 120 140 160 180 200 220 240 2600.0
0.5
1.0
1.5
2.0
2.5
3.0 (b) CW
PW
d[C
]/d
t|0 (
M/m
in)
Molar volume (mL/mol)
DP
IBU
CIPRO
PG
ATPSFT
CBZ
-14 -12 -10 -8 -6 -4 -2 00.0
0.5
1.0
1.5
2.0
2.5
3.0 (c) CW
PW
d[C
]/d
t|0 (
M/m
in)
log vapor pressure
DP
IBU
CIPRO
PG
ATP
SFT
CBZ
0 1 2 3 40.0
0.5
1.0
1.5
2.0
2.5
3.0 (d) CW
PW
d[C
]/d
t|0 (
M/m
in)
log Kow
DPIBU
CIPRO
PG
ATP
SFTCBZ
-20 -18 -16 -14 -12 -10 -8 -60.0
0.5
1.0
1.5
2.0
2.5
3.0 (e)
log Henry's law constant
CW
PW
d[C
]/d
t|0 (
M/m
in)
DP
IBU
CIPRO
PG
ATP
SFTCBZ
9.0 9.5 10.0 10.5 11.00.0
0.5
1.0
1.5
2.0
2.5
3.0 (f)
CW
PW
d[C
]/d
t|0 (
M/m
in)
log kOH
ATP
PG
IBUDP
CBZSFT
CIPRO
Figure 4.6: Degradation rates of the PPCPs under CW and PW ultrasound were plotted
against molecular weight (a), molar volume (b), log vapor pressure (c), log Kow (d), log
Henry’s law constant (e) and log •OH rate constant (f).
153
-14 -12 -10 -8 -6 -4 -2 0
-20
0
20
40
60
80
100
log vapor pressure
PE
()
ATP
PG
DPSFT
IBU
CBZ
CIPRO
(a)
0.0 0.5 1.0 1.5 2.0 2.5 3.0 3.5 4.0
-20
0
20
40
60
80
100
(b)
log Kow
PE
()
ATP
PG DPSFT
IBUCBZ
CIPRO
-20 -18 -16 -14 -12 -10 -8 -6
-20
0
20
40
60
80
100
(c)
log Henry's law constant
PE
()
ATP
PG
DPSFT
IBU
CBZ
CIPRO
9.0 9.5 10.0 10.5 11.0
-20
0
20
40
60
80
100
(d)
IBU
SFTDP
CIPRO
ATP
log kOH
PE
()
PG
CBZ
Figure 4.7: PE of the PPCPs were plotted against log vapor pressure (a), log Kow (b), log
Henry’s law constant (c) and log •OH rate constant (d).
154
Chapter 5: Probing the Utility of Pulsed Wave Ultrasound for the Degradation of
Pharmaceuticals in Municipal Wastewater
5.1 Abstract
Both the octanol-water partition coefficient (Kow) and diffusivity (Diw) have been
correlated to the rate of contaminant degradation by ultrasound. However, the relative
importance and a mechanistic understanding of the importance of both properties on
cavitational systems are poorly understood. In this study, a series of six pharmaceuticals
were selected based on their Kow and Diw values to probe the importance of both
properties on pharmaceutical degradation during continuous wave (CW) and pulsed wave
(PW) ultrasound. In deionized water, pharmaceuticals with the highest Diw (i.e.,
fluorouracil (5-FU)) and Kow (i.e., lovastatin (LOVS)) exhibited the greatest enhancement
in degradation rates in PW mode as compared to CW mode. This result suggests that a
pharmaceutical with either high diffusivity or hydrophobicity more readily populates the
bubble-water interface during the silent cycle of PW ultrasound. However, in municipal
wastewater effluent, the pulse enhancement (PE) for 5-FU and LOVS disappeared. The
presence of matrix inorganics (i.e., bicarbonate and sulfate anions) did not affect the PE,
indicating that the wastewater effluent organic matter (EfOM) is the cause of the
disappearance of PE. Irradiating 5-FU and LOVS in hydrophobic (HPO), transphilic
(TPI), and hydrophilic (HPI) fractions of EfOM showed that the TPI fraction reduced the
155
pulse enhancement the most, followed by HPI and HPO fractions. The smaller molecular
weight of TPI suggests that the TPI fraction is able to diffuse to cavitation bubble
surfaces, causing the reduction in pharmaceutical degradation in the presence of EfOM.
5.2 Introduction
Many studies have reported the occurrence of various pharmaceuticals in different
waters around the world, including tap water, wastewater, and receiving streams, ranging
from parts-per-trillion to parts-per-billion concentrations (1-6). Although the knowledge
of adverse effects of chronic low-dose exposure to pharmaceutical mixtures on human
health is limited, many toxicological studies have shown that exposures of fish and other
aquatic organisms to these anthropogenic compounds causes reproductive and behavioral
disorders (2, 7-9).
To minimize risk, efforts are underway to reduce human and aquatic organism
exposure to pharmaceuticals in waters, especially from municipal wastewater effluent,
the main route for pharmaceuticals to enter natural waters (10-13). Activated sludge
treatment in wastewater treatment plants has been shown to improve the removal of some
pharmaceuticals by using long sludge retention times (14-15). However, the majority of
pharmaceuticals are not completely degraded, and most existing municipal wastewater
treatment plants are not designed for operation using prolonged sludge retention times
(16). Finding alternative technologies that effectively remove pharmaceuticals from
wastewater effluents has, therefore, been of research interest (17-19).
Ultrasound, an advanced oxidation process (AOP), shows great potential to degrade
pharmaceuticals in wastewater as a tertiary treatment technology (20-23). Ultrasound has
156
unique advantages as compared to other technologies, such as no addition of chemicals,
ease of use, and short contact times (24-25). Ultrasonic irradiation induces chemical
reactions in water from the collapse of cavitation bubbles (26). The collapse of cavitation
bubbles generates localized hot spots (27-28), and these localized hot spots initiate
thermolytic and oxidation reactions with pharmaceuticals.
Properties of compounds, including surface excess (Γ) (29-30), octanol-water
partition coefficient (Kow) (31-32), vapor pressure (p) (31, 33), Henry’s law constant (KH)
(34-35), second-order rate constant with •OH (k•OH) (36-37), and diffusivity (Diw) (38-39),
have been shown to affect their degradation rates by ultrasound. These properties are
either thermodynamic (e.g., Γ, Kow, p, and KH) or kinetic (e.g., k•OH and Diw) parameters.
The thermodynamic parameters describe the tendency of pharmaceuticals to reach
equilibrium with specific regions of the cavitation bubbles. The kinetic parameters
determine the rates at which these compounds move to cavitation bubbles and react with
ultrasound-induced reactivity (i.e., heat and •OH). Although the dependence of both
thermodynamic and kinetic parameters on overall compound degradation has been
reported (26-36), the question regarding which aspect has a more profound influence on
the cavitational system remains unclear.
Pulsed wave (PW) ultrasound, whose ultrasonic signal train is separated by gaps of
no signal (40), features the discontinuation of bubble growth during the silent cycle.
Previous studies (41-44) found that PW ultrasound, under certain optimal conditions,
enhances the degradation of a compound, because it allows time (i.e., silent cycle) for the
surface active or small sized compound to diffuse to bubble-water interfaces, the sources
of reactivity. For instance, Xiao et al., (44) observed that degradation rates of seven
157
different pharmaceuticals and personal care products (PPCPs) by PW ultrasound are
compound dependent with degradation faster for smaller compounds or slower for larger
compounds than that under CW ultrasound. They linked the pulsed enhancement (PE) of
PPCPs to their Diw, specifically, a smaller sized PPCP is more able to diffuse to bubble
surfaces as compared to a larger sized PPCP. Yang et al. (45) investigated the sonolysis
of the surfactant 4-octylbenzene sulfonate, and the non-surfactant 4-ethylbenzene
sulfonate (EBS) in CW and PW mode. PW ultrasound was more effective than CW
ultrasound for OBS with a PE of 94 %; however, the increase in degradation rate of PW
over CW was not statistically significant for EBS. Enhancement was attributed to the
accumulation of the surfactant on cavitation bubble surfaces, resulting in reduced surface
tension and lowering the cavitation threshold (45). Our previous work evaluated PE for a
group of PPCPs. We observed a link between Diw and PE over a limited range of Diw and
Kow in deionized (DI) water.
In this study, pulsed wave (PW) ultrasound was used to systematically study the
roles that kinetic (i.e., Diw) and thermodynamic (i.e., Kow) parameters play in determining
pharmaceutical degradation. We selected six pharmaceuticals, namely fluorouracil (5-
FU), ibuprofen (IBU), clonidine (CLND), estriol (ESTO), nifedipine (NIFE), and
lovastatin (LOVS), on the basis of Kow and Diw. The Kow of these six compounds is in the
increasing order and Diw is in decreasing order (Figure 5.6 in the Supporting Information),
allowing us to observe trends in PE with Diw, Kow and PE. We investigated the
degradation of these six pharmaceuticals in DI and in a wastewater effluent to evaluate
the effect of environmentally relevant matrices on PE.
