Photochemical Transformation of Three Polycyclic Aromatic Hydrocarbons, Ibuprofen, and Caffeine in Natural Waters

Home , Formic acid, Hydroxyl radical, Phenanthrene

PHOTOCHEMICAL TRANSFORMATION OF THREE POLYCYCLIC AROMATIC HYDROCARBONS, IBUPROFEN, AND CAFFEINE IN NATURAL WATERS

DISSERTATION

Presented in Partial Fulfillment of the Requirements

for the Degree Doctor Philosophy in the

Graduate School of The Ohio State University

Laura Elizabeth Jacobs, M.S.

The Ohio State University

2008

Approved by

______

Dissertation Committee: Advisor

Professor Yu-Ping Chin, Advisor Approved by

Professor Linda K. Weavers ______

Professor Terry L. Gustafson Advisor

Professor Harold Walker Environmental Science Graduate Program

ABSTRACT

The photolysis of three polycyclic aromatic hydrocarbons (PAHs), pyrene, phenanthrene, and naphthalene were studied in waters taken from Gary, Indiana (GIN) and Wilmington, North Carolina (WNC). Direct photolysis of PAHs was observed with pyrene degrading at a faster rate than either phenanthrene or naphthalene. When compared to direct photolysis, phenanthrene degradation increased in GIN water, but decreased in the WNC water due to higher levels of dissolved organic carbon (DOC) for

WNC (9.29 mg/L vs 6.73 mg/L for GIN) and less nitrate (0.046 mM vs 0.205 mM) for

GIN. The slightly lower rate of phenanthrene degradation in WNC water, corrected for light attenuation effects, is statistically the same as the direct photolysis experiments. We attribute the lower rate of degradation in the presence of WNC water to light screening by

DOC, while we believe that the faster reaction rate observed for GIN is the result of nitrate generated hydroxyl radical (OH•) chemistry. Overall photo-reaction rates decrease for the lower molecular weight PAHs as the fastest naphthalene photolytic rate was roughly two orders of magnitude slower than the photolysis of pyrene.

The photolysis of ibuprofen and caffeine was studied in solutions of fulvic acid isolated from Pony Lake, Antarctica (PLFA); Suwannee River, GA (SRFA); and Old

Woman Creek Natural Estuarine Research Reserve, OH (OWCFA). At 10µM initial concentration ibuprofen and caffeine degrade slowly by direct photolysis, but we

ii observed enhanced photodegradation in solutions of each fulvic acid. Quenching studies suggest OH• plays a prominent role in both caffeine and ibuprofen photolysis.

Spectroscopic techniques reveal the formation of multiple hydrophobic photo-products upon photolysis of ibuprofen, the dominant byproduct identified as 1-(4- isobutylphenyl)ethanol and a minor derivative isobutylacetophenone. Caffeine and ibuprofen photolysis reactions proceed even more quickly in fulvic acid solutions (6 mg/L DOC) at lower, more environmentally relevant concentrations (0.1 µM) where presumably reaction kinetics are controlled by both short and long lived reactive species.

When probing the responsible reactive transients under suboxic conditions, fulvic acid mediated photolysis of caffeine and ibuprofen slows suggesting the influence of an oxygen dependent long lived radical (peroxyl or phenoxyl radicals) playing a role at

0.1µM.

iii

DEDICATION

To Chris.

For Dad.

ACKNOWLEDGMENTS

I would like to thank my advisors Yu-Ping Chin and Linda K. Weavers for their belief my abilities. It has been a pleasure working with them both over the past four years. A special thanks to Ryan L. Fimmen and Heath Mash. Both the ibuprofen work and my education would not be what it is without their input and patience. I thank the additional members of my PhD committee for their intellectual support, Terry L.

Gustafson and Harold Walker.

For technical support I thank Tanya Young and Kathy Welch. For unlimited field assistance I thank Old Woman Creeks senior research scientist Dave Klarer. Additional colleagues providing intellectual support include Kristopher McNeill, Eric Weber, Beate

Escher, Richard Zepp, Silvio Canonica, William J. Cooper and Kristin Schirmer.

Additionally, I am indebted to the Chin and Weavers Research Groups, whose support, friendship, and collaboration made this undertaking possible.

Most importantly, I would like to thank Chris, my father, mother, sister, and Ale. I love each of you dearly.

Finally, I would like to thank the sponsors of this research: National Science

Foundation Grant BES-0504434 and NOAA’s National Estuarine Research Reserve

System Graduate Research Fellowship (Laura E. Jacobs).

VITA

B.A. Geology, Clemson University…………………………………...………………2000

M.S. Geology, Vanderbilt University…………………………………………………2003

Graduate Research Associate, The Ohio State University……………….……..2004-2005

Graduate Research Fellow, The Ohio State University………………………...2005-2008

FIELD OF STUDY

Major Field: Environmental Science

TABLE OF CONTENTS

Page

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

Dedication………………………………………………………………………….. iv

Acknowledgements………………………………………………………………… v

Vita…………………………………………………………………………………. vi

List of Tables………………………………………………………………………. x

List of Figures……………………………………………………………………… xi

Chapters:

1. Introduction……………………………………………………………………… 1 1.1 Nature of Scope of Research…………………………………………… 1 1.2 Photochemistry in Natural Waters……………………………………… 2 1.3 Dissolved Organic Matter………………………………………………. 5 1.4 Objectives of This Dissertation………………………………………… 7 1.5 References……………………………………………………………… 9 1.6 Figures…………………………………………………………………. 12

2. Direct and Indirect Photolysis of Polycyclic Aromatic Hydrocarbons in Nitrate-rich Surface Waters………………………………………………………... 15 2.1 Introduction…………………………………………………………….. 15 2.2 Methods………………………………………………………………… 18 2.2.1 Materials……………………………………………………... 18 2.2.2 Photolytic Reactions…………………………………………. 18 2.3 Results and Discussion…………………………………………………. 20 2.4 Conclusions…………………………………………………………….. 27 2.5 References……………………………………………………………… 29 2.6 Tables…………………………………………………………………... 32

vii 2.7 Figures…………………………………………………………………. 33

3. Ibuprofen Photolysis in the Presence of Three Fulvic Acids…………………… 38 3.1 Introduction……………………………………………………………. 38 3.2 Methods……………………………………………………………….. 41 3.2.1 Chemicals and Fulvic Acids…………………………………. 41 3.2.2 Photolytic Reactions………………………………………… 41 3.2.3 Natural Sunlight Experiment………………………………... 43 3.2.4 HPLC……………………………………………………….. 43 3.3 Results and Discussion………………………………………………... 44 3.3.1 10 µM Photolysis Experiments……………………………… 44 3.3.2 0.1 µM Photolysis Experiments…………………………….. 47 3.4 Conclusions………………………………………………………….... 49 3.5 References…………………………………………………………….. 51 3.6 Tables…………………………………………………………………. 53 3.7 Figures………………………………………………………………… 54

4. Ibuprofen Photolysis: A Byproduct Analysis and Proposed Chemical Mechanism……………………………………………………………………….. . 59 4.1 Introduction………………………………………………………….... 59 4.2 Methods……………………………………………………………….. 60 4.2.1 Chemicals………………………………………………..….. 60 4.2.2. Photolytic Reactions……………………………………….. 61 4.2.3 LC-MS…………..………………………………………….. 61 4.3.4 GC-MS……………………………………………………... 59 4.3.5 1H and COSY NMR………………………………………... 62 4.3 Results and Discussion………………………………………………… 63 4.4. Conclusion…………………………………………………………… 67 4.5 References……………………………………………………………. 68 4.6 Tables…………………………………………………………………. 69 4.7 Figures………………………………………………………………… 70 4.8 Schemes……………………………………………………………….. 81

5. Caffeine as a Wastewater Tracer: A Photochemical Analysis………………… 82 5.1 Introduction…………………………………………………………… 82 5.2 Methods………………………………………………………………. 85 5.1.1 Materials……………………………………………………. 85 5.2.2. Photolytic Reactions……………………………………….. 86 5.3 Results and Discussion……………………………………………….. 87 5.4 Conclusion……………………………………………………………. 92 5.5. References……………………………………………………………. 94 5.6 Figures………………………………………………………………… 97

6. Conclusions and Future Research……………………………………………… 103

viii 6.1 Conclusions…………………………………………………………. 103 6.2 Future Research…………………………………………………….. 107

Bibliography……………………………………………………………………. 109

Appendices…………………………………………………………………….. 117 A. Chemical Actinometry…………………………………………….... 117 A.1 Solar Simulator Chemical Actinometry……………………. 117 A.2 Natural Photolysis Chemical Actinometry……………….... 121

LIST OF TABLES

Table Page

2.1 Selected Gary, Indiana (GIN) and Wilmington, North Carolina (WNC) - water parameters (dissolved organic carbon (DOC), nitrate (NO3 ), total -1 iron (Fe)) and observed photolytic degradation rate constants (kobs h ) in Milli-Q water and our samples; (-) Denotes not applicable………………….. 32

3.1 10 µM and 0.1 µM racemic and S-(+) (when indicated) ibuprofen -1 degradation rate constants (kobs h ) in the presence of simulated and natural sunlight, OWCFA (Old Woman Creek Fulvic Acid), SRFA (Suwannee River Fulvic Acid), and PLFA (Pony Lake Fulvic Acid) at indicated dissolved organic carbon levels (DOC), in the absence of molecular oxygen, and in the presence of isopropanol. (–) Denotes not applicable…………… … 53

4.1 Gas Chromatography-Mass Spectrometry (GC-MS) dominant ion peaks (1-11) from 10 µM ibuprofen solution photolyzed 48 h confirming the presence of numerous byproducts in solution, several mw ~176. (-) Denotes not applicable………………………………………………………………. … 69

LIST OF FIGURES

Figure Page

1.1 Illustration of the dissolved organic matter source continuum using the two end members Suwannee River Fulvic Acid (SRFA) and Pony Lake Fulvic Acid (PLFA); and Old Woman Creek Fulvic Acid (OWCFA)……… 12

1.2 Aerial photo of Old Woman Creek National Estuarine Research Reserve, Huron, OH. Courtesy of Dave Klarer……………………………………….. 13

1.3 Chemical structures of chosen non-point source contaminants…………….. 14

2.1 Photoinduced degradation of pyrene in Milli-Q and Wilmington, North -1 Carolina (WNC) and Gary, Indiana (GIN) sample waters where kobs h is the observed rate constant per hour. Rate constants are corrected for light screening for both natural waters……………………………………………. 33

2.2 Photoinduced degradation of phenanthrene in Milli-Q and Wilmington, North Carolina (WNC) and Gary, Indiana (GIN) sample waters where -1 kobs h is the observed rate constant per hour. Rate constants are corrected for light screening for both natural waters………………………………….. 34

2.3 Light absorption of selected water samples: Gary, Indiana (GIN) water (0.2 mM nitrate), 0.2 mM nitrate solution, and Wilmington, North Carolina (WNC) water (0.04 mM nitrate)………………………………….. 35

2.4 Photoinduced degradation of phenanthrene in Wilmington, North Carolina water at natural nitrate levels (0.046 mM nitrate) and spiked nitrate levels -1 of 0.2 mM and 0.4 mM where kobs h is the observed rate constant per hour.  Wilmington water;  Wilmington water plus 0.2 mM nitrate;  Wilmington water plus 0.4 mM nitrate………………………………… 36

2.5 Photoinduced pseudo first order degradation (Ln(Concentration/Initial Concentration)) of pyrene in Gary, Indiana water (0.2 mM nitrate) and GIN water with 25 mM methanol as a hydroxyl radical scavenger.  Gary water; ♦ Gary water plus 25 mM methanol……………………… 37

xi 3.1 Pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) in Milli-Q (Direct) photolysis degradation and in the presence of Old Woman Creek Fulvic Acid (OWCFA) with and without isopropanol………………………………………………………………….. 54

3.2 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) direct photolysis degradation and in the presence of Suwannee River Fulvic Acid (SRFA) with and without isopropanol…………………………………………………………………. 55

3.3 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) direct photolysis and in the presence of Pony Lake Fulvic Acid (PLFA) with and without 20 mM isopropanol……. 56

3.4 10 µM and 0.1 µM ibuprofen pseudo first order degradation rate constants -1 (kobs h ) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA)…………………………………………………………………… 57

-1 3.5 0.1 µM ibuprofen pseudo first order degradation rate constants (kobs h ) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) under oxic and suboxic systems and with 20 mM isopropanol (Iso)…………………... 58

4.1 a. UV-Vis scan of photolyzed solution at various time points (hours) demonstrating byproduct accumulation at 264 nm; b. HPLC chromatogram (λ = 264nm) of solution photolyzed for 48 hours (ibuprofen retention time ~5minutes;“major” photolysis byproduct retention time ~7 minutes, additional byproduct retention time ~8 minutes)…………………………………………………………………..... 70

4.2 1H 1-D NMR spectra of 500 µM ibuprofen……………………………….. 71

4.3 1H NMR of 10 µM photolyzed (48 h) ibuprofen solution………………… 72

4.4 Liquid Chromatography Electrospray Ionization Quadropole Time of Flight Mass Spectrometer (LC-ESI-QTOF-MS) of 10 µM ibuprofen solution photolyzed 48 h and proposed byproduct structures…………...... 73

4.5 Light absorption of 100 µM Isobutylaectophenone (IBAP) in Milli-Q…… 74

4.6 1H NMR of purchased isobutylacetophenone standard (100 µM)……….. 75

4.7 1H 1-D NMR spectra of isolated major byproduct fraction of photolyzed ibuprofen solution and proposed byproduct structure……………………. 76 xii

4.8 1H NMR COSY of HPLC waste line effluent (retention time 7 minutes) of photolyzed ibuprofen solution (48 h)……………………………………… 77

4.9 1H NMR COSY of 500 µM purchased isobutylacetophenone…………….. 78

4.10 Formation of major byproduct (retention time 7 minutes) measured in peak area units (HPLC with a UV detector) upon the photolysis of 10 µM ibuprofen (Direct) in the presence of Pony Lake Fulvic Acid (PLFA), Old Woman Creek Fulvic Acid (OWCFA), and Suwannee River Fulvic Acid (SRFA) with 20 mM isopropanol (Iso) and suboxic conditions.. 79

4.11 Formation of major byproduct (retention time 7 minutes) measured in peak area units (HPLC with a UV detector) upon the photolysis of 0.1 µM ibuprofen (Direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Old Woman Creek Fulvic Acid (OWCFA), and Suwannee River Fulvic Acid (SRFA) with 20 mM isopropanol (Iso) and suboxic conditions………. 80

5.1 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) in Milli-Q (direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and -1 Old Woman Creek Fulvic Acid (OWCFA) where kobs h is the observed rate constant per hour…………………………………………………………….. 97

5.2 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) in Milli-Q (Direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) with and without 20 mM isopropanol (Iso) and with visual trendlines…………………………………. 98

5.3 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) photolysis in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) with and without various levels of nitrate (NO3)………………………………………………………………… 99

5.4 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) photolysis in the presence of Old Woman -1 Creek Fulvic Acid (OWCFA) with and without 31 µM iron where kobs h is the observed rate constant per hour………………………………………. 100

5.5 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM; 0.1 µM) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old -1 Woman Creek Fulvic Acid (OWCFA) where kobs h is the observed rate xiii constant per hour…………………………………………………………… 101

5.6 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (0.1 µM) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) under oxic and suboxic conditions where -1 kobs h is the observed rate constant per hour……………………………… 102

A.1 Chemical actinometry (p-nitroanisole (PNA)/pyridine system) for the Altas CPS+ Solar Simulator, June 2006…………………………………… 117

A.2 Chemical actinometry (p-nitroanisole (PNA)/pyridine system) for the Altas CPS+ Solar Simulator, April 2007………………………………….. 118

A.3 Chemical actinometry (p-nitroanisole (PNA)/pyridine system) for the Altas CPS+ Solar Simulator, July 2007…………………………………... 119

A.4 Chemical actinometry (p-nitroanisole (PNA)/pyridine system) for the Altas CPS+ Solar Simulator, April 2008…………………………………. 120

A.5 Chemical actinometry (p-nitroanisole (PNA)/pyridine system) at Old Woman Creek National Esturaine Research Reserve, Huron, OH, June 2006…………………………………………………………………. 121

xiv

LIST OF SCHEMES

Schemes Page

4.1 Proposed reactions 1) Ibuprofen photodecarboxylation 2) Oxygen addition to carbon centered radical and subsequent rearrangement resulting formation of isobutylacetophenone 3) Hydroxylation of carbon centered radical………………………………………………….. 81

CHAPTER 1

INTRODUCTION

1.1 Nature and Scope of Research

In 1999-2000 Kolpin et al., produced a nationwide reconnaissance of non-point source (NPS) contaminants that potentially threaten the health of our surface water. This study joined others (Ternes, 1998; Daughton and Ternes, 1999) documenting numerous contaminants such as antibiotics, other prescription drugs, non-prescription drugs, steroids, reproductive hormones, pharmaceuticals and personal care products (PPCPs), and products of oil use and combustion which labeled these contaminants as “emerging contaminants” in natural waters. The presence of these emerging contaminants in the environment can be traced to wastewater treatment systems, septic systems, regulated and unregulated industrial discharges, large-scale animal feeding operations, combined sewage overflow (CSO), urban runoff and other non-point source pathways (Miller,

2001). These compounds are currently gaining much attention from the public and the scientific community as a need arises to understand the occurrence, bioavailability, fate, and ultimate impact on environmental systems (CNN Special Report, March 10, 2008).