158
5.3 Experimental Methods
5.3.1 Materials
5-FU (99%), IBU (99%), CLND (99%), ESTO (97%), NIFE (98%), sodium
bicarbonate (99%), and sodium sulfate bicarbonate (99%) from Sigma-Aldrich, LOVS
(98%) from TCI, methanol (HPLC Grade), acetonitrile (HPLC Grade), HCl (Trace metal
Grade) and NaOH (97%) from Fisher Scientific, were used as received. Table 5.1 in the
Supporting Information lists physicochemical properties of the pharmaceuticals.
SupeliteTM
XAD8 and Amberlite® XAD4 resins were purchased from Sigma Aldrich. A
Chromaflex Chromatography column (1085 mL, 4.8 cm × 60 cm) with 0.20 mm bed
supports was purchased from Kontes (Vineland, NJ). Water used was from a Millipore
System (Millipore, MA) with a resistivity R = 18.2 MΩ cm.
5.3.2 Wastewater effluent collection
The wastewater effluent was collected in December 2011 from a wastewater
treatment plant, located in the vicinity of Columbus, Ohio, U.S.A. The facility was
established in 2001 and has a treatment capacity of 10 million gallons per day. The plant
uses traditional activated sludge treatment and denitrification technologies, but no
primary clarification. The wastewater is mainly composed of municipal wastewater. The
wastewater sample was collected from a tertiary filtered effluent stream and poured into a
50 L polyethylene carboy with minimal headspace. Using a Masterflex pump (Cole
Parmer), the wastewater was consecutively prefiltered through 5 and 0.45 μm
groundwater filtration capsules (Pall Gelman). Capsules were pre-rinsed with Milli-Q
water before use. Filtered wastewater effluent was acidified to pH 2 with HCl, air
159
stripped for 2 hours to remove H2S and NH3, and stored in the dark at 4 °C until use.
Before use the pH of a wastewater effluent aliquot was readjusted the field pH using
NaOH. The presence of Na+ and Cl
- ions due to pH adjustment (from pH 2 to 7.7) is
believed not to interfere with the cavitational system; the lowest reported concentration
affecting cavitation is 0.1 M (46-49).
5.3.3 Wastewater effluent organic matter isolation
The XAD8 and XAD4 resins were cleaned and packed according to Standley and
Kaplan (50). Organic matter from wastewater effluent was isolated using XAD 8 and
XAD 4 resin columns following Quaranta et al. (51). The XAD8 and XAD4 resin
columns were used to retain operationally defined hydrophobic (HPO) and transphilic
(TPI) fractions of organic matter, respectively. The organic matter that passed through
both columns consisted of the hydrophilic (HPI) fraction of organic matter and inorganic
salts. The two columns were connected consecutively by Teflon tubing. Approximately
40 empty bed volumes of filtered wastewater effluent passed through the columns with a
flow rate of 15 empty bed volumes per hour to ensure that a significant amount of organic
matter was sufficiently retained. The organic matter was back-eluted from each column
individually using 0.1 M NaOH and the wastewater effluent organic matter fractions
were stored in the dark at 4 °C prior to use.
5.3.4 Wastewater effluent organic matter characterization
Characterization of the original wastewater effluent was conducted before pH
adjustment and is summarized in Table 5.2 in the Supporting Information. The
160
concentration of dissolved organic carbon in the solution was quantified by a Shimadzu
TOC-5000A analyzer. Specific ultraviolet absorbance was measured at 280 nm
(SUVA280) with a photospectrometer (model UV-2401, Shimadzu), since SUVA280 is
strongly correlated to the aromaticity of organic matter (52). Concentrations of common
elements were measured by inductively coupled plasma Atomic Emission spectroscopy,
ICP-AES (Vista AX, Varian). The dominant cations included calcium and sodium. Anion
concentrations were determined by ion chromatography (DX-120, Dionex). The
dominant anions included chloride and sulfate. The dominant ions are believed not to
interfere with cavitational systems at the measured concentration. Molecular weight (MW)
distributions of the original and fractionated wastewater effluent organic matter were
determined with size exclusion chromatography (SEC) (Hewlett Packard 1050). A 1
mL/min mobile phase of pH 7.0 buffer solution of 0.1 M
tris(hydroxymethyl)methylamine and HCl flowed through a Protein-Pak 125 size
exclusion column (Waters Associates) and into a UV detector analyzing at 235 nm.
Calibration was performed using sodium polystyrene sulfonates (Polysciences) with
molecular weights of 18K, 8K, 4.6K, and 1.8K, and acetone, following the method of
Chin et al. (53). A linear calibration curve (R2 = 0.99) was obtained between log MW and
elution volume.
5.3.5 Sonochemical experiments
Ultrasound at 205 kHz was emitted from a USW 51-52 ultrasonic flat plate
transducer (A = 23.4 cm2) (ELAC Nautik, Inc., Kiel, Germany) into a water jacketed
glass vessel with a volume of 300 mL. The temperature of the reactor was maintained at
161
20 °C. A SM-1020 Function/Pulse generator (Signametrics Corp.) delivered sound waves
with an operation mode of 100 ms on and 100 ms off. The acoustic energy density to the
reactor, determined by calorimetry, was 45 W/L.
During experiments, 0.5 mL samples for chemical analysis were taken from the
reactor at designated times using a 1 mL glass syringe (Gastight 1001, Hamilton Corp.).
The total sample volume taken during the course of sonication did not exceed 1 % of the
initial volume.
5.3.6 Chemical analysis
A Hewlett-Packard 1100 HPLC equipped with a diode array detector (DAD) and a 5
µm, 150 × 2.1 mm SB-C18 column (Agilent Technologies) was used to quantify the
concentration of the pharmaceuticals. Eluents consisted of 20 mM pH 3 phosphoric
buffer and acetonitrile with a flowrate of 0.5 mL/min. The eluent ratios were different for
different compounds. The UV wavelengths were set at 265, 220, 220, 225, 238, and 238
nm for 5-FU, IBU, CLND, ESTO, NIFE, and LOVS, respectively.
5.3.7 Computational modeling
A non empirical calculation at the Hartree Fock (HF) level of theory with 6-31+G*
basis set was applied to determine the molar volume of the molecules (mL/mol) in water,
using the polarizable continuum model solvation method. HF/6-31+G* is considered to
be a reliable level of theory to calculate the geometry of large organic molecules (54-55).
All calculations were performed using Gaussian 09 (56) at the Ohio Supercomputer
Center.
162
5.4 Results and discussion
5.4.1 Sonochemical degradation in DI water
Sonochemical degradation rates of 5-FU, IBU, CLND, ESTO, NIFE, and LOVS in
DI water are shown in Figure 5.1a. The initial degradation rates of LOVS under CW and
PW ultrasound are 0.55 and 0.67 µM/min, respectively, which is faster than the other
compounds. 5-FU degraded slowest, with the degradation rates 0.11 and 0.13 µM/min
under CW and PW ultrasound, respectively.
Linear correlations have been reported between sonochemical degradation rates of
contaminants and their log Henry’s law constants (35), log Kow values (31), and
molecular weight (57) in aqueous solutions. For instance, Colussi et al., (35) found that at
various frequencies the sonolytic degradation rate of chlorinated hydrocarbons, including
CCl4, CHCl3, C2Cl6, and CH2Cl2, was faster with larger Henry’s law constants. Nanzai et
al., (31) reported that the initial degradation rate of twelve monocyclic aromatic
compounds linearly correlated to their log Kow values. Fu et al., (2007) observed a
negative correlation between the rate of sonolysis of eight different estrogen compounds
and their molecular weight. The dependence of degradation rates of the pharmaceuticals
under both CW and PW ultrasound on physicochemical properties, including log Henry’s
law constants, log Kow, molecular weight, and molar volume was evaluated by non-
parametric correlations (Spearman’s rho) to analyze for any possible relationships.
Surprisingly, there is a positive correlation between degradation rate and molar volume
(Figure 5.7d in the Supporting Information), suggesting that a larger sized pharmaceutical
is destroyed faster than the small sized one. The suggestion is inconsistent with Fu et al.
163
(57) and may result from the set of PPCPs chosen in our study. Although the degradation
rate is not perfectly correlated (R2>0.8) with Kow, Figure 5.7b shows a hydrophobic
compound degrades faster than hydrophilic one. The observation is consistent with
Nanzai et al. (31).
In Figure 5.1a the initial degradation rates for 5-FU, NIFE, and LOVS are faster
under PW mode as compared to CW mode. For IBU, CLND, and ESTO, exhibit the
opposite trend. In order to evaluate the effect of pulsing ultrasound on degradation
kinetics, PE, a measure used to compare the difference in initial degradation rates
between CW and PW ultrasound, was determined by eqn. 5.1.
％ 100
dt
Cd
dt
Cd
dt
Cd
(%) PE
CW
CWPW (5.1)
where CW
dt
Cd and
PWdt
Cdare the initial degradation rates of a pharmaceutical under
CW and PW ultrasound, respectively.
The PE of the six pharmaceuticals in DI water is illustrated in Figure 5.2a. Pulsing
ultrasound is beneficial for 5-FU, NIFE, and LOVS, while IBU, CLND, and ESTO have
negative PE values. The U-shaped curve of PE vs. molecular weight in DI water shows
that, a pharmaceutical with either the smallest MW (i.e., 5-FU, MW =131 g/mol) or the
greatest Kow (i.e., LOVS, Kow = 4.26) was positively affected by pulsing. For the
compounds with medium diffusivity and hydrophobicity, PW ultrasound did not increase
the degradation kinetics.
164
In Chapter 4, we investigated the effect of pulsing on seven PPCPs which covered a
range of physicochemical properties and a diversity of structures. The PE of PPCPs was
linked to the diffusivity of the PPCP by observing an inverse relationship between the
molar volume of a PPCP and its PE. We did not explore the impact of hydrophobicity on
PE explicitly. Thus, this study verifies and builds on the previous work.
The benefit of PW ultrasound is the retardation of bubble growth during the silent
cycle, providing time for a pharmaceutical to diffuse to bubble-water interfaces and react
with the cavitation bubbles within the subsequent pulse. Thus, a strong hydrophobic or
highly diffusive compound has the preference or momentum to migrate to bubble-water
interfaces during the silent cycle, resulting in more molecules reacted by ultrasound-
induced reactivity. Our results suggest that a pharmaceutical with either high diffusivity
or hydrophobicity more readily populates the bubble-water interface during the silent
cycle of PW ultrasound. However, for the compounds with medium hydrophobicity and
diffusivity, their PE values are smaller suggesting that there is little to no benefit to using
PW ultrasound.
This can be explained by Fick’s law, defined in eqn. 5.2.
dz
dCDJ i
iwi (5.2)
where Ji is the mass flux of chemical i in direction of concentration gradient with; Ci is
the concentration of i; and z is distance in the direction of concentration gradient (i.e.,
thickness of bubble interface). As shown in eqn. 5.2, the flux towards to bubble-water
surface is a function of diffusivity and concentration gradient. The concentration gradient
depends on the enrichment of a pharmaceutical on bubble surfaces, a hydrophobic
165
compound accumulates more on bubble surfaces than a hydrophilic one; thus, for the
same initial concentration and diffusivity, a hydrophobic compound results in a greater
concentration gradient, ultimately a greater flux into bubbles. Similarly, two compounds
at the same concentration and hydrophobicity will accumulate on a bubble surface to the
same degree at equilibrium but the compound with a high Diw will result in a higher flux
and consequently, higher PE value. Therefore, pharmaceuticals with a combination of
diffusivity and concentration gradient result in the largest flux exhibit greater PE values.
Diffusivity is an inverse function of the molar volume of the diffusing compound
(58). However, molar volume information is not readily available. Easily obtainable
molecular weight generally correlates well with molar volume, resulting in its use (Figure
5.8 in the Supporting Information). Although molecular weight correlates well with
molar volume, the smoother curve of Figure 5.2b demonstrates that molar volume is a
better parameter than molecular weight for investigating relative effects of diffusivity.
5.4.2 Sonochemical degradation in wastewater effluent
The sonochemical degradation rates of 5-FU, IBU, CLND, ESTO, NIFE, and LOVS
in wastewater effluent are shown in Figure 5.1b. Unlike the trend observed in DI water,
NIFE degraded faster than the other compounds with the initial degradation rates 0.35
and 0.34 µM/min under CW and PW, respectively. 5-FU decomposed slowest,
approximately one order of magnitude slower than that of NIFE. The slowest degradation
kinetics of 5-FU in wastewater effluent is similar to that in DI water, suggesting that 5-
FU is recalcitrant to ultrasound-induced reactivity.
166
As expected, the overall slower degradation of the pharmaceuticals in the
wastewater effluent indicated that environmentally relevant matrices inhibit degradation.
For example, the initial degradation rates of LOVS and CLND in wastewater were
reduced by approximately 90% and 60%, respectively, as compared to in DI water.
Previous work has observed a similar effect (59-62). Sanchez-Prado investigated the
sonochemical degradation of triclosan in DI water and domestic wastewater (62). The
degradation rate constant was reduced by 76.8% in the matrix as compared to in DI water.
Cheng et al., investigated the effects of organic and inorganic matrices in groundwater on
sonochemical degradation of perfluorooctane sulfonate (PFOS) and perfluorooctanoate
(PFOA) (59-60). They concluded that both matrices negatively affect the sonochemical
kinetics due to organic matter adsorption to the bubble-water interface and inorganic ions
partitioning to and interaction with cavitation bubbles, hence decreasing the degradation
rates.
In wastewater effluent, the enhancement of degradation rates of the pharmaceuticals
in PW ultrasound disappeared. This effect was most prominent for 5-FU and LOVS
(Figure 5.2a). Therefore, the benefit of pulsing ultrasound in DI water may not be
applicable to municipal wastewater effluent. This lack of benefit of pulsing ultrasound in
wastewater effluent may be due to salts or effluent organic matter interfering with
degradation.
5.4.3 Effect of inorganic and organic matrix on PE of pharmaceuticals
When AOPs are investigated in water treatment, the water matrix containing
inorganic and organic components significantly slows the degradation of the target
167
contaminants (59-60, 63) due to matrix competition for •OH and/or sequestration of
contaminants from AOP reactivity. In ultrasound, the matrix may react with •OH, diffuse
to cavitation bubbles, quenching bubble dynamics, and alter the availability of
compounds to ultrasound-induced reactivity (64-65). However, there is little known about
how environmentally relevant matrices influence the enhanced degradation kinetics under
PW ultrasound. To this end, we evaluated the effect of inorganic and organic matrices,
independently.
5.4.3.1 Effect of inorganics on PE of pharmaceuticals
Previous studies (49, 66) indicated that sulfate and bicarbonate significantly affect
bubble dynamics as compared to other anions, such as chloride and nitrate (49).
Therefore, sulfate and bicarbonate anions were selected as matrix inorganics to elucidate
their effect on the PE of pharmaceuticals. Decreased degradation of PFOA and PFOS was
attributed to sulfate and bicarbonate decreasing the negative charge at the bubble-water
interface and altering the interfacial water structure and adsorption kinetics of target
compounds to bubble surfaces. Sun et al., (66) investigated the degradation kinetics of
acid black 1 (AB1) in the presence of 0.5 g/L sulfate and bicarbonate in the
ultrasound/Fenton system and they attributed the inhibited degradation to anions reacting
with •OH (eqn. 5.3 and 5.4).
3HCO + •OH •
3CO + H2O (5.3)
2
4SO + •OH •
4SO + OH
- (5.4)
168
We expected that the presence of 2
4SO and
3HCO would either react with •OH or
partition to the bubble-water interface, reducing the temperature of collapsing cavitation
bubbles, ultimately resulting in reduction of PE.
The PE of the six pharmaceuticals in the presence of 2 mM 2
4SO and
3HCO is
shown in Figure 5.3. Similar to the results of PE of the pharmaceuticals in DI water, the
PE values for 5-FU and LOVS are highest among the six compounds. In the presence of
sulfate anion, the PE for 5-FU and LOVS were 28.4 % and 20.4 %, respectively. With the
presence of bicarbonate anion, the PE for 5-FU and LOVS were 39.6 % and 12.3 %,
respectively. The presence of anions reduced the PE of NIFE, from 9.3% in the absence
to -35.3% and -44.7% for 2
4SO and
3HCO , respectively. Cheng et al., reported 2
4SO
and3
HCO could accumulate on bubble-water interface and change bubble dynamic (49).
The accumulation causes localized concentration of the anion within a certain distance
from bubble interfaces higher than in bulk solution. The hydrophobic NIFE accumulates
on the bubble surfaces to the same extent as 2
4SO and
3HCO , thus resulting in the most
pronounced interference by the anions, ultimately reducing PE of NIFE. For the other
compounds, the presence of anions seems to not decrease their PE values. The little
influence of the anions may be due to more accumulation of target compounds (e.g. 5-FU
and LOVS) or less accumulation of target compounds (e.g. IBU, CLND, and ESTO) on
bubble surfaces as compared to anions. The U-shaped curve (Figure 5.3) in the presence
of the anions resembles that of the DI water curve, indicating that inorganics do not
significantly affect the accumulation behavior of the pharmaceuticals on bubble surfaces
during the silent cycle, thus not altering the effect of pulsing on degradation kinetics.