1 Non-point source contamination, by definition, is difficult to control and can be widespread. Thus, a holistic watershed approach is needed to ameliorate this problem.

One such approach, wetland remediation, involves large-scale collection and diversion of runoff followed by treatment of the surface waters through natural processes i.e. attenuation, microbial degradation, redox active processes in sediments, phytoprocesses, and photochemistry (Gray and Sedlak 2005; Mehrotra and Sedlak, 2005; Miller and Chin

2002; Miller and Chin 2005; Hakala et al., 2007). Because of this passive, relatively inexpensive, and potentially effective treatment ability, interest regarding the preservation of natural wetlands and engineering new wetlands is increasing.

1.2 Photochemistry in Natural Waters

Photochemical processes are initiated when a particular constituent absorbs a photon, an ability directly related to the constituents structure. Molecular moieties that absorb photons are defined as chromophores, and include functional groups such as alkenes, carbonyls, and aromatics. An ultraviolet-visible light (UV-Vis) scan measures the ability of a constituent to absorb light as a function of wavelength. When a chromophore bearing organic molecule directly adsorbs light and is subsequently transformed, it is defined as direct photolysis (Schwarzenbach, 2003; Leifer, 1988; Zepp and Cline 1977).

Furthermore, naturally present constituents defined as photosensitizers can absorb light and produce reactive transients that subsequently react and transform/degrade other organic molecules, a process defined as indirect photolysis (Schwarzenbach, 2003;

Leifer, 1988; Zepp and Cline 1977).

2 Examples of photosensitizers include dissolved organic matter (DOM) and nitrate

(NO3), which can produce a cocktail of reactive transients capable of degrading organic molecules. Specifically, photo-excited DOM or 1DOM can undergo inner system crossing

(ISC) to a more stable, lower energy state, 3DOM, which can 1) react directly with contaminants (Gerecke et al., 2001), 2) generate hydroxyl radical (OH•) (Vaughn and

Blough; 1998), or 3) transfer energy to dioxygen generating other reactive transients

1 - singlet oxygen ( O2), hydrogen peroxide (H2O2), superoxide (O2• ), and peroxyl radicals

(ROO•) (Hoigne et al., 1989; Canonica and Freiburghaus 2001; Buschmann et al.,

2005a/b; Cooper et al., 1989; White et al., 2003; White, 2000; Haag and Hoigne 1986).

Taken together, these radicals (listed as 2 and 3) are called reactive oxygen species

- (ROS). Superoxide (O2• ) can then continue to react with itself to produce H2O2 in a process called dismutation (Cooper and Zika 1983). H2O2 generated in this fashion can futher react in the presence of slightly acidic iron bearing natural waters to produce OH•; a reaction deemed the photo-Fenton pathway:

- Fe(II) + H2O2 → Fe(III) + OH + OH•

(Southworth and Voelker, 2003). If nitrate is present in sufficient quantities, OH• is also produced via:

- -* - 3 NO3 + light → NO3 → NO2 + O( P)

- → NO2 + •O

- - •O + H2O → OH• + OH

(Zepp at al., 1987). Finally, irradiated DOM itself is also responsible for producing OH•, but by pathways that have not been well defined (White, 2000). These reactions result in

3 a steady-state concentration of reactive transients generated by irradiation of natural waters that have the ability to degrade contaminants in the water column.

The actual concentration of these transients remains low and is a function of both the type and amount of different photosensitizers present and the presence of scavengers responsible for keeping their levels in check. Many authors have probed the reaction kinetics of these transients with drinking water and surface water contaminants (Haag and

Yao 1992; Leifer, 1988; Schwarzenbach et al., 2003). The viability of photochemical degradation as an attenuation mechanism in surface waters or wetlands is typically deemed favorable (Miller and Chin, 2002; Miller and Chin 2005; Boreen et al., 2003).

But others have also identified limitations to this approach in the form of photo- derivatives (Edhlund et al., 2006; Boreen et al., 2005; Pérez-Estrada et al., 2005, and others). Finally, the environmental impact of these photo-derivatives must also be taken into account.

Photolysis in natural waters occurs in upper regions of the water column (photic zone) and is affected by depth and attenuation. Wetlands tend to have higher surface photoreactivity than other waters (Valentine and Zepp, 1993) generating a higher probability for successful remediation through photochemical reactions. Moreover, the presence of abundant photosensitizers (DOM, nitrate) in wetland waters further enhances the ability of these water bodies to degrade organic contaminants. Conversely, DOM can also absorb reactive photons in the water column, a phenomenon known as light screening and scavenge reactive transients (Torrents et al., 1997; Brezonik and

Fulkerson-Brekken, 1998). The turbidity of the water column will also affect the efficacy

4 of photo-remediation as suspended materials (phytoplankton, clays) limits light penetration.

When comparing photodegradation conducted using simulated sunlight to those that occur in the natural environment, correction factors, defined in the following chapters, account for difference in light source. Further, one must take into account diurnal cycles, sun angle (latitude specific), and other variables such as cloud cover, which is a more difficult variable to constrain. A corrector factor must be utilized to account for DOM light screening within reactor vessels if solutions are optically dense (Miller and Chin,

2002; Miller and Chin, 2005). Thus, all rate constants reported here represent photochemical reactions occurring on the surface of the water column (~ 1 cm) as depth and light attenuation are not taken into account. Furthermore, I did not account for diurnal cycles therefore reported results represent photodegradation under the most optimistic conditions ie. constituents are irradiated by a light source that is several times more intense than noon sunlight at 40° N in June.

1.3 Dissolved Organic Matter

Dissolved organic matter is a complex, reactive substance that is present ubiquitously in surface waters. It is the most abundant pool of organic matter on Earth and varies in composition both temporally and spatially (Hedges and Oades, 1997). In surface water, organic matter is derived from plant, animal, and microbial sources. In order to understand the range of reactions this substance can produce; I choose to study the photoreactivity of two end members which anchor opposite ends of a ‘source continuum’ where most aquatic DOM composition is an intermediate between the two materials 5 (Figure 1.1). The first is a fulvic acid from Suwannee River, GA, USA a blackwater river sourced in the Okefenokee Swamp deriving its organic matter from decaying terrestrial plants. The second fulvic acid is from Pony Lake, Antarctica a eutrophic lake whose dissolved organic matter is from microbial matter i.e. non-lignin precursors.

When interested in the role of dissolved organic matter in water chemistry, it is often necessary to extract this substance from natural waters. Different extraction techniques target different portions of the DOM pool. Resin adsorption methods isolate humic substances, defined accordingly as the portion of the DOM pool that is soluble in aqueous base (Schwarzenbach, 2003; Leenheer, 1981; Thurman & Malcolm, 1981). Further subdividing and isolating the DOM pool with this method, fulvic acid is the portion soluble in both basic and acidic solutions, humic acid is the portion insoluble in acidic solutions but soluble at high pHs (Schwarzenbach, 2003; Leenheer, 1981; Thurman &

Malcolm, 1981), and the water soluble hydrophilic neutrals (or carbohydrates) of DOM pass through the resin without acid/base elution. In this study, a non-ionic XAD-8 resin protocol is used (Leenheer, 1981; Thurman & Malcolm, 1981). This procedure is used to obtain the hydrophobic portion of the organic matter (soluble in both basic and acidic solutions), termed the fulvic acid (FA) fraction. As the more hydrophobic portion, fulvic acids are the more photochemically active fraction of organic matter allowing extrapolation from fulvic acids to natural waters.

6 1.4 Objectives of This Dissertation

I hypothesize that photochemical pathways in surface waters, specifically wetlands, can facilitate the photo-degradation of non-point source contaminants. I tested this hypothesis with the following objectives regarding the chosen contaminants:

• identify the principal photochemical degradative pathway (direct vs. indirect) in

chosen environmental matrices;

• isolating a large amount of dissolved organic matter from Old Woman Creek

National Estuarine Research Reserve, Huron, OH (Figure 1.2) in the form of

fulvic acids using the XAD-8 resin protocol for photolytic reactivity studies;

• probing the role of different sources of DOM using the above-mentioned fulvic

acid, deemed Old Woman Creek Fulvic Acid (OWCFA);

• compare OWCFAs efficacy as a photosensitizer to two International Humic

Substance Society (IHSS) fulvic acid standards: Pony Lake Fulvic Acid a

microbially derived material and Suwannee River Fulvic Acid a terrestrially

derived material;

• investigating the role of light screening by DOM by applying a correction factor

when necessary;

• probing the role of nitrate on the photodegradation of chosen contaminants using

the known scavengers;

• identifying and quantifying additional relevant reactive transients responsible for

indirect photolysis of the contaminants of interest through the use of varying

initial contaminant concentration, scavengers, and under suboxic conditions; and

7 • elucidating important photo-products derived from select reactions using

chromatographic and spectroscopic techniques

My NPS contaminants of choice represent commonly found pollutants that span a wide range of photoreactivities including those compounds that readily undergo direct photolysis and those that are less susceptible to this pathway. These include the polycyclic aromatic hydrocarbons (PAHs) pyrene, naphthalene, and phenanthrene, as well as ibuprofen and caffeine (Figure 1.3). The above objectives were achieved through a matrix of photochemical experiments with appropriate

- constituents/conditions (DOM, NO3 , scavengers, suboxic conditions, analyte concentrations, and statistical assays). Data were analyzed with appropriate reaction kinetics. When appropriate, reaction solutions were subjected to spectroscopic analysis to probe photo-derivatives and a transformation mechanism.

8 1.5 References

Boreen, A. L.; Arnold W. A.; McNeill K. (2003) Photodegradation of pharmaceuticals in the aquatic environment: A review. Aquat. Sci. 65, 320-341.

Boreen, A.L.; Arnold, W.A.; McNeill, K. (2005) Triplet-sensitized photodegradation of sulfa drugs containing six-membered heterocyclic gropuds: Identification of an SO2 extrutsion photoproduct. Environ. Sci. Technol. 39, 3630-3638.

Brezonic PL, Fulkerson-Brekken J. (1998) Nitrate-induced photolysis in natural waters: Controls on concentrations of hydroxyl radical photo-intermediates by natural scavenging agents. Environ Sci Technol 32: 3004-3010.

Buschmann, J.; Canonica, S.; Sigg, L. (2005a) Photoinduced oxidation of Antimony(III) in the presence of humic acid. Environ. Sci. Technol. 39, 5335-5341.

Buschmann, J.; Canonica, S.; Lindauer, U.; Hug, S.J.; Sigg, L. (2005b) Photoirradiation of dissolved humic acid induces arsenic(III) oxidation. Environ. Sci. Technol. 39, 9541- 9546.

Canonica, S.; Freiburghaus, M. (2001) Electron-rich Phenols for probing the photochemical reactivity of freshwaters. Environ. Sci. Technol. 35, 690-695.

Cooper, W.J; Zika, R.G. (1983) Photochemical Formation of Hydrogen Peroxide in Surface and Ground Waters Exposed to Sunlight. Science. 220, 711-712.

Cooper, W.J.; Zika, R.G.; Petasne, R.G.; Fischer, A.M. (1989) Sunlight Induced Photochemistry of Humic Substances in Natural Waters: Major Reactive Species. ACS Symposium Series. 219, 333-362.

CNN Special Report, Prescription Drugs Found in Drinking Water Across the U.S. March 10, 2008.

Daughton C.G.; Ternes, T.A. (1999) Pharmaceutical and personal care products in the environment: Agents of subtle change? Environ. Health Persp. 107, 907-938.

Edhlund, B.L.; Arnold, W.A.; McNeill, K. (2006) Aquatic photochemistry of nitrofuran antibiotics. Environ. Sci. Technol. 40, 5422-5427.

Gray, J.L.; Sedlak, D.L. (2005) The Fate of Estrogenic Hormones in an Engineered Treatment Wetland with Dense Macrophytes. Wat Environ Res. 77 (1), 24-31.

Gerecke AC, Canonica S, Muller SR, Scharer M, Schwarzenbach RP. (2001) Quantification of dissolved natural organic matter (DOM) mediated phototransformation of phenylurea herbicides in lakes. Environ. Sci. Technol. 35: 3915-3923. 9

Hakala, J.A.; Chin, Y-P.; Weber, E.J. (2007) Influence of Dissolved Organic Matter and Fe(II) on the Abiotic Reduction of Pentachloronitrabenzene. Environ. Sci. Technol. 41, 7337-7342.

Haag, W.R.; Hoigné, J. (1986) Photochemical Formation and Steady-State Concentrations in Various Types of Waters. Environ. Sci. Technol. 20, 341-348.

Haag, W.R.; Yao, C.C.D. (1992) Rate Constant for Reaction of Hydroxyl Radicals with Several Drinking Water Contaminants. Environ. Sci. Technol. 26, 1005-1013.

Hedges, J.I.; Oades, J.M. (1997) Comparative organic geochemistries of soils and marine sediments. Org. Geochem. 27, 7/8, 319-361.

Hoigne, J.; Faust, B.C.; Haag, W.R.; Scully, F.E.; Zepp, R.G. (1989) Aquatic Humic Substances and Sinks of Photochemically Produced Transients Reactants. ACS Symposium Series. 219, 363-381.

Kolpin, D.W.; Furlong, E.T.; Meyer, M.T.; Thurman, E.M.; Zaugg, S.D.; Barber, L.B.; Buxton, H.T. (2002) Pharmaceuticals, hormones, and other organic, wastewater contaminants in U.S. streams, 1999-2000: A national reconnaissance Environ. Sci. Technol. 36, 1202-1211.

Leenheer J.A. (1981) Comprehensive approach to preparative isolation and fractionation of dissolved organic carbon from natural waters and wastewaters. Environ. Sci. Technol. 15, 578-587.

Leifer A. 1988. The Kinetics of Environmental Aquatic Photochemistry. American Chemical Society, Washington, DC.

Mehrotra, A.S.; Sedlak D.L. (2005) Decrease in Net Mercury Methylation Rates Following Iron Amendment to Anoxic Wetland Sediments Sluries. Environ. Sci. Technol. 39, 2564-2570.

Miller, P. (2001) Photohemical transformation of agricultural pesticides promoted by natural water constituents in wetland surface waters. PhD Dissertation. The Ohio State University, Columbus OH, USA.

Miller, P.L.; Chin, Y.P. (2002) Photoinduced degradation of carbaryl in a wetland surface water. J. Agric. and Food Chem. 50, 6758-6765.