169
5.4.3.2 Effect of organics on PE of pharmaceuticals
Matrix organics have been shown to have different effects on sonochemical
degradation of the target compounds (61, 67-69). Matrix organics are substances with
diverse size and polarity; the influence of various components on sonochemical
degradation rates in CW or PW ultrasound is unknown. In order to gain a mechanistic
understanding of the role of matrix organics in wastewater effluent on sonolysis of
pharmaceuticals, we fractionated the effluent organic matter (EfOM) into operationally
defined HPO, TPI, and HPI fractions to determine how size and hydrophobicity of EfOM
affects sonolysis of target compounds. As illustrated in Table 5.3 in the Supporting
Information, the weight-averaged molecular weights of EfOM were smaller than
observed for natural DOM (53). In addition, the SUVA280 value of the TPI fraction was
larger than the corresponding HPO and HPI fractions, suggesting that the TPI fraction
may have more aromatic character than the HPO and HPI fractions (51). Sonolysis of 5-
FU and LOVS was conducted in these three fractions with a DOC of 3 mgC/L in each
fraction.
Figure 5.4 shows the initial sonochemical degradation rates of 5-FU and LOVS in
different fractions of EfOM. It is clear that the TPI fraction negatively affects the initial
degradation rates of both pharmaceuticals more than the HPO and HPI fractions of EfOM.
For 5-FU, compared to DI water, the presence of the TPI fraction reduced the rates by
53.1% and 66.9% in CW and PW ultrasound, respectively. HPO and HPI fractions of
EfOM both had a similar effect on the degradation kinetics of 5-FU. For LOVS, the HPI
fraction inhibited the degradation to a greater extent than the HPO fraction.
170
The pulsing effect on the degradation of 5-FU and LOVS in different fractions is
illustrated in Figure 5.5. Both compounds exhibited a negative PE in all fractions of
EfOM, while the TPI fraction reduced the PE the most.
The reduction in PE can be explained by both molecular size and the hydrophobicity
of the TPI portion of EfOM. The molecular weight of TPI (146 g/mol) is considerably
less than HPO (474 g/mol), suggesting that the size of transphilic fraction of EfOM is
significantly smaller that the hydrophobic fraction. Consistent with our previous
observations (70), small sized matrix organics perturb the accumulation of
pharmaceuticals during the silent cycle, quenching the ultrasound induced reactivity
under PW ultrasound. Xiao et al., (70) sonolyzed ciprofloxacin and ibuprofen in the
matrix organics of terephthalic acid (TA) and Suwannee River Fulvic Acid (SRFA).
Smaller-sized matrix organics (i.e., TA) had a greater impact on the degradation rates of
target contaminants by ultrasound compared to larger-sized matrix organics (i.e., SRFA).
Although with molecular weight of TPI (146 g/mol) slightly greater than HPI (128 g/mol),
the PE of LOVS in presence of TPI is less than that in the presence of HPI, suggesting
that the size of pharmaceutical may not be the exclusive reason for PE reduction.
As illustrated in eqn. 5.2, the diffusivity and concentration gradient co-determine the
flux of organic matrix towards to bubble surfaces. The molecular weight of TPI is similar
to HPI, indicating that the diffusivity of these two fractions is similar. In addition, the TPI
fraction is operationally more hydrophobic than HPI based on our isolation method.
Therefore, with the same DOC and similar diffusivity of TPI and HPI, the concentration
gradient for TPI fraction is greater than HPI, resulting in a greater flux of TPI into
171
bubbles than that of HPI during the silent cycle. Thus, the TPI fraction reduces the PE to
a greater extent than HPI.
Figure 5.5 shows that, in the presence of the HPI fraction, the PE of LOVS (-9.44%)
is significantly less than that of the HPO fraction (-40.47%). Considering the molecular
weight for HPI is less than HPO (MWHPI=128 g/mol and MWHPO=474 g/mol) but the
HPO fraction is operationally more hydrophobic than HPI, the result of less PE in the
presence of HPI than HPO suggested that the effect of molecular weight of EfOM is
more influential than that of hydrophobicity in determining PE of target compounds.
Figure 5.5 also shows the PE for 5-FU is greater than that of LOVS in each fraction
of EfOM. For instance, in the HPO fraction of EfOM, the PE values for 5-FU and LOVS
are similar, but in TPI fraction, PE values for 5-FU and LOVS were -11.5 and -69.3%,
respectively. The discrepancy indicates that the size and hydrophobicity of target
pharmaceuticals influence PE in the presence of EfOM. We suspected the size may be
more influential than hydrophobicity in determining the PE in different fractions of
EfOM. For the HPO fraction, the PE of two compounds was similar. This is due to the
larger molecular weight of HPO (MWHPI=474 g/mol) than LOVS (MWLOVS = 404 g/mol)
and 5-FU (MW5-FU= 103 g/mol), resulting in the presence of HPO fraction does not
change their PE values. For the TPI and HPI fractions, the molecular weight of TPI
(MWTPI=146 g/mol) and HPI (MWHPI=128 g/mol) are significantly less than LOVS but
greater than 5-FU, suggesting that a large sized pharmaceutical (i.e., LOVS) is less
steadily to diffuse to bubble surfaces, consequently resulting in a less PE value than small
size one (i.e., 5-FU). It is noted that, although 5-FU (log Kow= -1) is less hydrophobic
than LOVS (log Kow= 4.26), the Kow value may not correctly reflect the partitioning
172
behavior of 5-FU and LOVS to different fractions of EfOM in water, especially for the
compound with log Kow greater than 4 (71).
In summary, the presence of EfOM, especially small sized EfOM, negatively affects
the sonochemical degradation of pharmaceuticals under PW ultrasound. EfOM diffuses
or partition to bubble surfaces during the silent cycle, thus either competing the
adsorption sites for target compounds or interacting at the bubble-water interface,
lowering the bubble dynamics. These interferences will unavoidably reduce the pulsing
effect on pharmaceutical degradation, hence resulting in a negative PE value.
5.5 Environmental application
We evaluated the effect of inorganic and organic matrices to determine the
components in wastewater effluent reduced the pulsing effect on the sonolysis of the
pharmaceuticals. Our results showed that the organic matrix was the cause of the
reduction in PE. By isolating the EfOM, we further concluded that the TPI fraction
caused the most significant drop in PE during the sonolysis of pharmaceuticals. Small
sized EfOM diffuses to and accumulates at the bubble-water interface during the silent
cycle, thus impede the sonochemical degradation of the target pharmaceuticals under the
PW ultrasound the most. The inhibition effect is more prominent for hydrophobic target
pharmaceuticals. Overall, CW ultrasound is preferred over PW to enhance removal
efficiency of target pharmaceuticals by ultrasound in the presence of EfOM. It is optimal
to remove of matrix organics (especially small sized organics) before ultrasound
treatment to maximize its efficiency.
173
Acknowledgements
Funding from NOAA/Ohio Sea Grant College Program is gratefully acknowledged.
R. X. gratefully acknowledges Matthew Noerpel for the valuable comments on this
manuscript.
Supporting Information
Supporting information contains three tables and three figures. This material is
available free of charge via the Internet at http://pubs.acs.org.
174
5-FU IBU CLND ESTO NIFE LOVS
0.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
d[C
]/d
t|0 (
M/m
in)
CW
PW
a
5-FU IBU CLND ESTO NIFE LOVS
0.0
0.1
0.2
0.3
0.4
0.5
d[C
]/d
t|0 (
M/m
in)
CW
PWb
Figure 5.1: The initial sonochemical degradation rates of 5-FU, IBU, CLND, ESTO,
NIFE, and LOVS in DI water (a) and wastewater effluent (b) (pH 7.7, 20 °C,
[pharmaceutical]0 = 10 μM, and sonication power density at 45 W/L).
175
100 150 200 250 300 350 400 450-15
-10
-5
0
5
10
15
20
25
30
Molecular weight (g/mol)
Apparent log Kow
PE
(%
)
PE in WW
PE in DI
CLND
5-FU
IBU
ESTO
NIFE
LOVS
a
-1 0 1 2 3 4
5-FU
IBU
CLNDESTO
NIFE LOVS
50 100 150 200 250 300 350 400-15
-10
-5
0
5
10
15
20
25
30
b
Molar volume (mL/mol)
PE in DI
PE
(%
)
5-FU
IBUCLND
ESTO
NIFE
LOVS
Figure 5.2: The PE of the six pharmaceuticals in DI water and wastewater effluent (a). PE
in DI water was plotted as a function of molar volume (b) (pH 7.7, 20 °C,
[pharmaceutical]0 = 10 μM, and sonication power density at 45 W/L).
176
100 150 200 250 300 350 400 450
-60
-40
-20
0
20
40
60
2 mM SO2-
4
2 mM HCO-
3
in DI
Molecular weight (g/mol)
PE
(%
)
-1 0 1 2 3 4
Apparent log Kow
5-FU
IBU
CLND
ESTONIFE
LOVS
Figure 5.3: The PE of the six pharmaceuticals in DI water (pH 7.7, 20 °C,
[pharmaceutical]0 = 10 μM, and sonication power density at 45 W/L). In DI water, the
concentration of bicarbonate is 2.5 ×10-4
M, assuming water is equilibrated with the
atmosphere.