Miller PL, Chin YP. (2005) Indirect photolysis promoted by natural and engineered wetland water constituents: Processes leading to alachlor degradation. Environ Sci Technol 39: 4454-4462.

10 Pérez-Estrada, L.A.; Malato, S.; Gernjak, W.; Agüera, A.; Thurman, E.M.; Fernández- Alba, A.R. (2005) Photo-fenton degradation of diclofenac: Identification of main intermediates and degradation pathway. Environ. Sci. Technol. 39, 8300-8306.

Schwarzenbach RP, Gschwend PM, Imboden DM. 2003. Environmental Organic Chemistry, 2nd ed. John Wiley, Hoboken, NJ, USA.

Southworth, B.A.; Voelker, B.M. (2003) Hydroxyl radical production via the photo- Fenton reaction in the presence of eachfulvic acid. Environ. Sci. Technol. 37, 1130-1136.

Ternes, T.A. (1998) Occurrence of drugs in German sewage treatment plants and rivers. Wat. Res. 32, 3245-3260.

Torrents A, Anderson BG, Bilboulian S, Johnson WE, Hapeman CJ. (1997) Atrazine photolysis: Mechanistic investigations of direct and nitrate-mediated hydroxyl radical processes and the influence of dissolved organic carbon from the Chesapeake Bay. Environ Sci Technol 31: 1476-1482.

Thurman, E.M.; Malcolm, R.L. (1981) Preparative isolation of aquatic humic substances. Environ. Sci. Technol. 15, 463-466.

Valentine, R.L.; Zepp, R.G. (1993) Formation of carbon monoxide from the photodegradation of terrestrial dissolved organic carbon in natural waters. Environ. Sci. Technol. 27 (2), 409-412.

Vaughn, P.P.; Blough, N.V. (1998) Photochemical formation of hydroxyl radical by constituents of natural waters. Environ. Sci. Technol. 32, 2947-2953.

White, E. (2000) Determination of photochemical production of hydroxyl radical by dissolved organic matter and associated iron complexes in natural waters. M.S. Thesis. The Ohio State University, Columbus OH, USA.

White, E.M.; Vaughan, P.P.; Zepp, R.G. (2003) Role of the photo-Fenton reaction in the production of hydroxyl radicals and photobleaching of colored dissolved organic matter in a coastal river of the southeastern United States. Aquat. Sci. 65, 402-414.

Zepp, R.G.; Cline, D. M. (1977) Rates of Direct Photolysis in Aquatic Environment. Environ. Sci. Technol. 11, 4, 359-366.

Zepp, R.G.; Hoigné, J.; Bader, H. (1987) Nitrate-Induced Photooxidation of trace organic chemicals in water. Environ. Sci. Technol. 21: 443-450.

11 Figures 1.6

• • •

Figure 1.1 Illustration of the dissolved organic matter source continuum using the two end members Suwannee River Fulvic Acid (SRFA) and Pony Lake Fulvic Acid (PLFA); and Old Woman Creek Fulvic Acid (OWCFA).

Figure 1.2 Aerial photo of Old Woman Creek National Estuarine Research Reserve, Huron, OH. Courtesy of Dave Klarer.

Pyrene Phenanthrene Naphthalene

Ibuprofen Caffeine

Figure 1.3 Chemical structures of chosen non-point source contaminants

CHAPTER 2

DIRECT AND INDIRECT PHOTOLYSIS OF POLYCYCLIC AROMATIC HYDROCARBONS IN NITRATE-RICH SURFACE WATERS

2.1 Introduction

The environmental fate of polycyclic aromatic hydrocarbons (PAH) has been extensively studied because of their toxicity, persistence in the environment, and propensity to bioaccumulate (Schwarzenbach, 2003). Major anthropogenic sources of

PAHs include the combustion of fossil fuels, the release of petroleum products and coal tar, and the use of creosote as a wood preservative. The PAH concentrations in natural water are quite variable depending upon input sources (atmospheric deposition vs. direct discharge). Most of the environmental PAH burden remains near point sources and concentrations decrease with increasing distance from the source (Neff, 1979).

One important degradation pathway for PAHs in surface waters is direct photolysis (Miller and Olejnik, 2001; Chen et al., 2000; Lehto et al., 2000; Fasnacht and

Blough, 2002; Fasnacht and Blough, 2003a; Fasnacht and Blough, 2003b). The exact

PAH photolytic mechanism is controversial, with some investigators proposing a photoionization pathway to form a PAH radical cation and a hydrated electron (Chen et al., 2000; Zepp and Schlotzhauer, 1979), while others have suggested a reaction scheme 15 whereby the excited singlet or triplet state PAH can react with dioxygen. The latter pathway was deduced from the observed strong dependence of the photodegradation quantum yield on dissolved oxygen levels. The excited state PAH reaction with dioxygen can lead to recombination, the formation of a PAH cation radical and superoxide, or other oxygenated products formed by the collision complex (Fasnacht and Blough, 2003a;

Fasnacht and Blough, 2003b). Irrespective of pathway, the extent and quantum yields for this reaction are highly variable among different types of PAHs (Chen et al., 2000) with the more condensed aromatic ring compounds possessing greater reactivity due to their higher extinction coefficients at wavelengths present in sunlight (Bertilsson and

Widenfalk, 2002).

Several naturally occurring chemical constituents in natural waters are known to accelerate and/or retard PAH photolysis (Neff, 1979). Those substances capable of enhancing photoreactivity through the formation of reactive intermediates when irradiated are called photosensitizers. Two important photosensitizers that are ubiquitous in natural water bodies are dissolved organic matter (DOM) and nitrate. Both of these substances are known to indirectly enhance the photolysis of numerous organic compounds through the formation of chemical transients such as reactive oxygen species

(ROS) (Gerecke et al., 2001; Miller and Chin, 2002; Miller and Chin, 2005; Vaughan and

Blough, 1998). Important photogenerated reactants include the hydroxyl radical (OH•),

1 singlet oxygen ( O2) and photochemically excited states of DOM i.e., triplet DOM

(Gerecke et al., 2001; Canonica et al., 1995; Canonica and Freiburghaus, 2001).

Conversely, DOM can both screen reactive photons as well as scavenge ROS and other radicals (Torrents et al., 1997). In spite of the potentially high reactivity imparted to 16 natural waters by photo-excited DOM, work by Fasnacht and Blough (Fasnacht and

Blough, 2003a; Fasnacht and Blough, 2003b) has shown that it does not appear to play an important role in the photofate of PAHs.

Nitrate, when irradiated at wavelengths present in sunlight, is capable of forming

OH• (Zepp et al., 1987). Nitrate concentrations are typically low in surface waters and at these levels is not an important photosensitizer (Miller and Chin, 2002). In agricultural watersheds, however, nitrate levels can increase dramatically and remain at these levels for extended periods due to the extensive use of fertilizer. For example, nitrate levels in

Midwest, USA, surface waters range from 0.1 to greater than 1 mM (Miller and Chin,

2002; Miller and Chin, 2005; Brezonic and Fulkerson-Brekken, 1998). At these concentrations Zepp et al. (1987) showed that nitrate-generated OH• becomes an important ROS capable of degrading a number of contaminants. Thus, nitrate photolysis at these concentrations has been shown by a number of investigators to be the primary

OH• generated photosensitizer in natural waters (Miller and Chin, 2002; Miller and Chin,

2005; Torrents et al., 1997; Wilson and Mabury, 2000) and is much more important than

DOM. To date we are not aware, however, of any study that specifically examines the effect of nitrate present in natural waters on the photolytic fate of PAHs.

In the present study the photolysis of three PAHs, naphthalene, phenanthrene and pyrene, in water samples from two creosote-contaminated sites was studied. One site located near Gary, Indiana (41˚39’07”N, 87˚28’04”W) resides in a predominantly agricultural watershed (upper Indiana), while the second site is located in a diverse land use watershed near Wilmington, North Carolina (34˚12’56”N, 77˚57’07”W). The objective of the present study was to delineate the relative contribution of the direct and 17 indirect photolysis pathways to the photofate of the target PAHs in surface waters that contain different levels of nitrate and DOM.

2.2 Methods

2.2.1 Materials

Water samples from Gary, Indiana (GIN) and Wilmington, North Carolina

(WNC) were collected in December 2004 and acidified to a pH of approximately 2 for preservation. The sites were chosen based upon their proximity to PAH point sources

(both coal-tar based) and their contrasting water chemistries. Each sample was filtered to

0.45 µm, buffered using phosphate (5 mM), and adjusted to the measured field pH using

NaOH and HCl prior to photolysis. Both nitrate and dissolved organic carbon (DOC) in the natural water samples were respectively measured by ion chromatography and total organic carbon analysis (Shimadzu TOC-5000). Iron concentrations in our water samples were measured using the ferrozine method (Sottkey, 1970). Phenanthrene (Eastman

Kodak fluorescent grade), naphthalene (Aldrich 99%), and pyrene (Aldrich 99%) were used for the present study and represented a range of physicochemical properties from light absorption to aqueous solubility.

2.2.2. Photolytic Reactions

Photolysis of phenanthrene, pyrene, and naphthalene was conducted in Milli-Q

(18MΩ, Millipore) water (phosphate buffered as above) and in the two natural water samples using a solar simulator (Atlas Suntest CPS+). Stock solutions were made up in 18 acetonitrile and experimental working solutions had initial concentrations of 4.4 × 10-5 mol/L naphthalene, 1.0 × 10-6 mol/L phenanthrene, and 1.5 × 10-7 mol/L pyrene. Working solutions were made by injecting an appropriate volume of stock solution into an empty flask, followed by volatilization of the acetonitrile matrix, and the addition of Milli-Q water or the two natural water samples. In select photolysis experiments, nitrate was added to Milli-Q water to reach concentrations comparable to those present at the sampling sites (Table 1). Solutions (no headspace) were irradiated in borosilicate culture tubes (to screen most wavelengths <290 nm) in the solar simulator at 25°C using a 500 W xenon lamp until at least two half- lives of degradation were observed. Dark controls and triplicate samples were run concurrently. Radiometer and temperature readings show no change throughout. Finally, blanks revealed no detectable PAHs in the water samples despite the local coal-tar contamination with the exception of naphthalene, which occurred at less than 1% of our initial concentration used in the experiments.

The PAH concentrations were determined by high pressure liquid chromatography (Waters Corp. 2487, Breeze 3.3 software) with fluorescence spectroscopy detection at λexcitation (ex) = 252 nm/λemission (em) = 352 nm for phenanthrene,

λex = 330 nm/λem = 376 nm for pyrene, and λex = 276 nm/λem = 332 nm for naphthalene.

A direct aqueous injection of 50 µl was made from each vial and analytes were separated using a Novapak Waters C-18 reverse-phase chromatography column. Each sample was eluted isocratically at 1 ml min-1 using an 80% methanol 20% water mobile phase (v/v).

We also monitored any changes in pH over the course of the experiments. Chemical actinometry (p-nitroanisole (PNA)/pyridine system) was conducted to monitor photon flux and changes in the light source (Dulin and Mill, 1982). Results of chemical 19 actinometry for this and the following chapters are present in Appendix A.

Chemical actinometry showed no significant change in the light intensity of our solar simulator over the course of our experiments. To correlate our reaction rates to those that can occur in natural sunlight we used the following equation to estimate PNA’s equivalent rate constant (k) at noon in June at 40o N:

k = φλ∑ελLλ (1) where φ is PNA’s quantum yield, ε is the molar absorption coefficient of PNA, and L is sunlight irradiance for a specific wavelength (Leifer, 1988). We calculated a natural sunlight rate constant 2.83 d-1 over a range of 290 to 800 nm for the conditions stated above, which is approximately 3.7 times less than the PNA rate constant measured

-1 ([pyridine]o = 5 mM) in our solar simulator (10.51 d ). Thus, our PAH degradation rates should be interpreted within the context of the light source used in the present study.

Indeed, light scattering and attenuation in the water column as well as light angle would presumably retard photolysis even further in a natural water body.

2.3 Results and Discussion

The chemical composition of the two samples reflects the different land uses in the watersheds (Table 2.1). While both waters have near neutral pH, the GIN sample is characterized by a high nitrate concentration (0.205 mM). Conversely, the WNC water has negligible nitrate, but higher dissolved organic carbon (9.29 mg/L) and total iron (4

µM) concentrations. This contrast supplies insight into various indirect photolytic pathways of PAH degradation. 20 To date, direct photolysis has been shown to be the dominant pathway of degradation for PAHs in natural waters (Miller and Olejnik, 2001; Chen et al., 2000;

Lehto et al., 2000; Fasnacht and Blough, 2002). We found all three PAHs to be susceptible to direct photolytic degradation. Our PAH direct photolytic pseudo first-order rate constants (Table 2.1, denoted as Milli-Q) corroborate literature values reported by others (Fasnacht and Blough, 2002; Fasnacht and Blough, 2003a; Fasnacht and Blough,

2003b). Moreover, we observed enhanced photo-reactivity for the higher ring PAH compounds, which is consistent with their ability to absorb light at longer wavelengths in sunlight (Bertilsson and Widenfalk, 2002).

No enhanced pyrene or phenanthrene photodegradation was observed in the relatively high DOM WNC water sample (Table 2.1). Fasnacht and Blough (2002) observed a similar phenomenon and concluded that DOM mediated photodegradation of

PAHs was not an important pathway with the possible exception of anthracene. Even with this compound these investigators were unsure of the specific mechanism involved.

Indeed, relative to both the Milli-Q and GIN water experiments we observed slower reaction rates of pyrene and phenanthrene in the WNC sample and attribute a portion of this to light screening effects (Table 2.1).

To quantify light screening in natural water, we applied a correction factor that was calculated for each rate constant. Using an ultraviolet (UV)-Vis absorbance spectrum

for each sample, a light screening factor (Sλ) was calculated for each wavelength:

–(1.2)(A) Sλ = [1 – 10 ] / (2.3)(1.2)(A) (2) where A is absorbance and 1.2 is an average distribution function of light in natural

waters (Schwarzenbach, 2003; Zepp and Cline, 1977). We plotted Sλ verses wavelength 21 and integrated the area under the curve to determine the overall screening factor over the wavelengths of interest (290 – 350 nm). The WNC water has a screening factor of 0.82 while the GIN sample has a screening factor of 0.98. Reported rate constants are corrected for light screening when indicated (Figure 2.1 and Figure 2.2). The GIN water slowed the phenanthrene photolysis rate constant by 1.4% (kcorrected = rate constant

-1 corrected for light screening = 0.072 h ) and pyrene photolysis by 2.0% (kcorrected = 1.44 h-1), while the higher light screening of the WNC sample reduced the respective

-1 phenanthrene and pyrene rate constants by 18.0% (kcorrected = 0.050 h ) and 18.2%

-1 (kcorrected = 1.32 h ).

The PAH photolysis in GIN water resulted in varied effects for the target compounds. Pyrene photolysis in this sample was slower than in Milli-Q water, a portion of which is caused by light screening effects (Table 2.1, Figure 2.1). Surprisingly, phenanthrene photolysis was enhanced by GIN water (Table 2.1) even after we compensated for light screening (Figure 2.2). As stated previously the DOM level is lower for the GIN sample relative to the WNC water, but nitrate levels are significantly higher (Table 2.1). Indeed the GIN nitrate levels are in the range where it becomes the dominant OH• producing photosensitizer (Zepp et al., 1987). Because nitrate absorbs photons in the ultraviolet range present in sunlight (300 – 366 nm) we suspect that it is responsible for the enhanced degradation of phenanthrene in this water. Light absorption by the GIN sample shows that nearly all of the light-absorbing properties of the water appear to be dominated by nitrate when compared to a UV-Vis absorbance spectrum of a

0.2 mM nitrate solution (Figure 2.3). In contrast the WNC sample’s UV-Vis spectrum shows that DOM is the dominant chromophoric constituent especially at lower 22 wavelengths (~230-300 nm) (Figure 2.3).