177
DI HPO TPI HPI0.01
0.1
1
5-FU CW
5-FU PW
LOVS CW
LOVS PW
d[C
]/d
t|0 (
M/m
in)
Figure 5.4: The initial sonochemical degradation rates of 5-FU and LOVS in DI water
and different fractions of effluent organic matter (pH 7.7, 20 °C, [pharmaceutical]0 = 10
μM, [DOC]= 3 mgC/L, and sonication power density at 45 W/L).
178
HPO TPI HPI-80
-70
-60
-50
-40
-30
-20
-10
0
10
PE
(%
)
5-FU
LOVS
Figure 5.5: The PE of 5-FU and LOVS in different fractions of effluent organic matter
(pH 7.7, 20 °C, [pharmaceutical]0 = 10 μM, [DOC]= 3 mgC/L, and sonication power
density at 45 W/L).
179
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189
Supporting Information
190
Table 5.1: Selected physicochemical properties of 5-FU, IBU, CLND, ESTO, NIFE, and LOVS at 25°C.
1: The heat stability of the pharmaceutical was approximated by the lowest bond energy in that compound.
2: The apparent log Kow values for the pharmaceutical at pH 7.7 were determined by ACD/LogD version 6.00.
3: Molar volume was calculated at Hartree Fock (HF) level of theory with 6-31+G* basis set and PCM solvation method.
190
191
Table 5.2: Summary of main wastewater effluent characteristics
pH 7.7 3
4PO (mg/L) 5.8
TOC (mgC/L ) 8.88 Ca (mg/L) 75.85
alkalinity (mg/L as CaCO3) 39.4 K (mg/L) 12.65
conductivity (μS/cm) 845 Mg (mg/L) 19.3
Cl- (mg/L) 70.4 Na (mg/L) 57.8
2
4SO (mg/L) 94.5 S (mg/L) 35.4
192
Table 5.3: Summary of main characteristics for different fractions of EfOM
HPO TPI HPI
weight-averaged MW (g/mol) 474 146 128
SUVA280 (L mg-1
m-1
) 1.09 2.17 0.74
% total DOC 47.9 5.4 46.6
193
-2 -1 0 1 2 3 4 550
100
150
200
250
300
350
400
450
ESTO
NIFE
LOVS
CLND
Mo
lecu
lar
weig
ht
(g/m
ol)
Apparent log Kow
5-FU
IBU
Figure 5.6: The molar weight of 5-FU, IBU, CLND, ESTO, NIFE, and LOVS was plotted
against their apparent log Kow at pH 7.7.
194
-12 -11 -10 -9 -8 -7 -60.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
log Henry's law constant
CW
PWd
[C]/
dt|
0 (
M/m
in)
5-FU
IBU
CLND
ESTO
NIFE
LOVS(a)
-2 -1 0 1 2 3 4 50.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
(b) LOVS
NIFE
ESTO
CLND
IBU
5-FU
CW
PW
d[C
]/d
t|0 (
M/m
in)
Apparent log Kow
100 150 200 250 300 350 400 4500.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
(c) CW
PW
d[C
]/d
t|0 (
M/m
in)
Molecular weight (g/mol)
5-FU
IBU
CLND
ESTO
NIFE
LOVS
50 100 150 200 250 300 350 400 4500.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
(d)
5-FU
CW
PW
d[C
]/d
t|0 (
M/m
in)
Molar volume (mL/mol)
CLND
IBU
ESTO
NIFE
LOVS
Figure 5.7: Degradation rates of the pharmaceuticals under CW and PW ultrasound in DI
water were plotted against log Henry’s law constant (a), apparent log Kow (b), molecular
weight (c), and molar volume (d).
195
50 100 150 200 250 300 350 400100
150
200
250
300
350
400
450
Mo
lec
ula
r w
eig
ht
(g/m
ol)
Molar volume (mL/mol)
5-FU
CLND
IBU
ESTO
NIFE
LOVS
Figure 5.8: The molar volume of 5-FU, IBU, CLND, ESTO, NIFE, and LOVS was
plotted against their molecular weight. (Molar volume was calculated at Hartree Fock
(HF) level of theory with 6-31+G* basis set and PCM solvation method)
196
Chapter 6: Conclusions and Future Work
6.1 Conclusions
Ultrasound, as a water treatment technology, has been studied in this dissertation to
remove PPCPs under different conditions. The major conclusions we drew are the
following.
First, as discussed in Chapter 2, matrix organics play roles in sonochemical
degradation of CIPRO and IBU, an example for hydrophilic and hydrophobic
pharmaceuticals, respectively. In the absence of matrix organics, the degradation rates of
CIPRO and IBU depend on their initial concentrations and ultrasonic frequencies: a short
half-life is observed as the initial concentration decreases, and 620 kHz is more effective
as compared to 20 kHz. In the presence of matrix organics, the degradation of CIPRO
and IBU was affected by matrix organics to a different extent. A more hydrophobic target
compound with higher diffusivity is impacted little by the presence of matrix organics.
The diffusivity and concentration of matrix organics determine treatability by ultrasound.
A smaller size and higher concentration of matrix organics in water have a larger impact
on the treatability of target contaminants by ultrasound as compared to a larger size and
lower concentration of matrix organics.
Second, PW ultrasound was used to systematically study whether and how •OH
scavengers affect cavitation bubbles in Chapter 3. Based on the reduction in PE of the
197
probe compound, CBZ, all the tested scavengers (e.g. FA, CA, TA/TPA, KI, MS, BS, and
AA/acetate) except AA/acetate affect cavitation bubbles and are not recommended to use
as bulk •OH scavengers in cavitational systems. The range of utility of AA/acetate as a
bulk •OH scavenger was tested under different concentrations (0.5 mM to 0.1 M) and pH
values (3.5 to 8.9).
Third, in Chapter 4, five pharmaceuticals (CBZ, IBU, ATP, SFT, and CIPRO), and
two personal care products (PG and DP) have been degraded under CW and PW
ultrasound, respectively in aqueous solutions. The independence of degradation rates of
the PPCPs under both CW and PW ultrasound on their physicochemical properties
indicates that there is no single property that controls the degradation kinetics. For the
enhancement due to pulsing, there exists a relationship between PE and the molar volume,
indicating that a small size PPCP exhibits a positive PE, while a large size PPCP yields a
negative PE. This correlation implies that diffusion governs transport of the PPCPs to the
bubble surface during the silent cycle. To explore this phenomenon, we applied the
knowledge gained in Chapter 3. A bulk solution •OH scavenger, acetic acid, was added in
the sonicated system to differentiate the contribution of bulk •OH oxidation to the overall
degradation. The fraction of degradation occurring in bulk solution is positively
correlated with the molar volume of the compound. Small size molecules may quickly
diffuse to the cavitation bubbles, resulting in a higher portion of the PPCP in and around
cavitation bubbles. Our results suggest that PW ultrasound is an efficient method to
enhance degradation for the small size PPCPs, while CW ultrasound is more effective to
degrade large size PPCPs in aqueous solution.
198
In Chapter 5, we investigated the sonolysis of six pharmaceuticals, namely 5-FU,
IBU, CLND, ESTO, NIFE, and LOVS, in DI water and wastewater effluent. Unlike in
Chapter 4, our results showed that in DI water pharmaceuticals with the highest Diw (i.e.,
5-FU) and Kow (i.e., LOVS) exhibited the greatest enhancement in degradation rates in
PW mode as compared to CW mode. However, in municipal wastewater effluent, the PE
for 5-FU and LOVS disappeared. By evaluating the effect of inorganic and organic
matrices, we concluded that the organic matrix was the cause of the reduction in PE in
the wastewater effluent matrix. By isolating the EfOM, we confirmed that the TPI
fraction caused the most significant drop in PE during the sonolysis of pharmaceuticals,
because it had combined characteristics of both high diffusivity and aromaticity
compared to HPO and HPI fractions of EfOM, further corroborating that small sized
matrix organics inhibit the pulsing effect on degradation kinetics. This conclusion was
consistent with the conclusion in Chapter 2. Our results provide insight into the
application of ultrasound as a water treatment technology. Overall, CW ultrasound is
preferred over PW to enhance removal efficiency of target pharmaceuticals by ultrasound
in the presence of EfOM. If EfOM is removed before sonication of the pharmaceuticals,
PW ultrasound exhibits a faster degradation of pharmaceuticals with either high
diffusivity or great hydrophobicity.
6.2 Future work
Our study showed the ultrasound is able to degrade PPCPs under different
conditions and the degradation kinetics depend on the ultrasonic mode and
199
physicochemical properties of PPCPs. However, there are several questions need to
address to better understand the system.
(1) Extend the study of EfOM in determining PPCPs degradation kinetics under PW
ultrasound. We observed that the PE values for 5-FU and LOVS disappeared in
wastewater matrix as compared to that in DI water and attributed the disappearance of PE
to the presence of EfOM. The TPI fraction of EfOM diffuses to cavitation bubble
surfaces and interacts with pharmaceuticals during the silent cycle. Although the size
distribution and aromaticity for different fraction have been measured, the aromaticity of
EfOM is approximated to hydrophobicity but not necessarily equal to hydrophobicity.