To further corroborate our observations we conducted experiments whereby phenanthrene was photolyzed in the WNC water sample with added nitrate equivalent to and twice the concentration present in the GIN water (Fig. 2.4). We observed a linear increase in the reaction rate with increasing nitrate. In the absence of DOM the reaction rates are expected to be even higher due to DOM’s ability to scavenge OH• produced by photolyzed nitrate (Brezonic and Fulkerson-Brekken, 1998).

Dissolved organic matter has been shown to generate OH• when irradiated by sunlight. While the specific mechanism for this process is unknown, we do know that the photo-Fenton process plays an important role in the production of the hydroxyl radical in the presence of iron and under slightly acidic conditions (Southworth and Voelker, 2003;

White et al., 2003). Due to the absence of sufficient total iron in our samples coupled with circumneutral pH values (Table 2.1), we assumed that photo-Fenton pathways are not important for this system. Irrespective of the mechanism, the steady state hydroxyl

-17 radical concentration ([OH•]ss) is typically less than 10 M for natural waters in the absence of nitrate (Zepp et al., 1987; Brezonic and Fulkerson-Brekken, 1998). In the presence of nitrate at levels higher than 100 µM, however, [OH•]ss can be one or more orders of magnitude higher. In a separate study we conducted in Antarctica, we observed a roughly two order of magnitude difference in [OH•]ss between irradiated surface water from Pony Lake (a eutrophic lake on the Antarctic coast with nitrate levels that exceeded

-15 200 µM with a [OH•]ss of 4.0 × 10 M) and nitrate-free solutions prepared from the

-17 fulvic acid fraction of the Pony Lake DOM ([OH•]ss of 8.8 × 10 M) (Jennifer J.

23 Guerard, Ohio State University, Columbus, OH, USA, unpublished data). Thus, we suspect that the high nitrate levels in the GIN sample produce enough OH• to facilitate the degradation of phenanthrene (and possibly other similar ring PAHs), but not for all

- PAHs. Finally, Miller and Chin (Miller and Chin, 2002) observed NO3 mediated photolysis for the agrochemical, alachlor, whereby the photochemical degradation of alachlor becomes important in the presence of high nitrate levels (>100 µM).

Pyrene photolysis is dominated by direct photolysis with a rate constant of 2.52 h-

1 (Figure 2.1, Table 2.1). Pyrene’s reaction in the WNC and GIN samples was lower and we attribute some of this decrease to both light screening and other interferences caused by DOM in the samples e.g., binding of pyrene to the DOM phase (Lindsey and Tarr,

2000a; Lindsey and Tarr, 2000b). Chin et al. (1997) reported a pyrene partition coefficient (KDOC) of 10290 L/kgOC in the presence of Suwannee River fulvic acid. To calculate the percentage of bound pyrene in the presence of organic matter we use the equation:

% pyrene-DOC = [DOC] • Kdoc/(1 + [DOC] • Kdoc) • 100 (3) where % pyrene-DOC is the percentage of bound pyrene, [DOC] is the concentration of dissolved organic carbon (kg/L), and KDOC (L/kg) is the reported pyrene partition coefficient. The calculated fraction of free pyrene in the water is 93.5% for GIN and

91.3% for WNC. The observed reaction of pyrene in each sample is still slower than direct photolysis, even after one corrects for DOM bound pyrene (1.50 hr-1 for GIN water and 1.18 hr-1 for WNC). Additionally, when one considers the effects of DOM-bound pyrene and light screening, our measured rate constants are still lower than the direct photolysis rate constants (1.54 hr-1 for GIN water and 1.44 hr-1 for WNC). 24 The significantly higher level of nitrate in the GIN sample did not enhance the photodegradation of pyrene. In this case the extremely fast direct photolysis rate constant for pyrene exceeds its reactivity with OH• even though its reported second order rate

10 constant (kOH•) with OH• is 1.5 × 10 L/mol-s (Lindsey and Tarr, 2000b). To demonstrate this phenomenon we can express the rate of PAH photolysis as follows:

-d[PAH]/dt = kobs[PAH] = kdir[PAH] + kOH•[PAH][OH•]ss (4) where kdir is the first order rate constant for the direct photolytic pathway, t is time, and kobs is the observed rate constant. We estimated the pseudo first-order rate constant (k’) for the reaction between OH• and pyrene:

k’ = kOH• • [OH•]ss (5)

- Although we did not measure [OH•]ss for the WNC or GIN, we chose a value of 1.4 × 10

16 mol/L for GIN [OH•]ss based upon the value that we measured (using the method of

Zhou and Mopper (1990)) in Old Woman Creek, a Lake Erie coastal wetland (Huron,

OH, USA), which has a similar water chemistry profile (nitrate ~135 µM, DOM concentration of 2.8 mgC/L) as the GIN sample (E. Houtz. 2007. Hydroxyl radical production in Old Woman Creek national estuarine reserve. Undergraduate thesis, The

Ohio State University, Columbus, OH, USA). The resulting calculated rate of reaction between OH• and pyrene in GIN water (8 × 10-3h-1) is more than two orders of magnitude smaller than the direct photolysis rate constant. Therefore, the presence of OH• producing photosensitizers does not appear to impact the photolytic degradation of pyrene. To further support this conclusion, experiments run in the presence of an excess

OH• scavenger (methanol at 25 mM), did not influence the degradation kinetics of

25 pyrene (Figure 2.5).

Direct photolysis is an important degradation pathway for phenanthrene (Figure

2.2). However, our experiments show that OH• plays an important role as evidenced by the relatively higher GIN water degradation rate constant. This enhanced indirect photolysis of phenanthrene is most likely due to the nitrate-generated OH• in the GIN water. Indeed, the calculated first-order rate constant for phenanthrene and OH•

(Equation 5) is 1.16 × 10-2 h-1, based upon its reported second-order rate constant of 2.3 ×

10 -16 10 L/mol-s (Lindsey and Tarr, 2000a) and a [OH•]ss = 1.4 × 10 mol/L. The sum of the calculated OH• rate constant and the direct photolysis rate constant of 5.4 × 10-2 h-1

(Table 2.1) is 6.6 × 10-2 h-1 which is similar to the observed degradation rate constant of phenanthrene in GIN water (7.5 × 10-2 h-1, Figure 2.2). Therefore OH• producing sensitizers present in sufficient quantities affect the photolytic fate of phenanthrene, but not pyrene. As stated previously, this was demonstrated in the nitrate-spiked WNC sample where enhanced phenanthrene photolysis occurred (Figure 2.4). Indeed, the phenanthrene rate constant of WNC water spiked with 0.2 mM nitrate (the same nitrate concentration as the GIN sample) resulted in a very similar rate constant of 6.1 × 10-2 h-1

(Figure 2.4).

Direct photolysis is also an important pathway for the degradation of naphthalene with a rate constant of 1.5 × 10-2 h-1 (Table 2.1). We were unable to run GIN whole water with naphthalene experiments due to a lack of sample at this stage of the project.

However, our experiments show that OH• plays a strong role as evidenced by our experiments conducted in a 0.2 mM nitrate solution buffered with phosphate. The buffer

26 acts as an OH• scavenger at the concentrations used in this study and we observed pseudo-first order kinetics with our target PAH. Therefore, using Equation 5 and the reported second-order rate constant for naphthalene reaction with OH• (9.4 × 109 M-1 s-1;

-16 (Roder et al., 1990)) and our proxy [OH•]ss value of 1.4 x 10 mol/L we estimate the contribution from OH• in GIN water for naphthalene to be 4.7 × 10-3 h-1. The sum of this calculated rate constant due to OH• production and the direct photolysis rate constant is

2.0 × 10-2 h-1. The lower estimated naphthalene degradation rate constant relative to the

-2 -1 - observed value of 3.7 × 10 h , measured in a phosphate buffered NO3 solution, is attributable to both the higher nitrate levels present in the clean experiment relative to the

Old Woman Creek water sample (200 µM vs 135 µM) and the lack of other OH• scavengers e.g., DOM, which would lower the value of [OH•]ss (Breonic and Fulkerson-

Brekken, 1998). Nonetheless, we conclude that OH• producing sensitizers present in sufficient quantities will affect the photolytic fate of naphthalene and phenanthrene.

2.4 Conclusions

Direct photolysis of all PAHs was observed under simulated solar radiation with pyrene degrading at a faster rate than either phenanthrene or naphthalene. Phenanthrene degradation, when compared to its direct photolysis rate, increased in Gary, Indiana water, but decreased in the Wilmington, North Carolina water. The slightly lower rate of phenanthrene degradation observed for the North Carolina water sample, corrected for light attenuation effects, is statistically the same as the direct photolysis experiments.

Therefore, we attribute the lower rate of degradation in the presence of Wilmington,

27 North Carolina water to light screening by DOC, while we believe that the faster reaction rate observed for the Indiana water is the result of hydroxyl radical (OH•) chemistry generated by nitrate photolysis. Indeed corroborating this conclusion, degradation of the target compound increased when nitrate was added to the Wilmington sample. Overall rates of photolysis decrease with decreasing PAH molecular weights, confirming that larger aromatic ring compounds possess higher reactivity due to their higher extinction coefficients at wavelengths present in sunlight (Bertilsson and Widenfalk, 2002). Thus, direct photolysis is an important pathway for the degradation of PAHs in sunlit waters for larger ring PAHs, while waters rich in nitrate have increased photo-reactivity toward lower ring PAHs through reaction with OH•. Dissolved organic matter can inhibit this pathway by both light screening and the scavenging of photo-produced ROS.

28 2.5 References

Bertilsson S, Widenfalk A. (2002) Photochemical degredation of PAHs in freshwater and their impact on bacterial growth- Influence of water chemistry. Hydrobiologia 469: 23- 32.

Canonica S, Jans U, Stemmler K. (1995) Transformation kinetics of phenols in water- photosensitization by dissolved natural organic material and aromatic ketones. Environ Sci Technol 29: 1822-1831.

Canonica S, Freiburghaus M. (2001) Electron-rich phenols for probing the photochemical reactivity of freshwaters. Environ Sci Technol 35: 690-695.

Chen J, Peijnenburg WJGM, Quan X, Yang F. (2000) Quantitative structure-property relationships for direct photolysis quantum yields of selected polycyclic aromatic hydrocarbons. Sci Total Environ 246: 11-20.

Chin YP, Aiken GR, Danielsen KM. (1997) Binding of pyrene to aquatic and commercial humic substances: The role of molecular weight and aromaticity. Environ Sci Technol 31: 1630-1635.

Dulin D, Mill T. (1982) Development and evaluation of sunlight actinometers. Environ Sci Technol 16: 815-820.

Fasnacht MP, Blough NV. (2002) Aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ Sci Technol 36: 4364-4369.

Fasnacht MP, Blough NV. (2003) Kinetic analysis of the photodegradation of polycyclic aromatic hydrocarbons in aqueous solution. Aquat Sci 65: 352-358.

Fasnacht MP, Blough NV. (2003) Mechanisms of the aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ Sci Technol 37: 5767-5772.

Lehto K, Vuorimaa E, Lemmetyinen H. (2000) Photolysis of polycyclic hydrocarbons (PAHs) in dilute aqueous solutions detected by fluorescence. J Photochem Photobiol A Chem 136: 53-60.

29 Lindsey ME, Tarr MA. (2000) Inhibited hydroxyl radical degradation of aromatic hydrocarbons in the presence of dissolved fulvic acid. Water Res 34: 2385-2389.

Lindsey ME, Tarr MA. (2000b) Inhibition of hydroxyl radical reaction with aromatics by dissolved natural organic matter. Environ Sci Technol 34: 444-449.

Leifer A. (1988) The Kinetics of Environmental Aquatic Photochemistry. American Chemical Society, Washington, DC.

Miller JS, Olejnik D. (2001) Photolysis of polycyclic aromatic hydrocarbons in water. Water Res 35: 233-243.

Miller PL, Chin YP. (2002) Photoinduced degradation of carbaryl in a wetland surface water. J Agric Food Chem 50: 6758-6765.

Miller PL, Chin YP. (2005) Indirect photolysis promoted by natural and engineered wetland water constituents: Processes leading to alachlor degradation. Environ Sci Technol 39: 4454-4462.

Neff JM. (1979) Polycyclic Aromatic Hydrocarbons in the Aquatic Environment. Applied Science, London, UK.

Roder M, Wojnárovits L, Földiák G. (1990) Pulse radiolysis of aqueous solutions of aromatic hydrocarbons in the presence of oxygen. Radiation Physics and Chemistry 36: 175-176.

Schwarzenbach RP, Gschwend PM, Imboden DM. (2003) Environmental Organic Chemistry, 2nd ed. John Wiley, Hoboken, NJ, USA.

Southworth BA, Voelker BM. (2003) Hydroxyl radical production via the photo-fenton reaction in the presence of fulvic acid. Environ Sci Technol 37: 1130-1136.

Stookey LL. (1970) Ferrozine- A new spectrophotmetric reagent for iron. Anal Chem 42: 779- 781.

Vaughan PP, Blough NV. (1998) Photochemical formation of hydroxyl radical by constituents of natural waters. Environ Sci Technol 32: 2947-2953.

White EM, Vaughan PP, Zepp RG. (2003) Role of the photo-Fenton reaction in the production of hydroxyl radicals and photobleaching of colored dissolved organic matter 30 in a coastal river of the southeastern United States. Aquat Sci 65: 402-414.

Wilson RI, Mabury SA. (2000) Photodegradation of metolachlor: Isolation, identification, and quantification of monochloroacetic acid. J Agric Food Chem 48: 944- 950.

Zepp RG, Schlotzhauer PF. (1979) Photoreactivity of selected aromatic hydrocarbons in water. In Jones PR, Leber P, eds, Polynuclear Aromatic Hydrocarbons. Applied Science, London, UK, pp 141-158.

Zepp RG, Hoigné J, Bader H. (1987) Nitrate-induced photooxidation of trace organic chemicals in water. Environ Sci Technol 21: 443-450.

Zepp RG, Cline DM. (1977) Rates of direct photolysis in aquatic environment. Environ Sci Technol 11: 359-366.

Zhou XL, Mopper K. (1990) Determination of photochemically produced hydroxyl radicals in seawater and fresh-water. Mar Chem 30: 71-88.

2.6 Tables

Milli-Q water GIN water WNC water NO - 3 - 0.205 0.046 (mM) DOC (mg/L) - 6.73 9.29

Fetotal (µM) - 2.0 ±0.4 4.0 ±0.4 pH 6.0 7.4 6.8

Pyrene 2.52 ±0.13 1.41 ±0.15 1.08 ±0.13 k h-1 obs Phenanthrene -1 0.055 ±0.009 0.071 ±0.004 0.041 ±0.009 kobs h Naphthalene -1 0.015 ±0.0013 - - kobs h

Table 2.1 Selected Gary, Indiana (GIN) and Wilmington, North Carolina (WNC) water - parameters (dissolved organic carbon (DOC), nitrate (NO3 ), total iron (Fe)) and observed -1 photolytic degradation rate constants (kobs h ) in Milli-Q water and our samples. (-) Denotes not applicable.

2.7 Figures

Figure 2.1 Photoinduced degradation of pyrene in Milli-Q and Wilmington, North -1 Carolina (WNC) and Gary, Indiana (GIN) sample waters where kobs h is the observed rate constant per hour. Rate constants are corrected for light screening for both natural waters.

Figure 2.2 Photoinduced degradation of phenanthrene in Milli-Q and Wilmington, North -1 Carolina (WNC) and Gary, Indiana (GIN) sample waters where kobs h is the observed rate constant per hour. Rate constants are corrected for light screening for both natural waters.

Figure 2.3 Light absorption of selected water samples: Gary, Indiana (GIN) water (0.2 mM nitrate), 0.2 mM nitrate solution, and Wilmington, North Carolina (WNC) water (0.04 mM nitrate).

Figure 2.4 Photoinduced degradation of phenanthrene in Wilmington, North Carolina water at natural nitrate levels (0.046 mM nitrate) and spiked nitrate levels of 0.2 mM and -1 0.4 mM where kobs h is the observed rate constant per hour.  Wilmington water;  Wilmington water plus 0.2 mM nitrate;  Wilmington water plus 0.4 mM nitrate.

Figure 2.5 Photoinduced pseudo first order degradation (Ln(concentration/initial concentration)) of pyrene in Gary, Indiana water (0.2 mM nitrate) and GIN water with 25 mM methanol as a hydroxyl radical scavenger.  Gary water; ♦ Gary water plus 25 mM methanol.

CHAPTER 3

IBUPROFEN PHOTOLYSIS IN THE PRESENCE OF THREE FULVIC ACIDS

3.1 Introduction

Pharmaceuticals and personal care products (PPCPs) have been detected in natural waters (e.g., wetlands, lakes, and rivers) throughout the world and are widely recognized as emerging environmental contaminants (Ternes, 1998; Kolpin, 2002; Daughton and

Ternes, 1999). PPCPs find their way into the environment through application, subsequent wastewater introduction, and incomplete removal. Concern regarding their presence in the environment stems not from concentration; most of these compounds are present at low levels, but from chronic exposure to low doses and unknown toxicological impacts of unidentified derivatives. In 1999/2000 a United States Geological Survey study detected the pharmaceutical ibuprofen in ~10% of 139 streams sampled across the

United States at concentrations as high as 0.005 µM (Kolpin, 2002).

Ibuprofen is a non-steroidal anti-inflammatory drug (NSAID) contained in widely used medications. It is an alkyl benzene with a carboxylic acid functional group, a pKa of

~4.8, and a log Ko/w of 2.48 (Scheytt et al., 2005). Ibuprofen is chiral with pharmacological activity attributed to the S-isomer and is typically administered as a 38 racemic mixture (Hutt, 1983). Over 70% of ibuprofen is metabolized and excreted in urine. Hydroxylated and carboxylated compounds are the predominant metabolized forms

(Hutt, 1983). Despite this significant metabolism, 0.015 µM ibuprofen has been detected in wastewater influent (Buser, 1999).

Ibuprofen is chiral with pharmacological activity attributed to the S-isomer and is typically administered as a racemic mixture (Hutt, 1983). Once administered, the inactive

R-enantiomer undergoes racimization to the active S-(+)-ibuprofen in the body (Hutt,

1983; Buser, 1999) with a significant amount (70-80%) of the total ibuprofen being excreted as the S-enantiomer (Hutt, 1983; Lockwood, 1983). As a result, S-ibuprofen present in wastewater influent has been detected at concentrations up to eight times higher than the inactive R-enantiomer (Buser, 1999). Wastewater treatment removal is efficient (> 95%), yet incomplete, with ibuprofen found in wastewater treatment effluent at concentrations of 0.0004 µM and relatively lower S-enantiomer excess (up to 2 times more S than R) (Buser et al., 1999). Ibuprofen was detected in Swiss surface waters with a diminished excess of the S-enantiomer with a ratio similar to the racemic mixture, i.e., no excess of the S-enantiomer; this is explained by relatively rapid degradation of the S form (compared to the R) when incubated in lake water (Buser et al., 1999). It is also important to note that ibuprofen levels reported by Buser et al. in wastewater effluent are an order of magnitude below ibuprofen concentrations detected in the environment by

Kolpin et al., Ternes, and others. Higher ibuprofen concentration in natural waters leads to speculation that the dominant source of ibuprofen in the environment is combined sewage overflow and possibly septic tank leakage (Buser et al., 1999).

39 NSAIDs, as well as many other classes of pharmaceuticals, are known to degrade photochemically (Boreen et al., 2003; Packer et al. 2003). Packer et al. showed that ibuprofen is susceptible to direct and indirect photolytic pathways. Facilitating photolysis through indirect processes, humic substances are an important portion of the dissolved organic matter pool in surface waters that can generate responsible reactants such as reactive oxygen species (ROS) or photochemically generated excited states of dissolved organic matter (DOM) (Gerecke, 2001; Miller and Chin, 2002; Packer et al., 2003), which can subsequently react with ibuprofen. The XAD-8 isolated fulvic acid (FA) fraction of humic substances represents the chromophoric portion of DOM as the more aromatic, more hydrophobic portion of organic matter (Thurman and Malcolm, 1981) and thus the most photochemically active fraction (Brown et al., 2004; Williams et al., 2007).

Packer et al. showed ibuprofen indirect photolysis facilitates its degradation beyond that of direct photolysis through, in part, the reactive oxygen species hydroxyl radical (OH•).

These investigators suspected an additional reactive species of unknown origin that contributes to the photodegradation of ibuprofen in natural waters but this pathway has yet to be elucidated (Packer et al., 2003).

This study examines the mechanism(s) responsible for the photochemical degradation of ibuprofen in the aquatic environment and the effect of fulvic acids on this process. We studied the direct and indirect photolysis of ibuprofen by probing two initial concentrations using fulvic acids that represent the continuum of DOM derived from terrestrial and microbial processes.

40 3.2 Methods

3.2.1 Chemicals and Fulvic Acids

Milli-Q (18MΩ-cm, Millipore) water was used for all aqueous solutions. All solvents (acetonitrile, hexane, methanol, isopropanol and ether) and chemicals

(hydrochloric acid, etc.) were reagent-grade or higher. Racemic ibuprofen and S-(+)- ibuprofen with a purity of 99% were obtained from Acros Organics. Natural organic matter standards, Suwannee River Fulvic Acid (SRFA) and Pony Lake Fulvic Acid

(PLFA) as discussed in Chapter 1, were obtained from the International Humic Substance

Society. Additionally, we isolated a fulvic acid from a National Estuarine Research

Reserve wetland in Huron, OH according to standard XAD-8 non-ionic resin protocol

(Leenheer, 1981; Thurman & Malcolm, 1981). A 400 L water sample from Old Woman

Creek, OH in August 2006 was pre-filtered to 20 microns, filtered in a series of glass fiber filters with 5 micron and 0.5 micron nominal pore size, acidified onsite to pH 2 with

HCl. The water was then transferred to Columbus, OH where the XAD-8 fractionation technique was used to obtain the hydrophobic portion of the organic matter; which we term the fulvic acid fraction of Old Woman Creek (OWCFA).

3.2.2 Photolytic Reactions

Ibuprofen stock solutions were prepared in acetonitrile. Experimental working solutions of 10 µM or 0.1 µM were prepared by addition of an appropriate volume of stock, evaporation of the acetonitrile, re-constitution into the desired aqueous matrix, and pH adjustment to 7 (±0.1) with HCl and/or NaOH. Select solutions were prepared 41 anoxically by either: (1) sparging with argon for 1 min/mL and transferring to photolysis tubes within a glove box (95% N2 / 5% H2) or (2) subjecting solutions to 3 freeze-pump- thaw cycles in a Schlenk line followed by canula transfer into photolysis tubes. After this treatment, oxygen levels in photolysis solutions were below the limit of detection with a dissolved oxygen probe. To probe indirect photochemical pathways, each solution was spiked with the same weight of the desired fulvic acid and an aliquot was used for total organic carbon analysis (Shimadzu TOC-5000). Isopropanol (20 mM) was used as a hydroxyl radical scavenger (Packer et al., 2003). Air tight quartz tubes (1 cm pathlength capped with Teflon-lined quartz lids) were used for photolysis, screening wavelengths

<290 nm. Photolysis of ibuprofen was conducted using a solar simulator (Atlas Suntest

CPS+) with a Xenon arc lamp at 25°C and lamp energy of 500 Watts for a time duration equivalent to a minimum of one half-life. Select photolysis experiments were evaluated over three half-lives in order to determine the appropriate rate model for degradation kinetics. Chemical actinometry (p-nitroanisole/pyridine) (Dulin and Mill, 1982) was conducted periodically to monitor photon flux and changes in the light source intensity.

Actinometry showed no significant change in the light intensity of our solar simulator over the course of experiments and results are outlined in Appendix A. Dark controls were run concurrently. Temperature, pH, and radiometer readings were monitored during each experiment, with all three parameters remaining consistent.

To correlate our reaction rates to those that occur in natural sunlight we used a normalization procedure defined in Chapter 2 and determined that the solar simulator is

3.7 times stronger than natural sunlight. Thus, ibuprofen degradation rates should be interpreted within the context of the light source used in this study. Additionally, all 42 degradation experiments obeyed pseudo-first-order kinetics. An experimental procedure developed by Mopper and Zhou (1990) was used to quantify hydroxyl radical production in sample solutions. This method uses methanol as a probe for hydroxyl radical, the two reacting to produce formaldehyde which is subsequently derivatized with 2,4- dinitrophenylhydrazine (DNPH) for quantification by high pressure liquid chromatography (HPLC) (Mopper and Zhou, 1990).

3.2.3 Natural Sunlight Experiment

Ibuprofen working solutions were prepared as with simulated solar experiments in a matrix of Old Woman Creek Natural Estuarine Research Reserve (Figure 1.1) wetland water (41º22’54”N; 82º30’52”W) obtained from the wetland at the time of the experiment, June 2006, and filtered to 0.45 µm with glass fiber filters. Solutions were irradiated for one day (~14 hours) in borosilicate culture tubes with teflon lined lids on the banks of Old Woman Creek Natural Estuarine Research Reserve, OH on a black platform elevated 10 cm above the ground. Chemical actinometry using a p-nitroanisole

(PNA)/pyridine (PYR) system (Dulin and Mill, 1982) was conducted over the course of the experiment and showed minimal change in the natural light intensity (see Appendix

A).

3.2.4 HPLC

Ibuprofen and photolysis byproducts were analyzed by High Pressure Liquid

Chromatography (HPLC) with UV-Vis dual wavelength detector at λ = 223 nm and λ =

264 nm respectively (Waters Corp. 2487, Breeze 3.3 software). A direct injection of 50 43 µL from each vial was made into the HPLC and analytes were separated using a Restek

C-18 reverse-phase chromatography column. Each sample was eluted isocratically at 1 ml min-1 using a 60% acetonitrile and 40% water mobile phase (v/v) buffered with phosphate to pH 3.

3.3. Results and Discussion

3.3.1 10 µM Photolysis Experiments

Racemic ibuprofen (10 µM initial concentration) degrades by direct photolysis, with a pseudo first order rate constant of 0.0025 h-1 (Table 3.1). This translates to a half-life on the order of 200 hours, relatively slow when compared to direct photolysis of other organic contaminants (Chapter 2), yet is a similar rate when compared to other photolytic studies of ibuprofen (taking into account varying light source intensity) (Packer et al.,

2003; Lin and Reinhard 2005). S-(+)-ibuprofen (10 µM initial concentration) degrades by direct photolysis at the same rate as the racemic mixture (Table 3.1). Buser et al. (1999) documents a higher proportion of the S-(+) (vs R) isomer released in wastewater treatment effluent but a racemic mixture is found in the environment. Since both the racemic and S-(+)-Ibuprofen photodegrade at the same rate (Table 3.1), we attribute this phenomenon to other sources of degradation (i.e. microbial).

Ibuprofen degrades approximately 5 times faster than by direct photolysis (~45 hour half life) (Table 3.1) in the presence of each fulvic acid. I attribute this acceleration to

DOM generated reactive transients. However, as each fulvic acid is derived from a different source, this phenomenon is surprising. Organic matter can produce a cocktail of 44 reactive species, which varies with source (White, 2003; Hoigne et al., 1989; Cooper et al., 1989). Suwannee River, GA fulvic acid (SRFA) is sourced in decaying terrestrial plants while Pony Lake, Antarctica (PLFA) is a eutrophic lake whose dissolved organic matter is from microbial matter i.e. non-lignin precursors. Old Woman Creek fulvic acid’s (OWCFA) composition is derived from a combination of microbial and terrestrial inputs. Therefore, converse to demonstrated results; degradation rates are expected to vary with fulvic acid.

To further probe the role of fulvic acids, the known OH• scavenger, isopropaol, was added to photolysis experiments. Hydroxyl radical is shown to be an important reactive oxygen species generated by DOM, particularly with OWCFA as the degradation rate is

-1 slowed by 46% (kobs = 0.009 ± 0.001 h ), confirming Packer et al. (2003) (Figure 3.1).

Hydroxyl radical is also an important ROS generated by SRFA (k = 0.007 ± 0.001 h-1;

64% reaction reduction) (Figure 3.2), more important than in the presence of OWCFA.

Surprisingly, in the presence of PLFA, scavenging by isopropanol indicates that hydroxyl radical is not a relevant photosensitizer, as its addition does not change the rate of ibuprofen degradation (Figure 3.3). Our [OH•]ss measurements corroborate these observations because photolyzed PLFA solutions yielded [OH•]ss levels that are a factor of three lower than levels measured for OWCFA (8.80 × 10-17 M vs 2.42 × 10-16 M respectively).

The role of oxygen dependent reactive transients produced by organic matter was probed by photolyzing suboxic ibuprofen solutions (10 µM) in the presence of each fulvic acid. The rate of degradation in the presence of PLFA and SRFA remains

45 statistically the same as oxic solutions (Table 3.1). In the presence of OWCFA, suboxic conditions accelerate the degradation of ibuprofen (Table 3.1). In these reactions OH• is still an effective reactive transient as Vaughn and Blough (1998) demonstrate that OH• formation is not dioxygen dependent. Therefore, the reactive transient/s responsible for degradation in the presence of PLFA and the non-OH• catalyzed degradation in the

1 presence of SRFA, is dioxygen independent. I speculate that singlet oxygen ( O2), and

- superoxide (O2• ) are not relevant reactive oxygen species. In the presence of OWCFA, suboxic conditions accelerate the degradation of ibuprofen (Table 3.1). I speculate in the suboxic OWCFA system 3DOM is no longer as extensively scavenged by dioxygen and thus is allowed to play a larger role in degradation.

Next, I explored ibuprofen degradation in Old Woman Creek natural water under natural sunlight conditions. At 10 µM, ibuprofen undergoes photo-degradation in natural

-1 sunlit systems with a kobs of 0.0038 ± 0.001 h . After correction (equation 1) to account for different source intensities, we determined that ibuprofen degrades at a similar rate

-1 (kcor = 0.014 ± 0.001 h ) in Old Woman Creek water in natural sunlight as it degrades in solar simulator experiments with OWCFA (Table 3.1). When making this comparison it is important to note the lower DOC level which would slow relative degradation rates and the presence of 135 µM nitrate, a known hydroxyl radical generator (Zepp, 1987), which would accelerate relative degradation rates in the natural water. Therefore, the similar rate between natural and simulated systems is attributed to a lower DOM concentration and nitrate generated OH•. In addition to nitrate and fluctuating DOC levels, light

46 scattering and light attenuation in the water column as well as shifts in the incident light angle would presumably affect the rate of ibuprofen photo-degradation in natural waters.

3.3.2 0.1 µM Photolysis Experiments

The effect of initial ibuprofen concentration on indirect photolysis was also investigated. For all three fulvic acids the degradation rate constant increased compared to 10 µM concentrations as shown in Figure 3.4. However, only a slight increase was observed for OWCFA whereas rate constants increased in the presence of PLFA and

SRFA (× 5 and × 2, respectively) (Table 3.1, Figure 3.4). Canonica and Freiburghaus

2001 demonstrated that the photo-degradation of electron-rich phenols at low (0.1 µM) concentrations is accelerated 2-3 fold relative to higher (5 µM) concentrations in the presence of natural freshwaters and DOM isolates. They concluded that these compounds reacted with DOM generated photo-oxidants of various lifetimes. At high concentration

(5 µM) the pseudo-first order degradation is dominantly a function of the short-lived species such as excited triplet states of DOM (3DOM) while at low concentrations reaction kinetics reflect the combined effect of short-lived (≤2 µs) species, such as excited triplet states of DOM (3DOM) and long lived species (≥100 µs) which may include peroxyl or phenoxyl radicals, radical cations of aromatic structures, and excited states of DOM chromophoric constituents (Canonica and Hoigne 1995; Canonica et al.