Measuring the hydrophobicity of different fractions of EfOM can help us to be more
accurate to understand the intermolecular interaction between pharmaceuticals and each
fraction of EfOM. To confirm the role of each fraction of EfOM played, it is beneficial to
sonicate a pharmaceutical that does not affected by pulsing in DI water (e.g. IBU) in
different fractions of EfOM. It is expected that such a compound should have little
alteration of PE in the presence of EfOM. Based on our theory, a compound without a
positive PE in DI water should not have preference or momentum to diffuse to the
bubble-water interface during the silent cycle. Therefore, the presence of different
fractions of EfOM should not change its PE value.
(2) Extend the study of the role of PPCPs diffusivity in determining PE. In Chapter 5,
seven PPCPs were degraded under CW and PW ultrasound, respectively. We observed
that the degradation rates by PW ultrasound are compound dependent with degradation
either faster for smaller compounds or slower for larger compounds than that under CW
ultrasound. The diffusivity was linked to the PE: small sized PPCPs have higher PE
200
values than the large sized PPCPs. Although we have verified our conclusion by using a
bulk •OH scavenger, acetic acid, there may be an alternative way to elucidate the role of
diffusivity of the PPCP in PE. We could investigate the sonolysis of mixtures of PPCPs.
This study is interesting from both a fundamental and a practical point of view. The
sonolysis of mixtures of PPCPs will help us to better understand the role of diffusivity.
We expect that a small sized PPCP will degrade faster than a large sized PPCP in binary
mixtures based on our conclusion in Chapter 4. Additionally, when it comes to
application of ultrasound in wastewater, mixtures of PPCPs are expected. Thus, a better
understand of how the presence of other PPCPs impacts degradation kinetics is
practically necessary.
(3) Investigate the frequency effect on PPCPs removal under PW ultrasound. Our
study have showed that both small size and hydrophobic PPCPs are able to more readily
diffuse to bubble interface and are impacted more by pulsing ultrasound in aqueous
solutions. The enhanced degradation rate is attributed to the accumulation of small size or
hydrophobic PPCPs during the silent cycle. Therefore, increasing the accumulation of
target compounds on bubble interface will result in a faster degradation rate under PW
mode. We could test our theory at higher frequencies (up to a megahertz). Since the
surface to volume ratio of bubble is larger and bubble population is denser at higher
frequency as compared to low frequency. The increased bubble surface could increase the
accumulation of PPCPs during the silent cycle, which ultimately resulting in fast
degradation kinetics under PW ultrasound.
(4) Investigate the effect of solid particles on sonolysis of PPCPs. We examined the
effects of inorganic and organic matrices on PPCPs degradation. However, it is
201
interesting to examine the degradation kinetics in presence of total suspended solids.
Particles may serve as additional nuclei, thus increasing the numbers of cavitation events
and accelerating the degradation of PPCPs. The presence of solid particles may exhibit
different influences on degradation kinetics. On the other hand, the particles scatter the
ultrasonic wave in water, weakening the energy dissipating into water, and perturbing the
bubble distribution, eventually decreasing the removal efficiency of PPCPs.
Understanding the effect of solid particles will help us better design the application of
ultrasound in PPCPs removal in water treatment plants.
202
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Appendix: Thermodynamic and kinetic study of ibuprofen with hydroxyl radical: A
density functional theory approach
Ibuprofen, a common non-prescription non-steroidal anti-inflammatory drug
(NSAID) and a frequently-detected pharmaceutical in natural waters, was studied using
density functional theory (DFT) at the B3LYP/6-31+G**//B3LYP/6-31G* level of theory
in its reaction with hydroxyl radicals. The reaction pathways that were studied included
•OH addition to the aromatic ring, abstraction of a hydrogen atom, and nucleophilic
attack on the carbonyl group. The energy barriers for most additions are moderate, while
abstraction and nucleophilic attack reaction barriers are low or non-existent. The
hydrogen-atom - abstraction and nucleophilic attack pathways are more favorable than
•OH addition, which is in agreement with previous experimental observations. The
enthalpy change for H-atom abstraction reactions range from -35.31 to -17.32 kcal/mol,
with a median of -32.04 kcal/mol; the enthalpy change for nucleophilic attack on the
carbonyl group is -35.41; and the enthalpy change for •OH addition ranges from -18.17 to
-11.71 kcal/mol, with a median of -13.07 kcal/mol. The calculated rate constant, 7.4 × 109
M-1
s-1
, is in agreement with experimentally determined values. Our results suggest that
advanced oxidation processes which generate reactive hydroxyl radicals, as tertiary
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treatment technology in water and wastewater treatment plants, has a great potential to
remove ibuprofen in water.
A.2 Introduction
Thousands of tons of pharmaceuticals are consumed or discarded by the public
every year [1-3]. These compounds find their way into the water system as pollutants
from the pharmaceutical industry, excretory products of medically treated humans, and
incomplete removal in wastewater treatment plants [4]. The U.S. Geological Survey
(USGS) provided the first overview of the occurrence of pharmaceuticals in water
resources across the United States during 1999-2000 [5]. They found these anthropogenic
compounds were prevalent at 139 sampling sites, 80% of the streams sampled. In their
study, the frequency of detection of ibuprofen, a non-prescription non-steroidal anti-
inflammatory (NSAID), was 9.5% among all of the 84 samples, and the median
detectable concentration was 0.2 µg/L. More recent studies reported environmental
concentrations of ibuprofen in the range of 10 ng/L to 169 µg/L in an analysis of effluent
from waste treatment plants in Spain [6].
Although the concentrations are very low, they may represent a potential hazard for
human health. In addition, the metabolites can be more harmful than the parent organic
compounds [7-9]. Toxicological studies have shown that exposure of fish and other
aquatic organisms to the pharmaceuticals can cause adverse reproductive effects such as
reduced viability of eggs, sexual disruption, and changes in sperm density [10-11].
However, a full understanding of the comprehensive hazards of these pharmaceuticals at
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environmental concentrations, which may cause modification of genetic expression and
protein function, is limited.
Therefore, based on precautionary principles, pharmaceuticals in water sources,
especially in drinking water resources, should be sufficiently treated to minimize their
potential risk to the ecosystem and human health.
Efforts have been made to improve the pharmaceutical removal efficiency in water
treatment plants. Studies have shown the conventional adsorptive treatment technologies,
such as solids removal, activated carbon, and conventional activated sludge, are not
always capable of eliminating pharmaceuticals in water with satisfactory efficiency [12-
14]. Oulton et al. [12], investigated the removal efficiency of ibuprofen in the influent
and effluent wastewater from 65 traditional treatment plants around the world and found
that the efficiency was plant dependent, ranging from ca. 0 to 100%. On the other hand,
advanced oxidation processes (AOPs), processes in which hydroxyl radicals (•OH) are
formed at ambient temperature and atmospheric pressure, have been found to be
promising for degrading organic pollutants in waters. AOPs generate highly reactive •OH
with oxidation potentials of 2.74 V and 1.77 V in acidic and neutral/basic solution,
respectively [15]. Consequently, •OH could non-selectively and rapidly oxidize nearly all
organic pollutants, thus exhibiting a high potential to remove pharmaceuticals.
There are considerable numbers of experimental AOPs studies which have
investigated the degradation kinetics for pharmaceuticals, including ibuprofen [16-18].
Packer et al. [18] determined the degradation rate of ibuprofen in both direct photolysis
and •OH mediated indirect photolysis. The second order rate constant of ibuprofen
reacting with •OH was measured to be 6.5 ± 0.2 ×109 M
-1 s
-1. Xiao et al. [19] irradiated
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ibuprofen in the presence of natural organic matter (NOM) using ultrasound and
concluded that ibuprofen competes with NOM in water for ultrasound-induced •OH.
Mendez-Arriaga et al. [20] identified several byproducts of ibuprofen in the photo-Fenton
system and found that two major byproducts in their system were decarboxylated and
hydroxylated metabolites of ibuprofen. Although the removal of ibuprofen in water by
AOPs has been achieved, it is still unknown how ibuprofen reacts with •OH the on
molecular level, what the possible degradation pathways are, and whether the detected
metabolites are directly derived from the oxidation of ibuprofen or if they are
secondary/tertiary byproducts.
Quantum mechanical calculations provide unique insights about chemical reactivity
on the molecular level, thus it becomes very attractive to investigate the degradation of
organic pollutants by AOPs technologies. Among all of the quantum mechanics methods,
density functional theory (DFT) methods [21] have been reported to be a reliable and
affordable approach in predicting reaction mechanisms, byproduct formation, and
reaction kinetics [22-24]. For example, Minakata and Crittenden [25] modeled rate
constants between •OH and eight simple pollutants (i.e., CH3OH, CH3CHO, CH3OCH3,
CH3COCH3, CH3COOH, CH3F, CH3Cl and CH3Br). The estimated rate constants
calculated by the DFT approach were within a factor of 5 of the experimental values.