1995; Canonica and Freiburghaus 2001).

Our results for SRFA and PLFA are consistent with those of Canonica and

Freiburghaus (2001), Canonica and Hoigne (1995), and Canonica et al. 1995. This 47 increase in degradation rate at low ibuprofen concentrations (Figure 3.4), indicates lower concentrations reflect the influence of long-lived and short-lived scavengers generated by

PLFA and SRFA on ibuprofen photo-degradation. A slight increase in degradation in the presence of OWCFA demonstrates the same trend but the long-lived reative transients from this fulvic acid play a more minor role. Conversely, shorter-lived species, i.e.

3DOM, contribute to the photo-transformation at 10 µM evidenced by 1) relatively slower rates than the 0.1 µM photolysis and 2) acceleration of the degradation rate in the presence of OWCFA in suboxic systems.

To further probe reactive transients at low concentrations, suboxic ibuprofen photolyses at 0.1 µM were performed in the presence of each fulvic acid. Ibuprofen degradation rates in the presence of PLFA and SRFA solutions and low dioxygen concentrations decreased compared to dioxygen saturated conditions (Figure 3.5), indicating the presence of a fulvic acid generated dioxygen dependent short-lived reactive transient at low concentrations such as peroxyl and phenoxyl radicals. Conversely, in the presence of OWCFA suboxic conditions demonstrate no change in the rate of degradation

(Figure 3.5), which indicates the presence of a dioxygen independent reactive transient.

Additionally, whatever long-lived transients are responsible for the increased degradation at low concentrations, they seem to be produced in varying levels depending on fulvic acid as all oxic/suboxic rates vary (Figure 3.4 and Figure 3.5). This is consistent with the idea that the species will exist upon irradiation in various steady states depending on fuvic acid composition/source.

48 Isopropanol was used to further probe the reactive transients responsible for degradation at low concentrations. Isopropanol addition slows low concentration reaction rates significantly, generating rates similar to 10 µM ibuprofen photolyses with FA

(Table 3.1; Figure 3.5). Specifically, at low concentrations in the presence of OWCFA, isopropanol slows the reaction by 49%. The presence of isopropanol slows reaction rates with SRFA and PLFA by 33% and 86%, respectively. As PLFA has an [OH•]ss significantly lower than that of OWCFA yet the reaction rate is slowed 86% (vs 49% and

33%), I speculate that while isopropanol is a scavenger of OH•, at low concentrations it also acts as a scavenger to some of the long-lived reactive transient/s responsible for the increased rate of degradation.

3.5 Conclusions

Ibuprofen degrades by direct photolysis. Its degradation rate increases approximately 5-fold with the presence of each fulvic acid. Isopropanol indicates the importance of OH• in the presence of OWCFA and SRFA by slowing of the degradation rate. However, in the presence of PLFA and isopropanol the rate remains the same as without the scavenger, indicating that in this case OH• is irrelevant and another reactive transient is responsible for degradation. Suboxic experiments ([Ibuprofen]o = 10 µM) in the presence of SRFA and PLFA show no change in degradation rate indicating that short lived reactive transients involved in degradation at this concentration are dioxygen independent. However, suboxic experiments with OWCFA show a rate acceleration

49 suggesting the importance of 3DOM/molecular oxygen relationship upon degradation kinetics.

Photolyses at lower, more environmentally relevant concentrations (0.1 µM) of ibuprofen reveal an accelerated rate in the presence of each FA, with a slight increase in the presence of OWCFA, a doubling of the rate in the presence of SRFA, and an acceleration × 5 in the presence of PLFA when compared to experiments at 10 µM.

Coupled with suboxic experiments indicating a SRFA and PLFA produce dioxygen dependent reactive transient/s while OWCFA does not, I speculate various reactive transients playing a role for each organic matter. Isopropanol slowing the degradation rate to values converse to OH• steady state numbers indicates it scavengers not only

OH•, but other DOM generated reactive transients present at low concentrations as well.

50 3.5 References

Boreen, A. L.; Arnold W. A.; McNeill K. (2003) Photodegradation of pharmaceuticals in the aquatic environment: A review. Aquat. Sci. 65, 320-341.

Brown, A.; McKnight, D.M.; Chin, Y.P.; Roberts, E.C.; Uhle, M. (2004) Chemical characterization of dissolved organic matter in Pony Lake, a saline coastal pond in Antarctica. Mar Chem. 89, 327-337.

Buser, H.; Poiger, T.; Muller M. D. (1999) Occurrence and environmental behavior of the chial pharmaceutical drug ibuprofen in surface waters and in wastewater. Environ. Sci. Technol. 33, 2529-2535.

Canonica, S.; Freiburghaus, M. (2001) Electron-rich Phenols for probing the photochemical reactivity of freshwaters. Environ. Sci. Technol. 35, 690-695.

Canonica, S; Jans, U.; Stemmler K.; Hoigné J. (1995) Transformation kinetics of phenols in water: Photosensitization by dissolved matural organic material and aromatic ketones. Environ. Sci. Technol. 29, 1822-1831.

Canonica, S.; Hoigné J. (1995) Enhanced oxidation of methoxy phenols at micromolar concentration photosensitized by dissolved organic matter. Chemosphere. 30, 2365-2374.

Cooper, W.J.; Zika, R.G.; Petasne, R.G.; Fischer, A.M. (1989) Sunlight Induced Photochemistry of Humic Substances in Natural Waters: Major Reactive Species. ACS Symposium Series. 219, 333-362.

Daughton C.G.; Ternes, T.A. (1999) Pharmaceutical and personal care products in the environment: Agents of subtle change? Environ. Health Persp.107, 907-938.

Dulin, D.; Mill T. (1982) Development and Evaluation of Sunlight Actinometers. Environ. Sci. Technol. 16, 815-820.

Gerecke, A.C.; Canonica, S.; Muller, S.R.; Scharer, M.; Scharzenbach, R.P. (2001) Quantification of dissolved organic matter (DOM) mediated phototransformation of phenylurea herbicides in lakes. Environ. Sci. Technol. 35, 3915-3923.

Hoigne, J.; Faust, B.C.; Haag, W.R.; Scully, F.E.; Zepp, R.G. (1989) Aquatic Humic Substances and Sinks of Photochemically Produced Transients Reactants. ACS Symposium Series. 219, 363-381.

Hutt, A.J.; Caldwell J. (1983) The metabolic chiral inversion of 2-arylpropionic acids- a novel route with pharmacological consequences. J. Pharm. Pharmacol. 35, 693-704.

51 Kolpin, D.W.; Furlong, E.T.; Meyer, M.T.; Thurman, E.M.; Zaugg, S.D.; Barber, L.B.; Buxton, H.T. (2002) Pharmaceuticals, hormones, and other organic, wastewater contaminants in U.S. streams, 1999-2000: A national reconnaissance Environ. Sci. Technol. 36, 1202-1211.

Leenheer J.A. (1981) Comprehensive approach to preparative isolation and fractionation of dissolved organic carbon from natural waters and wastewaters. Environ. Sci. Technol. 15, 578-587.

Lin, A.Y.; Reinhard, M (2005) Photodegradation of common environmental pharmaceuticals and estrogens in river water. Environ. Tox. Chem. 24, 6, 1303-1309.

Lockwood, G.F.; Albert, K.S; Gillespie, W.R.; Bole, G.G.; Harkcom, T.M.; Szpunar, G.J.; Wagner, J.G. (1983) Pharmacokinetics of ibuprofen in man. I. Free and total area/dose relationships. Clin. Pharmacol. Ther. 34, 97-103.

Mopper, K.; Zhou X. (1990) OH• photoproduction in the sea and its potential impact on marine processes. Science. 250, 661-664.

Miller, P.L.; Chin, Y.P. (2002) Photoinduced degradation of carbaryl in a wetland surface water. J. Agric. and Food Chem. 50, 6758-6765.

Packer, J.L.; Werner, J.J.; Latch, D.E.; McNeill, K.; Arnold, W.A. (2003) Photochemical fate of pharmaceuticals in the environment: Naproxen, diclofenac, clofibric acid, and ibuprofen. Aquat. Sci. 65, 342-351.

Scheytt, T.; Mersmann, P.; Lindstädt, R.; Heberer, T. (2005) 1-Octanol/water partition coefficients of 5 pharmaceuticals from human medical care: Carbamazepine, clofibric acid, diclofenac, ibuprofen, and propyphenazone. Water Air Soil Poll. 165, 3-11.

Ternes, T.A. 1998. Occurrence of drugs in German sewage treatment plants and rivers. Wat. Res. 32, 3245-3260.

Thurman, E.M.; Malcolm, R.L. (1981) Preparative isolation of aquatic humic substances. Environ. Sci. Technol. 15, 463-466.

Vaughn, P.P.; Blough, N.V. (1998) Photochemical formation of hydroxyl radical by constituents of natural waters. Environ. Sci. Technol. 32, 2947-2953.

Williams, M.W.; Knauf, M.; Cory, R.; Caine, N.; Liu, F. (2007) Nitrate content and potential microbial signature of rock glacier outflow, Colorado Front Range. Earth Surf. Proc. Land. 32, 1032-1047.

Zepp, R.G.; Hoigné, J.; Bader, H. (1987) Nitrate-Induced Photooxidation of trace organic chemicals in water. Environ. Sci. Technol. 21: 443-450. 52 3.6 Tables

[Racemic Ibuprofen]0 Fulvic Acid + 20 mM -1 -1 = 10 µM [DOC] mgC/L kobs (h ) Isopropanol kobs(h ) Direct - 0.0025 ±0.0004 - Direct (S-(+)- Ibuprofen) - 0.0031 ±0.0006 - OWCFA 5.56 ±0.18 0.015 ±0.001 0.009 ±0.001 SRFA 7.2 ±0.2 0.016 ±0.001 0.007 ±0.001 PLFA 5.45 ±0.19 0.014 ±0.001 0.016 ±0.001

SRFA low O2 7.2 ±0.2 0.017 ±0.002 -

PLFA low O2 5.45 ±0.19 0.015 ±0.002 -

OWCFA low O2 5.56 ±0.18 0.019 ±0.001 Natural Sunlight; OWC water 2.8 ±0.2 0.014 ±0.001 -

[Racemic Ibuprofen]0 Fulvic Acid + 20 mM -1 -1 = 0.1 µM [DOC] mgC/L kobs (h ) Isopropanol kobs(h ) PLFA 5.45 ±0.19 0.079 ±0.013 0.013 ±0.006 SRFA 7.2 ±0.2 0.029 ±0.004 0.020 ±0.004 OWCFA 5.56 ±0.18 0.019 ±0.001 0.011 ±0.002

PLFA low O2 5.45 ±0.19 0.056 ±0.009 -

SRFA low O2 7.2 ±0.2 0.015 ±0.002 -

OWCFA low O2 5.56 ±0.18 0.022 ±0.006 -

Table 3.1 10 µM and 0.1 µM racemic and S-(+) (when indicated) ibuprofen degradation -1 rate constants (kobs h ) in the presence of simulated and natural sunlight, OWCFA (Old Woman Creek Fulvic Acid), SRFA (Suwannee River Fulvic Acid), and PLFA (Pony Lake Fulvic Acid) at indicated dissolved organic carbon levels (DOC), in the absence of molecular oxygen, and in the presence of isopropanol. (–) Denotes not applicable.

53 3.6 Figures

Figure 3.1 Pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) direct photolysis degradation and in the presence of Old Woman Creek Fulvic Acid (OWCFA) with and without isopropanol.

Figure 3.2 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) direct photolysis degradation and in the presence of Suwannee River Fulvic Acid (SRFA) with and without isopropanol.

Figure 3.3 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of Ibuprofen (10 µM) in Milli-Q (Direct) photolysis and in the presence of Pony Lake Fulvic Acid (PLFA) with and without 20 mM isopropanol.

Figure 3.4 10 µM and 0.1 µM ibuprofen pseudo first order degradation rate constants -1 (kobs h ) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA).

-1 Figure 3.5 0.1 µM ibuprofen pseudo first order degradation rate constants (kobs h ) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) under oxic and suboxic systems and with 20 mM isopropanol (Iso).

CHAPTER 4

IBUPROFEN PHOTOLYSIS: A BYPRODUCT ANALYSIS AND PROPOSED CHEMICAL MECHANISM

4.1 Introduction

Several authors have documented derivatives resulting from photochemical pathways in aqueous media and proposed chemical mechanisms that give rise to these derivatives (Edhlund et al., 2006; Boreen et al., 2005; Pérez-Estrada et al., 2005, and others). Indeed, the light-induced transformation, or photosensitized decarboxylation

(photodecarboxylation) of NSAIDs and resulting derivative formation has been addressed by several studies in non-aqueous systems (Bosca, 2001; Bosca, 1994; Monti, 1997;

Budac, 1992; Castell et al., 1987) and natural systems (Pérez-Estrada et al., 2005).

Photodecarboxylation involves light-induced decarboxylation whereby a benzylic radical is formed followed by the subsequent ejection of CO2 and a solvated electron.

This electron can then be scavenged by molecular oxygen (O2), generating superoxide

• - ( O2 ) leading to a series of oxygenated byproducts (Bosca, 1994). Castell et al. (1987) proposed the same mechanism for the decomposition of ibuprofen when photolyzed by a medium pressure Hg lamp in methanol. This irradiation gave rise to several byproducts

59 via light-induced excitation of ibuprofen, cleavage of the C-C bond α to the carboxyl group generating a benzylic radical intermediate, and finally hydrogen abstraction, dimerzation, methanol addition, or reaction with dioxygen generating various byproducts

(Castell, 1987).

As demonstrated in Chapter 3, ibuprofen undergoes direct and indirect photolysis in natural systems leading to speculation regarding photo-derivatives. To my knowledge no study has probed the specific mechanism of sunlight induced photochemical transformation of ibuprofen in natural systems. Furthermore, photo-derivatives generated upon the photolysis of ibuprofen are uninvestigated as well. In this study, I examine the chemical mechanism of ibuprofen degradation through identification of photo-derivatives and an analysis of derivative production in the presence of various fulvic acids. Then, using this information and literature review, I propose a mechanism of sunlight-mediated transformation of ibuprofen in natural waters.

4.2 Methods

4.2.1 Chemicals

All solvents (hexane, methanol, acetonitrile) and chemicals used (formic acid, isobutylacetophenone) were reagent grade or better. CDCl3 (99.8% -D) were obtained from Acros Organics. Additional materials used are as described in Chapter 3.

60 4.2.2 Photolytic Reactions

Ibuprofen solutions were prepared as described in Chapter 3. Direct and indirect photolysis of ibuprofen, correlation of reaction rates with natural sunlight, and actinometry was conducted as described in Chapter 2. HPLC is as described in Chapter 3 with the following change: 264 nm UV absorbance detection wavelength for photoderivatives.

4.2.3 LC-MS

A Bruker Liquid Chromatography Electrospray Ionization Quadropole Time of

Flight Mass Spectrometer with corresponding acquisition software (Hystar version 3.1,

ULTROTOF) was used to ascertain photolysis byproduct structures. The analysis was performed in positive mode with a gradient mobile phase beginning at 50/50 0.1% formic acid/methanol and ramping to 20/65/15 0.1% formic acid/methanol/acetonitrile over 8 minutes. Each sample run totaled 11.5 minutes using a 0.4 mL min-1 flow rate with a

Waters C18 column (2.1 × 100 mm).

4.2.4 GC-MS

After irradiation, the byproduct was extracted from solution using methanol and evaporated to concentrate. These solutions were subjected to electron impact, Gas

Chromatography with a MS/MS detector (Varian Star 3400CX with Saturn 2000

MS/MS). GC-MS/MS parameters include an initial temperature of 50ºC held for 6 minutes, ramped to 200ºC at 10ºC per minute and held for 1 minute, and ramped to 270ºC

61 and held for 2 minutes. A 4 µL inject of sample in splitless mode 10 to 1 with a Supelco

Equity 1701 30 meter column and helium carrier gas resulted in a 9 minute sample run.

4.2.4 1H and COSY NMR

Standard solutions of ibuprofen and isobutylacetophenone were prepared at approximately 0.5 M solutions in CDCl3 (99.8%-D), which was also used as an internal reference. Solutions for analysis of photochemical byproduct formation were prepared by photolyzing a 100 µM ibuprofen solution in DI water for 48 h, and exaction of photolyzed solution with hexane. Hexane extracts were desiccated with anhydrous sodium sulfate, decanted and evaporated to yield a colorless residue, which then was dissolved in CDCl3 prior to NMR analysis. Nuclear Magnetic Resonance (NMR) spectra were acquired at the NMR laboratory, within The Ohio State University Chemistry

Department. All spectra were acquired on a Bruker Avance DRX 500 MHz NMR spectrometer, except for isobutylacetophenone, which was measured using a DPX 400

MHz NMR spectrometer. One-dimensional NMR spectra were acquired using a 14.8

µsec (90°) pulse, with a 1 second relaxation delay, for 128 scans at 294 K. Spectra were processed using 0.3 Hz line broadening. Gradient enhanced 2-D homonuclear correlation spectra (COSY) were used to reveal adjacent protons in a single spin system. Pulse sequence for COSY acquisition includes a 1.5 second delay followed by a 14.8 µs (90°) pulse, a delay incremented by the inverse of the sweep window (typically ~1 µsec), a second 90° pulse, and a 0.23 second acquisition. The incremented delay was used to allow for the evolution of the spin-spin coupling, and thereby detect vicinal protons. A

62 total of 8 scans were acquired with 256 increments for each scan, and a sweep width was set at 9 ppm or approximately 4500 Hz.

4.3 Results and Discussion

Upon irradiation, multiple byproducts are formed and accumulate during direct photolysis, in the presence of natural sunlight, and in the presence of SRFA, OWCFA, and PLFA under suboxic conditions and in the presence of the OH• scavenger isopropanol (all photochemical experiments conducted in Chapter 3). Surprisingly, some of these byproducts are more non-polar than ibuprofen based on HPLC retention time, with one more non-polar byproduct dominating both the HPLC and UV-Vis assay (264 nm absorbance) (Figure 4.1) that accumulates as the experiment proceeds. Further probing these photoderivatives, 1H NMR (Figures 4.2 and 4.3) and LC-MS (Figure 4.4a) spectra of unreacted ibuprofen and its photolyzed solution reveal multiple compounds, consistent with our HPLC data.

LC-ESI-QTOF-MS (positive mode) identified a molecular weight of 176.2 and suggests two possible derivatives in the photolysis solution: isobutylacetophenone

(IBAP), an intermediate in the synthesis of ibuprofen, and 1-(4-isobutylphenyl)ethanol

(4IBPE) (Figure 4.4b/c). GC-MS corroborates the presence of multiple photolysis byproducts of approximately the same molecular weight in the photolysis solution (Table

4.1). Moreover, the identification of IBAP agrees with Castell et al. who identified this compound as one of the byproducts of ibuprofen when irradiated in methanol, where

~8% IBAP is formed ([Ibuprofen]o = 16 mM). Once purchased, a UV-Vis spectrum of

63 IBAP confirms the 264 nm absorbance (Figure 4.5). Additionally, a NMR spectrum of the IBAP standard is similar to the photolyzed solution spectra (Figure 4.2 and 4.6) further confirming this structural assignment; note the quartet at 2.5 ppm and doublet at

7.8 ppm in both spectra. Indeed, isobutylacetopheone has been detected at concentrations of 0.00023 µM in waters near residential properties, citing ibuprofen degradation as the environmental precursor (Zorita et al. 2007).

We isolated the major byproduct peak (Figure 4.1b; peak retention time 7 minutes) from the HPLC waste line effluent and analyzed it with both 1-D and 2-D

(COSY) 1H NMR. The HPLC isolated 1D spectra while still complex, reveals a lack of methyl group quartet at 2.5 ppm and doublet at 7.8 ppm (Figure 4.7). This is inconsistent with the IBAP NMR spectrum (Figure 4.6) indicating it is not the major byproduct

(Figure 4.6 and 4.7). Additionally, the 1-D NMR spectra’s complexity leads us to speculate that multiple compounds are co-eluting from the HPLC column at 7 minutes

(Figure 4.1b). 2-D 1H NMR (COSY) of the HPLC isolated fraction reveals structural connectivity between the quartet at 4.1 ppm and a doublet at 1.2 ppm (Figure 4.7 and

4.8). This quartet at 4.1 ppm, when compared to the ibuprofen scan (Figure 4.2), has shifted upfield from 3.7 ppm upon photolysis, but is still well resolved. This shift, combined with COSY connectivity, indicates the presence of a more electronegative functional group at the carboxylic acid group of ibuprofen, which has been altered due to photolysis.

Taken together, NMR reveals the structure of the major byproduct as 1-(4- isobutylphenyl)ethanol (mw = 176.2; 4IBPE) confirming mass spectrometry results,

64 where the carboxylic acid has undergone hydroxylation (Figure 4.7). Further confirming this structural assignment of 4EDMS is a COSY of isobutylacetophenone a lack of the same connectivity in the photolyzed solution (Figure 4.9). We also observed a shift in the aromatic doublets from the left of the solvent peak (CHCl3 ~7.4 ppm) prior to photolysis to the right after photolysis (Figure 4.7) indicating the attachment of a more electronegative moiety to the benzene ring or specifically, hydroxylation of the aromatic spin system. However, NMR spectra complexity compiled with GC-MS and LC-ESI-

QTOF-MS suggest a derivative with a molecular weight of 176, indicating the attachment of –OH to the benzene ring is part of another unidentified byproduct, not 4IBPE.

Photosensitized decarboxylation (photodecarboxylation) of NSAIDs in non- environmental systems has been extensively probed (Bosca, 2001; Bosca, 1994; Monti,

1997; Budac 1992; Castell et al., 1987). Specifically, Castell indicates that IBAP and 1-

(4-isobutylphenyl)ethanol, are the result of reaction with dioxygen. Therefore, the structure of both identified byproducts (isobutylacetophenone and the ‘4IBPE’) indicate that in natural systems ibuprofen, when irradiated, photodecarboxylates as had been demonstrated in methanol (Castell, 1987) and with other similarly structured NSAID’s such as ketoprofen (Boscá et al., 2001) (Scheme 4.1). Additionally we have also shown that hydroxylation of the benzene ring takes place (Figure 4.7). Castell et al. documents equal percent yield of IBAP and 1-(4-isobutylphenyl)ethanol, upon irradiation of ibuprofen in methanol. In contrast, our environmental systems demonstrate that the

4IBPE derivative is the dominant photoproduct.

65 Fulvic acid composition can also affect photoinduced byproduct formation. At

[Ibuprofen]o = 10 µM in the presence of SRFA and OWCFA, HPLC peak area units of the major byproduct do not significantly change when compared to direct photolysis

(Figure 4.10). However, in the presence of PLFA the major byproduct peak formation increases relative to direct photolysis and in the presence of OWCFA and SRFA (Figure

4.10, triangles). I speculate that PLFA is forming different short-lived reactive transient/s compared to SRFA and OWCFA, which produce more of the photo-derivative/s. The

OH• scavenger isopropanol does not slow the rate of degradation at [Ibuprofen]o = 10 µM in the presence of PLFA (Chapter 3) nor do degradation experiments in the presence of

PLFA and isopropanol slow the generation of the byproduct (Figure 4.11: dark triangles).

It is then thought that photo-derivative formation is OH• independent.

Formation of the major byproduct peak (Figure 4.1b) in the presence of fulvic acids at the low initial ibuprofen concentrations (0.1 µM) does not show the same trend

(Figure 4.11). Indeed, no correlation is apparent between byproduct peak area generation and aqueous medium used (Milli-Q, PLFA, OWCFA, and SRFA). Moreover, the linearity of photo-derivative generation demonstrated at 10 µM (Figure 4.10) is not replicated at the lower concentration (Figure 4.11). As discussed in Chapter 3, low concentration experiments demonstrate the dominance of long- and short-lived reactive transients at this concentration, and variance in low concentration ibuprofen degradation kinetics suggests importance of at least two (dioxygen dependent and dixoygen independent) long-lived species from each fulvic acid. As NMR spectroscopy indicates the major derivative peak includes more than one compound, at low concentrations

66 reactive pathways responsible for the photodegradation of ibuprofen become more complicated and as a result the byproduct formation becomes more complex.

4.4. Conclusion

Irradiation of ibuprofen in natural systems produces multiple photo-derivates, two of which I have assigned a chemical structure, isobutylacetophenone and ‘4IBPE’.

Additionally, NMR demonstrates the addition of a more electronegative moiety to the former ibuprofen benzene ring as a third structural assignment. NMR spectra of the dominant HPLC byproduct peak indicates more than one compound eluding at this time, one of which is 4IBPE. This peak area production upon photolysis of [Ibuprofen]o = 10

µM, is magnified in the presence of PLFA when compared to direct photolysis, SRFA, and OWCFA demonstrating the importance of DOM origin in photo-derivative production. Furthermore, I conclude that at environmentally relevant concentrations of ibuprofen DOM produced short and long-lived excited state species further complicate photo-derivative prodcution.

67 4.5 References

Boscá, F.; Marín M. L.; Miranda M. A. (2001) Photoreactivity of the non-steroidal anti- inlammatory 2-arylpropionic acids with photosensitizing side effects. Photochem. Photobiol. 74, 637-655.

Boscá, F.; Miranda, M.A.; Carganico, G.; Mauleón, D. (1994) Photochemical and photobiological properties of ketoprofen associated with the benzophenone chromophore. Photohemistry and Photobiology. 60, 96-101.

Budac, D.; Wan, P. (1992) Photodecarboxylation: mechanism and synthetic utility. J. Photochem. Photobio. A: Chem. 67, 135-166.

Castell, J.V.; Gomez, M.J.; Miranda, M.A.; Morera, I.M. (1987) Photolytic Degradation of Ibuprofen. Toxicity of the Isolated Photoproducts on Fibroblasts and Erythrocytes. Photochem Photobiol. 46, 991-996.

Edhlund, B.L.; Arnold, W.A.; McNeill, K. (2006) Aquatic photochemistry of nitrofuran antibiotics. Environ. Sci. Technol. 40, 5422-5427.

Monti, S.; Sortino, S.; De Guidi, G.; Marconi, G. (1997) Photochemistry of 2-(3- benzoylphenol)proponic acid (ketoprofen) Part 1: A picosecond and nanosecond time resolved study in aqueous solution. J. Chem. Soc., Faraday Trans. 93, 13, 2269-2275.

Pérez-Estrada, L.A.; Malato, S.; Gernjak, W.; Agüera, A.; Thurman, E.M.; Fernández- Alba, A.R. (2005) Photo-fenton degradation of diclofenac: Identification of main intermediates and degradation pathway. Environ. Sci. Technol. 39, 8300-8306.

Zorita, S.; Barri, T.; Mathiasson, L. (2007) A novel hollow-fibre microporous membrance liquid-liquid extraction for determination of free 4-isobutylacetophenone concentration at ultra trace level in environmental aqueous samples. J. Chromatogr. 1157, 30-37.

68 4.6 Tables

e z 10

m 2

- - - - 9 g m ------

om 1 a t r r n f i n

s t

11) 1 5 . . -

- - - 1 1 9 ------( 9 1 oduc

1 2 ks a

bypr

8 9 9 pe .

. .

- 6 2 1 ------ous on 6 9 7 r i

1 1 e 1 nt m

9 nu 8 9 i

. . .

- -

2 1 6 4 - - - - of

6 6 7 1 e dom 1 1 1

) nc S e

s M e 1 1 1 1 8 8 - ......

- pr C 1 1 1 1 9 5 - - - -

6 3 6 6 4 0 . G ( 1 1 1 1 1 2 e he t

y bl r a t ng c

e i

9 8 1 9 1 8 ......

m 7

r - om 6 9 5 6 5 3 - 1 ppl i - - r 0 1 1 4 3 6 a t 1

c 2 1 1 1 1 1 onf c pe not

h s

8 9 9 9 9 8 1 s e 9 ......

s 5 . - - 48 9 8 1 0 0 5 1 a 0

0 not 5 1 6 2 2 7 6 1 d 9 M e 1 1 1 1 1 1 1 e - D

yz ) -

phy

ol (

1 9 8 2 a

7 . . . . . 7 1 9 . 2 1 7 8 4 2 9 1 6 4 6 9 9 7 ogr 7 0 3 1 1 1 1 phot 5 t 176

1 1 1 1 a ~

on i w om

9 8 m 9 hr

1 9 1 9 . . . 5 4 l ol . . . . 7 7 C 4 6 6 s 1 0 a

0 2 5 0 7 7 r 0 s 7 7 1 1 9 4 5 9 n a 1 1 1 e ve G e of s

1 4.

on, 0 1 bupr e i 1 2 3 4 5 6 7 8 9 i l 1 1

ut ab M ol T µ s

69 4.7 Figures

Figure 4.1 a. UV-Vis scan of photolyzed ibuprofen solution at various time points (hours) demonstrating byproduct accumulation at 264 nm; b. HPLC chromatogram (λ = 264 nm) of solution photolyzed for 48 hours (ibuprofen retention time ~5minutes; “major” photolysis byproduct retention time ~7 minutes, additional byproduct retention time ~8 minutes).

Figure 4.2 1H 1-D NMR spectra of 500 µM ibuprofen.

Figure 4.3 1H NMR of 10 µM photolyzed (48 h) ibuprofen solution.

72 IBAP

MW = 176.2 4IBPE

Figure 4.4 Liquid Chromatography Electrospray Ionization Quadropole Time of Flight Mass Spectrometer (LC-ESI-QTOF-MS) of 10 µM ibuprofen solution photolyzed 48 h and proposed byproduct structures.

Figure 4.5 Light absorption of 100 µM Isobutylaectophenone (IBAP) in Milli-Q.

Figure 4.6 1H NMR of purchased isobutylacetophenone standard (100 µM).

Figure 4.7 1H 1-D NMR spectra of isolated major byproduct fraction of photolyzed ibuprofen solution and proposed byproduct structure.

Figure 4.8 1H NMR COSY of HPLC waste line effluent (retention time 7 minutes) of photolyzed ibuprofen solution (48 h).

Figure 4.9 1H NMR COSY of 500 µM purchased isobutylacetophenone.

Figure 4.10 Formation of major byproduct (retention time 7 minutes) measured in peak area units (HPLC with a UV detector) upon the photolysis of 10 µM ibuprofen (Direct) in the presence of Pony Lake Fulvic Acid (PLFA), Old Woman Creek Fulvic Acid (OWCFA), and Suwannee River Fulvic Acid (SRFA) with 20 mM isopropanol (Iso) and suboxic conditions.

Figure 4.11 Formation of major byproduct (retention time 7 minutes) measured in peak area units (HPLC with a UV detector) upon the photolysis of 0.1 µM ibuprofen (Direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Old Woman Creek Fulvic Acid (OWCFA), and Suwannee River Fulvic Acid (SRFA) with 20 mM isopropanol (Iso) and suboxic conditions.

80 4.8 Schemes

Scheme 4.1 Proposed reactions 1) Ibuprofen photodecarboxylation 2) Oxygen addition to carbon centered radical and subsequent rearrangement resulting formation of isobutylacetophenone 3) Hydroxylation of carbon centered radical and conversion of isobutyl side chain to alkene derivative

CHAPTER 5

CAFFEINE AS A WASTEWATER TRACER: A PHOTOCHEMICAL ANALYSIS

5.1 Introduction

A molecular tracer is defined as a conservative compound used to distinguish the contributions of various sources to hydrologic processes. Required properties of a good tracer include ubiquitous use to the source in question, that it is non-volatile and highly soluble, and it is recalcitrant in nature (Eganhouse, 1997; Strauch, 2008). The xanthine alkaloid caffeine (1,3,7-trimethylxanthine) has been proposed as a molecular tracer of wastewater contributions to streatmflow (Seiler, 1999; Strauch, 2008) due to its very low octanol-water partition coefficient (log Ko/w 0.01) (Gossett, 1983), high aqueous solubility (22 mg/mL), and ubiquity. Indeed, caffeine is one of the most popular if not the most popular drug in the world as it is contained in coffee, tea, chocolate, soft drinks

(Bunker and Williams, 1979; Ogunseitan, 1996) and pharmaceuticals (prescription and over the counter) (Graham, 1978). For example, caffeine concentrations in home-brewed coffee range from 8 to 91 mg per cup (Gilbert et al., 1976), and the average concentration in cola beverages is 47 mg per serving. The levels are somewhat lower with an average of

10 to 40 mg contained in one serving of tea (Bunker and Williams, 1979).