Morales-Roque et al. [26] studied •OH oxidation process of phenol in water with DFT
method and demonstrated that the ortho position is more readily attacked by •OH
compared to para and meta positions based on free energy change before and after the
reaction. However, modeling pharmaceuticals, especially ibuprofen, with •OH through
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the DFT method has not been explored sufficiently as indicated by the lack of data in the
literature.
The first step of the oxidation reaction between ibuprofen and •OH in water was
modeled using DFT. Ibuprofen was selected for study because it is a commonly detected
pharmaceutical in aquatic environments [5, 27]; and experimental evidence exists which
can be compared to calculated results; finally ibuprofen, containing 33 atoms, is a
reasonable target compound for the DFT method. Both thermodynamic and kinetic
behavior of the possible reaction pathways were modeled, providing insights into the
thermodynamics of product formation, reaction energy profiles and the potential energy
surface, branching ratio, and ibuprofen degradation rate constant on the molecular level.
A.3 Methodology
The structure of ibuprofen is illustrated in Figure A.1 and the physiochemical
properties regarding its environmental behavior and fate are summarized in Table A.1.
Geometry optimizations and transition state searches were performed using
Gaussian 09 (Revision A.01) [28]. The geometries of the reactants, transition state (TS)
species, and products were optimized at the B3LYP/6-31G* level. Single-point energies
were calculated for the optimized B3LYP/6-31G* geometries with the same functional
and a more flexible basis set, i.e. B3LYP/6-31+G**. The solvent (water) effect on the
reaction was modeled with the integral equation formalism (IEF) version of the
polarizable continuum model (PCM) [29] as part of the B3LYP/6-31+G** single-point
energy calculation. The <S2> values for the various TS structures typically show minimal
spin contamination, ranging from 0.75 to 0.80. The nature of all stationary points, either
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minima or transition states, was determined by calculating the vibrational frequencies at
the B3LYP/6-31G* level. Intrinsic reaction coordinates (IRC) calculations were
performed for all TS species to confirm that they connect to the anticipated reactants and
products on the potential energy surface. The thermal and entropic contributions to the
free energies were also obtained from the B3LYP/6-31G* vibrational frequency
calculations, using the unscaled frequencies.
The likely sites of reaction can be based on the free radical stability of the resulting
carbon radical. In the case of hydrogen abstraction, the benzylic position (C11 and C24 in
Figure A.1) form stable carbon radicals and would be expected to be highly reactive to
•OH. Tertiary carbon radicals are also stable so H abstraction at C14 could be expected.
C16, C20, and C24 are all primary carbons and do not form stable radicals so H
abstraction is not as likely at these positions.
In the case of •OH addition, this is most likely to occur at the aromatic ring with its
“high” electron density. Addition forms a resonance stabilized carbon radical in the ring
and could be expected to be a favorable process. Since the two faces of the benzene ring
are asymmetric the thermodynamics of addition to both faces needs to be determined. In
Figure A.1, the “top” or upper face of the benzene ring is anti to the carboxylic acid
hydroxyl group and represents an “anti” addition, while addition to the syn face (or
bottom face in Figure A.1) represents “syn” additions.
Finally the carboxylic acid group is susceptible to two modes of attack. First the
•OH could react with the acidic proton which would lead to a decarboxylation reaction
ultimately leaving a carbon radical at C24. Second the •OH could react via a
228
“nucleophilic” attack on C30 resulting in the loss of carbonic acid and again leaving a
carbon radical at C24. All of these processes will be modeled and compared.
The possible pathways for the reactions of ibuprofen with •OH based on these three
mechanisms are illustrated in Figure A.2. These reactions represent the first step of the
oxidative degradation of ibuprofen. Subsequent steps could involve the addition of •OH
to the carbon radical or the loss of a second H atom and formation of a carbon-carbon
double bond. Multiple steps would lead to the oxidative products seen by Mendez-
Arriaga et al., [20] Zheng et al., [30] and Madhavan et al. [31].
The enthalpies (R
ΔH ) and free energies (R
ΔG ) of the reactions corrected by
thermal effects at 298 K were calculated using eqns. A.1 and A.2.
reactants
reactantscorected0
products
productscorrected0R)H(E)H(E(298K)ΔH (A.1)
reactants
reactantscorrected0
products
productscorrected0R)G(E)G(E(298K)ΔG (A.2)
where E0 is the total electronic energy at T = 0 K, Gcorrected and Hcorrected are the thermal
correction to Gibbs free energy and enthalpy, respectively.
For the transition state for each reaction pathway, the relative activation energy, ΔG‡,
was calculated as indicated in eqn. A.3
ΔG‡
= reactants
reactantscorrected0TScorrected0)G(E)G(E (A.3)
229
The energy of activation is the minimum amount of energy needed for a reaction to
proceed.
Conventional transition state theory (TST) was employed to predict the rate of
reaction (M-1
s-1
) between ibuprofen and •OH at 298 K in water through eqn. A.4
/RT)ΔGexp(h
TkΓ(T)k(T)
‡
0
B (A.4)
where h is Planck’s constant, kB is Boltzmann’s constant, and ‡
0ΔG is the kinetic energy
barrier. Γ(T), temperature-dependent factor, corresponds to quantum mechanical
tunneling and is approximated by the Wigner method, eqn. A.5 [32].
2i ]T
)[1.4424
1(1Γ(T) (A.5)
where vi is the imaginary frequency, whose vibrational motion is the direction of the
reaction.
A.4 Result and discussion
A.4.1 Thermodynamics
The free energy and enthalpy of the reactions were calculated at the B3LYP/6-
31+G**//B3LYP/6-31G* level of theory using the IEFPCM water model. The enthalpies
and free energies of the reactions for the three different mechanisms, •OH addition, H
abstraction, and nucleophilic attack, are tabulated in Table A.2. All of the reactions are
230
thermodynamically favorable processes (R
ΔG < 0). In addition, all the reactions are
exothermic (R
ΔH < 0). The activation energy ΔG‡, imaginary vibrational frequencies,
rate constants and branching ratios are tabulated in Table A.3. Figure A.3 plots the
relative TS and product energies for all reactions.
A.4.1.1 •OH Addition
For •OH addition, the enthalpy change ranges from -18.17 to -11.71 kcal/mol, with a
median of -13.07 kcal/mol while the free energy change ranges from -4.99 to -1.76
kcal/mol. The difference can be attributed to a large negative change in entropy as the
reaction proceeds from two reactants to a single product.
The activation barrier ranges from 7.58 to 11.56 kcal/mol (average: 9.67 kcal/mol)
with the exception of C4 “syn” which has an activation barrier of only 2.08 kcal/mol.
This very low activation barrier results from an intermolecular hydrogen bond which
forms with the acid hydroxyl group as the •OH approaches the C4 carbon atom, see
Figure A.4. The hydrogen bond forms before the TS is reached and lowers the barrier
significantly, approximately 7.6 kcal/mol compared to other additions to either face of the
aromatic ring. The product for the addition at C4 “syn” also maintains a hydrogen bond
resulting in a significantly lower free energy, -7.23 kcal/mol compared to an average of -
3.17 kcal/mol for the other addition products. Of the eight possible addition reactions,
addition to C4 “syn” is clearly the most favorable either under kinetic or thermodynamic
conditions. However all addition reactions are less favorable than the other reaction
pathways due to the loss of aromaticity in the addition products.
231
A.4.1.2 H Abstraction
The hydrogen abstraction reactions can produce two different benzylic carbon
radicals at C11 and C24 (H12, H13; and H25 respectively), one tertiary radical C14 (H15)
and abstraction of the acidic hydrogen (H32). The benzylic radicals have an enthalpy
change ranging from -35.31 to -32.08 kcal/mol (average -33.12 kcal/mol) and free energy
change from -36.11 to -33.50 kcal/mol (average -34.40 kcal/mol). The tertiary radical has
an enthalpy of -23.68 kcal/mol and free energy change of -26.14 kcal/mol. The acidic
hydrogen abstraction has an enthalpy of -17.32 kcal/mol and free energy change of -
18.46 kcal/mol. In contrast to the •OH addition reactions, there is a more favorable
entropy change for H abstraction since the number of molecules/radicals does not change.
The activation barrier ranges from 6.15 to 8.07 kcal/mol (average 7.04 kcal/mol) for
the benzylic radicals. After extensive searches TS geometries could not be located for
either the tertiary radical or for the abstraction of the acidic proton at this level of theory.
The implication is that while the formation of the benzylic radicals are under kinetic
control, and would be faster than the OH addition reactions, the tertiary and acidic proton
abstractions are barrierless in terms of enthalpy and the reaction will proceed as soon as
the •OH encounters an ibuprofen molecule in solution.
All hydrogen abstraction products are more thermodynamically stable than the
addition products; along with their lower activation barriers the abstraction reactions
would dominate over addition reactions.
Subsequence secondary reaction of the C24 carbon radical would lead to
decarboxylation and secondary oxidation leading to a ketone group. This product, and its
232
further degradation products have been detected as oxidation products of ibuprofen using
sonolysis [31], gamma irradiation [30], and the photo-Fenton process (AOPs) [20].
A.4.1.3 Reaction at the Carboxylic acid
The final mode of interaction studied was the “nucleophilic” attack of •OH on the
carboxylic acid carbon, C30. This mode results in the formation of carbonic acid and
leaves a carbon radical at C24, see Figure A.2 c. The enthalpy for this reaction is -35.41
kcal/mol with a free energy change of -29.48 kcal/mol. Again, after an extensive search it
proved impossible to locate a stable TS complex at this level of theory. Again this would
imply that this reaction is barrierless in terms of enthalpy and the reaction will occur as
soon as a •OH encounters an ibuprofen molecule in solution.
Subsequence secondary reaction of the carbon radical would lead to hydroxylation
followed by additional rounds of oxidation leading to a ketone, similar to the H
abstraction of H32.
A.4.2 Comparison to experiment
The formation of hydroxylated ibuprofen has been detected in several different
AOPs systems, such as sonochemical, photochemical, and photochemical-catalytical
degradation of ibuprofen. Madhavan et al. [31], sonolyzed ibuprofen at the frequency of
213 kHz with a power output of 55 W/L and identified the structure of sonochemical
byproducts with electrospray ionization mass spectrometric technique. They found
sonochemical degradation of ibuprofen formed mono- and di-hydroxylated intermediates,
thus corroborating our results that the major and first step of ibuprofen reacting with •OH
233
is to undergo H-atom abstraction and then an additional •OH reacts with the ibuprofen
radical. Although they did not report the abundance of the decarboxylation byproduct 4-
isobutylacetophenone, its formation has been detected, suggesting that nucleophilic
attack is also thermodynamically favorable. Mendez-Arriaga et al. studied the
photochemical degradation of ibuprofen with TiO2 and also found the formation of
hydroxylated intermediates [33]. They indicated that the hydroxylation process can be the
first step followed by demethylation or decarboxylation. However based on our
calculations, both decarboxylation and hydroxylation processes via a carbon radical
intermediate are equally thermodynamically favorable.
The •OH addition to ibuprofen is thermodynamically less favorable than H
abstraction or nucleophilic attack as can clearly be seen in Table A.2. The R
ΔG and
RΔH for •OH addition is substantially less negative or exothermic than either abstraction
of nucleophilic attack. One would assume from these trends that the •OH addition should
occur less frequently than either abstraction or nucleophilic attack. This is confirmed by
experiment where only a two •OH addition products was detected [31], both
corresponding to addition ortho to the propanoic acid group, likely via a C4 hydrogen
bonded mechanism, see Figure A.4.
A.4.3 Kinetics
The transition state for each reaction pathway was located at the B3LYP/6-31G*
level of theory and the energy of the TS species was calculated at the B3LYP/6-31+G**
level of theory. Relative ΔG‡, rate constants
and imaginary frequency for the TS species
involved in the reactions of ibuprofen with •OH in aqueous solution are tabulated in
234
Table A.3. The ΔG‡ for TS species the ranges from 0.00 to 11.56 kcal/mol, with a median
9.33 kcal/mol. Energy profiles for the potential energy surface of the •OH addition, H
abstraction and nucleophilic attack reactions are plotted in Figure A.3. Figure A.3 shows
that the energy barrier height for H abstraction and nucleophilic attack pathways are
lower than that of •OH addition, indicating the rate constants for these two modes is
generally faster than •OH addition. Calculated rate constants for H abstractions where the
TS could be located range from 8.57 × 106 M
-1s
-1 to 1.95 × 10
8 M
-1s
-1 (average 9.14 × 10
7
M-1
s-1
) compared to 9.80 × 104 M
-1s
-1 to 1.75 × 10
7 M
-1s
-1 (average 3.00 × 10
6 M
-1s
-1) for
the addition reactions. The C4 syn addition has an anomalously high TST rate constant of
1.87 × 1011
M-1
s-1
. This is likely due to limitations both in TST theory which fails for
reactions with low barrier highs [39] and limitations in the calculation. Explicit solvation
of the acid group was not included which would have to be desolvated for the hydrogen
bond to form. In this case collision or diffusion theory would give a better approximation,
although likely an over estimation of the rate constant.
After an exhaustive search the TS species for H abstraction occurring at H15 and
H32 and nucleophilic attack at C30 could not be located at B3LYP/6-31G* level of
theory, implying that the energy barrier for •OH reacting with ibuprofen at these
positions is zero and hence the reaction is controlled by the diffusion of reactants in
aqueous solution.
Since the reaction is diffusion-controlled, TST is not applicable. Thus, the simplified
version of the Roman Smoluchowski equation was used to predict the rate constant [34],
calculated in eqn. A.6
235
)D)(Dr(rπ4NkIBUOHIBUOHAD
(A.6)
where NA is the Avogadro constant; OH
r and IBU
r are radii for •OH and ibuprofen,
respectively. DOH and DIBU are the diffusion coefficients for •OH and ibuprofen,
respectively. The second order rate constant for the reaction is calculated to be 7.4 × 109
M-1
s-1
which is comparable to the experimental value 6.5 ± 0.2 ×109 M
-1 s
-1 [18].
The branching ratios, referring to the contribution of each reaction pathway to the
overall degradation, for the products through different pathways are also tabulated in
Table A.3. H abstraction at H15 and H32, syn addition at C4 and nucleophilic attack on
C30 are equally dominant compared to other positions, corroborating that decarboxylated
and hydroxylated products were experimentally detected in the reaction between
ibuprofen and •OH [20, 30-31]. The syn addition ortho to the carboxylic acid group is
observed experimentally, but as noted the rate constant is likely an over estimation as
explicit solvation was ignored in the calculation due to the complexity it added.
A.5 Conclusion
In this study, the first step of the oxidation reaction between ibuprofen and •OH in
water was calculated. We proposed three possible pathways for the reactions of ibuprofen
with •OH, •OH addition, H abstraction, and nucleophilic attack on carbonyl group. The
reactions are generally controlled by thermodynamics, more specifically enthalpy. The
hydroxylation at the benzylic and tertiary carbon atoms and decarboxylation processes
are the dominant degradation pathway. Based on diffusion theory, the calculated second
order rate constant of ibuprofen with •OH is in agreement with the experimental value.
236
Our results theoretically support that AOPs, as tertiary treatment technology in water and
wastewater treatment plants, has a great potential to remove ibuprofen in waters.
237
Table A.1: Physiochemical properties of ibuprofen
use class anti-inflammatory
solubility (mg L-1
) 21 a
log Henry’s law constant at 25 °C (atm-m3 mol
-1) -6.82
b
true log Kow 3.97 c
vapor pressure (mm Hg) 0.0756 d
pKa 4.90 e
second-order •OH rate constant (M-1
s-1
) 6.5 ± 0.2×109, f
a: [35]; b: estimated value [36]; c: [37]; d: [36] ; e: [38]; and f: [18].
238
Table A.2: R
ΔG and R
ΔH (kcal/mol) for the products involved in the reactions of
ibuprofen with •OH in aqueous solution
reaction mechanism Position/atom number R
ΔG (kcal/mol) R
ΔH (kcal/mol)
•OH addition
C5 anti -3.76 -13.81
C5 syn -4.99 -14.80
C4 anti -2.48 -12.33
C4 syn -7.23 -18.17
C2 anti -1.76 -11.83
C2 syn -4.12 -14.70
C1 anti -2.10 -11.71
C1 syn -2.99 -12.12
H abstraction
H12 (C11) -33.50 -32.08
H13 (C11) -33.58 -32.04
H15 (C14) -26.14 -23.68
H25 (C24) -36.11 -35.31
H32 (O31) -18.46 -17.32
nucleophilic attack C30 -29.48 -35.41
239
Table A.3: ΔG‡ (kcal/mol) and imaginary frequency for the transition state species
involved in the reactions of ibuprofen with •OH in aqueous solution.
a: TS does not exist at B3LYP/6-31G* level of theory; N.A: not available.
b: see section 3.3 of the text.
c: these values calculated by diffusion theory.
240
Figure A.1: Optimized geometry of ibuprofen at B3LYP/6-31+G** level of theory with
IEFPCM water model. The “syn” face is on the bottom of this diagram and the “anti”
face the top.
241
Figure A.2: Possible pathways for the reactions of ibuprofen with •OH based on (a) H
abstraction, (b) •OH addition, and (c) nucleophilic attack.
242
Figure A.3: Profile of the potential energy surface (a. •OH addition pathway; b. H
abstraction and nucleophilic attack pathways) for the reaction between ibuprofen and
•OH at the B3LYP/6-31+G** level of theory.
243
Figure A.4: Product of “syn” addition to C4 with hydrogen bonding.
244
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