82 Once ingested, 3% of the total caffeine intake is excreted in urine (Tang-Liu,

1983). Despite this small amount and efficient (> 99%) wastewater treatment removal

(Strauch et al., 2008; Buerge, 2003) caffeine has been documented extensively in the environment. For example, Buerge et al. (2003) found between 2 to 250 ng/L caffeine in

Swiss surface waters while Gardinali and Zhou (2002) document between 6 and 41 ng/L in Biscayne Bay, FL and more recently between 10 to 100 ng/L has been discovered in

German surface waters (Strauch, 2008). Caffeine is also found in groundwater with concentrations ranging from 10 to 230 ng/L (Strauch, 2008; Seiler, 1999).

Buerge et al. (2003) evaluated caffeine as an anthropogenic marker of wastewater contamination in surface waters (Buerge, 2003) and found indirect photolysis via the hydroxyl radical (OH•) as the primary pathway of its degradation. This was concluded by a comparison of the 2nd-order rate constant of reaction between OH• as well as observed caffeine degradation in a batch photo-reactor. Despite this important removal mechanism for caffeine in sunlight waters, the authors conclude that caffeine is a suitable wastewater treatment tracer, but identify a need for more precise information on the compounds’ behavior in natural waters (Buerge et al., 2003).

Hydroxyl radical (OH•) can be produced from various photosensitizers upon irradiation including nitrate (NO3), dissolved organic matter (DOM; outlined in detail in previous chapters), and the presence of Fe(II) and DOM by the photo-Fenton pathway

(Southworth and Voelker, 2003; White et al., 2003). The photo-Fenton process plays an important role in the production of OH• when slightly acidic, Fe-rich natural waters are irradiated. Southworth and Voelker (2003) document an OH• steady-state concentration

83 -16 ([OH•]ss) of 2.3 × 10 M at pH 6 with 10 µM Fe in the presence of organic matter via the pathway:

- Fe(II) + H2O2 → Fe(III) + OH + OH• (1)

Where hydrogen peroxide (H2O2) is formed from irradiated DOM via superoxide dismutation and Fe(II) is formed from photoreduction of Fe(III)-oxides or Fe(III)-DOM complexes (Southworth and Voelker, 2003; White et al., 2003).

Dissolved organic matter can produce reactive transients other than OH•, such as

1 - H2O2 (mentioned above), as well as singlet oxygen ( O2), superoxide (O2• ), peroxyl

(ROO•), and phenoxyl radicals (Canonica and Freiburghaus ,2001; Buschmann et al.,

2005a/b; Faust and Hoigne, 1987; Haag and Hoigne, 1986). Moreover, DOM from different sources vary in chemical composition. Therefore, type of reactive transients formed can differ based upon the composition of the target DOM. For example, [OH•]ss from a microbially-derived fulvic acid such as the International Humic Substance Society

(IHSS) reference material Pony Lake Fulvic Acid (PLFA) is 8.8 × 10-17 M while a terrestrially derived fulvic acid such as such as the IHSS Suwannee River Fulvic Acid

-16 (SRFA) generates an [OH•]ss of 2.9 × 10 M (White, 2003).

In addition to a source specific mixture, reactive transients from DOM have various lifetimes (demonstrated in Chapter 3). Using alkyl and methoxy phenols as probe molecules, Canonica and Freiburghaus (2001) found enhanced photo-degradation (× 2-3) in the presence of DOM isolates and natural waters at low (0.1 µM) probe molecule concentrations relative to higher concentrations (5 µM). They attributed this phenomenon to the reaction of short-lived scavengers at higher probe concentrations and a

84 combination of short- and long-lived reactive transients at lower, more environmentally relevant concentrations (Canonica and Hoigne, 1995; Canonica et al. 1995; Canonica and

Freiburghaus, 2001). Therefore, when evaluating the photochemistry of an organic contaminant, in relation to its acceptability as a wastewater marker, concentration of the contaminant as well as DOM source specific reactive transients must be considered.

To my knowledge, no investigation has probed, in depth, caffeine photochemistry in natural waters. Such a study is necessary to evaluate the efficacy of caffeine as a wastewater treatment marker. In this study I investigated the role of the photosensitizers

- NO3 , DOM, and Fe(III) on the photodegradation of caffeine in the presence of fulvic acids. Additionally, I probed the effect of the initial concentrations 10 µM and 0.1 µM of caffeine on degradation kinetics. Finally, I investigated oxygen dependent reactive transients under suboxic experimental conditions.

5.2 Methods

5.1.1 Materials

All materials used (methanol, isopropanol) were reagent grade or better. Caffeine

99% reagent grade was obtained from Acros Organics. Anhydrous Ferrous Chloride

(FeCl3) was obtained from Fisher Scientific. We isolated a fulvic acid from a National

Estuarine Research Reserve wetland in Huron, OH according to standard XAD-8 non- ionic resin protocol (Leenheer, 1981; Thurman & Malcolm, 1981), a protocol further defined in Chapter 3. We deem this fulvic acid Old Woman Creek Fulvic Acid

(OWCFA). Additional fulvic acid isolates from Suwannee River, GA (SRFA) and Pony

85 Lake, Antarctica (PLFA) were acquired from the International Humic Substance Society

(IHSS).

5.1.2. Photolytic Reactions

A 10 mM caffeine stock solution was prepared in Milli-Q (18MΩ, Millipore) and diluted down to the desired µM levels in the appropriate matrix (Milli-Q, fulvic acid, nitrate, etc.) with pH adjustment to 7 (±0.1) with NaOH and/or HCl. Select solutions were prepared anoxically by sparging with argon for 1 min/mL and transferring to photolysis tubes within a glove box (95% N2 / 5% H2). The photo-Fenton pathway was probed by introduction of FeCl3 into reaction solutions by a spike of iron stock (0.04 M

HCl and 0.4 mM Fe(III)).

As previously described, air tight quartz tubes (1 cm pathlength capped with Teflon- lined quartz lids) were used for photolysis, screening wavelengths <290 nm. Photolysis of ibuprofen was conducted using a solar simulator (Atlas Suntest CPS+) with a Xenon arc lamp at 25°C and lamp energy of 500 Watts for a time duration equivalent to a minimum of one half-life. To correlate our reaction rates to those that can occur in natural sunlight we used the procedure defined in Chapter 2 finding the solar simulator is 3.7 times stronger than natural sunlight. Thus, caffeine degradation rates should be interpreted within the context of the light source used in this study unless text indicates the correction factor has been applied. Additionally, all reaction kinetics obeyed pseudo first order fits. Chemical actinometry was conducted as described in Chapter 2 and showed no significant light change over the course of each experiment (Appendix A).

86 HPLC is as described in Chapter 3 with the following change: 50/50 methanol/water mobile phase and 275 nm UV absorbance detection wavelength. OWCFA and PLFA

[OH•]ss used were determined and reported in Chapter 3.

5.3 Results and Discussion

Caffeine degrades slowly by direct photolysis (half life of approximately 150 hours) at an initial concentration of 10 µM, and the addition of fulvic acids accelerates this degradation (Figure 5.1). Specifically, Old Woman Creek Fulvic Acid (OWCFA) and

Suwannee River Fulvic Acid (SRFA) increase the rate threefold while Pony Lake Fulvic

Acid (PLFA) accelerated the rate even further (× 4) (Figure 5.1). Based upon our

-17 -16 measurements of [OH•]ss (PLFA 8.80 × 10 M; OWCFA 2.42 × 10 M) and of SRFA

2.90 × 10-16 M (White, 2003), I expected slower degradation in the presence of PLFA if degradation was solely caused by the reaction between caffeine and OH•, but the accelerated rate suggests that other pathways might be involved.

I was able to delineate the hydroxyl radical contribution to overall photodegradation of caffeine using [OH•]ss values. I express the rate of caffeine photodegradation as follows:

-d[Caffeine]/dt = kobs[Caffeine] = kdir[Caffeine] + kOH•[Caffeine][OH•]ss (2) where kdir is the observed first order rate constant for the direct photolytic pathway, t is time, and kobs is the observed rate constant in the presence of fulvic acids. I estimate the pseudo first-order rate constant (k’) for the reaction between OH• and caffeine:

k’ = kOH• • [OH•]ss (3)

87 nd using [OH•]ss data for each fulvic acid and the 2 order rate constant between OH• and

9 -1 -1 caffeine (6 × 10 M s : kOH•) (Shi, 1981). The resulting calculated rate of reaction between OH• and caffeine (k’) when summed with kdir provides a calculated estimate of

-1 -1 -1 kobs (PLFA 0.006 h ; SRFA 0.010 h ; OWCFA 0.009 h ). When comparing these values to my experimental rate constants, the photodegradation of caffeine in SRFA and

OWCFA solutions have very similar values (k of 0.0115 h-1 and 0.0101 h-1, respectively) indicating the pathway in the presence of these two fulvic acids is OH• dominated. In

-1 contrast, the experimental PLFA rate of degradation (kobs 0.0148 h ) is threefold higher than the calculated estimate, indicating that some other reactive transient/s have a strong influence of caffeine photodegradation.

To further probe this phenomenon, the known OH• scavenger, isopropanol, was added to these reactions. The results confirm that OH• plays a dominant role (Figure 5.2) yet they are inconsistent with the above calculation because degradation slowed to the rate of direct photolysis for each fulvic acid, including Pony Lake (Figure 5.2). I speculate that in addition to OH•, isopropanol may scavenge other responsible reactive transients emitted from irradiated PLFA, but assessing the effectiveness of isopropanol as a OH• scavenger is beyond the scope of this project. Nonetheless, these data do demonstrate the importance of nitrate generated OH• in the photodegradation of caffeine in natural waters, confirming Buerge et al., (2003).

- NO3 generated OH• chemistry further accelerates caffeine degradation (Figure 5.3).

- As discussed in Chapter 2, agricultural watershed NO3 levels can vary from 100 µM to 1 mM (Miller and Chin, 2002; Miller and Chin, 2005; Brezonic and Fulkerson-Brekken,

88 - 1998). As NO3 is a known OH• generator (Zepp, 1987), caffeine photolysis is accelerated with the addition of 100 µM nitrate and continues to increase with increasing

- NO3 concentration (up to 900 µM) in the presence of each fulvic acid (Figure 5.3). As discussed earlier, the rate of caffeine degradation in the presence of PLFA is faster than with OWCFA and SRFA (Figure 5.1). Not surprisingly, when compared to other FAs caffeine degradation is fastest in the presence of nitrate and PLFA (Figure 5.3: triangles).

This phenomenon appears to be specific to this unique fulvic acid (entirely microbially derived), which generates an unidentified reactive transient that is more reactive than

– NO3 .

- - Rates of degradation in the presence of NO3 (Figure 5.3) are higher, kobs up to 0.05 h

1 -1 , relative to direct photolysis (kobs 0.0039 h ) and reaction with fulvic acids (kobs up to

-1 - 0.015 h ). Indeed, caffeine in the presence of PLFA and 884 µM NO3 yields a pseudo first order rate constant of 0.05 h-1, representing a half-life of approximately 13 hours (vs

150 hours for direct photolysis). The correction factor (defined in Chapter 3) to translate our solar simulator to natural sunlight, deems this a half-life of 48 hours. Therefore,

– caffeine degradation from NO3 generated OH• is significant and should be taken into account when using caffeine as a wastewater marker in agricultural watersheds.

In addition to the role of nitrate, I examined the effect of the photo-Fenton pathway on caffeine degradation kinetics. Fe(III) was added to solutions containing caffeine and

OWCFA in order to activate the photo-Fenton pathway. At pH 7 and high Fe(III) concentrations (31 µM, the upper limit of dissolved Fe(III) in natural waters), there is no statistical increase related to OWCFA solutions in the absence of iron. These data show that the photo-Fenton pathway plays a minimum role in caffeine degradation (Figure 5.4). 89 Nontheless, I suspect it would play a more important role in caffeine photolysis in

-16 slightly acidic natural waters as [OH•]ss at pH 6 is on the order of 10 M (Southworth and Voelker 2003).

Photolysis at low caffeine concentrations were performed to investigate the role of long lived reactive transients responsible for degradation. Caffeine has been documented at a range of 1 × 10-5 to 1 × 10-3 µM (2 to 250 ng/L) in surface and groundwater (Buerge et al., 2003). Thus, photolyses were conducted an initial concentration of 0.1 µM caffeine, the limit of detection for our system which is still two orders of magnitude higher than the highest reported concentration in natural waters. At 0.1 µM, caffeine photolysis is accelerated in the presence of each FA relative to 10 µM (Figure 5.5). SRFA

-1 and PLFA rate constants at this level were statistically the same rate (kobs ~0.020 h ) or approximately a 24 hour half-life in the solar simlator. Relative to their respective rates at

10 µM, SRFA increases × 2.5 while the PLFA acceleration rate almost doubles (× 1.8).

At the low caffeine concentrations, photolysis in OWCFA solutions accelerates as well

-1 (kobs of 0.017 h or × 1.5 faster than the corresponding 10 µM experiment) (Figure 5.5).

As discussed in Chapter 3 this indicates that both long and short lived species play a role in degradation at low concentrations; while at high concentrations the short lived species are dominant in facilitating transformation. Given the difference in rate constants for each fulvic acid, I suspect more than one long lived reactive transient plays a role.

To probe the identity of these long lived species, anoxic experiments were performed at low concentrations (0.1 µM) with each fulvic acid. Compared to oxic conditions,

-1 suboxic experiments slow photolysis rates in the presence of SRFA (kobsoxic 0.028 h vs

90 -1 -1 -1 kobssuboxic 0.013 h ) and OWCFA (kobsoxic 0.017 h vs kobssuboxic 0.009 h ) while the degradation rate in the presence of PLFA remains statistically the same (Figure 5.6). This indicates one or more dioxygen-dependent long-lived transient/s produced by OWCFA and SRFA increases caffeine degradation at low concentrations, such as peroxyl or phenoxyl radicals. Faust and Hoigné (1987) found photo-reactivity of phenols and DOM, most likely with the peroxyl radical, increased with increasing phenol ring substitution with electron-donating methyl groups (Faust and Hoigné, 1987). Since both phenoxyl and peroxyl radicals are strong electrophiles, it would stand to reason that caffeine, bearing two methyl groups (–CH3), would react with both species (Figure 1.2).

As discussed earlier (Chapter 1 & 3) the two DOM end members, SRFA and PLFA, represent terrestrially-derived and microbial-derived organic matter, respectively while

OWCFA is some combination of both sources. Correspondingly, SRFA has a more aromatic composition than PLFA (Thurman and Malcom, 1981; Chin et al., 1994), while

OWCFA falls somewhere in between these two substances. This varying aromaticity is consistent with the conclusion that fulvic acids (or in this case SRFA) containing a higher degree of aromatics more effectively degrade compounds sensitive to the long lived phenoxyl radicals at environmentally relevant concentrations. However, since the suboxic rate of degradation is not slowed to the rate of direct photolysis for any of the fulvic acids, I suspect an additional reactive transient, whose formation is dioxygen independent, plays a role.

91 5.4 Conclusion

At 10 µM, caffeine is susceptible to direct photolysis and the presence of fulvic acids accelerates its photodegradation rate, of which a significant fraction can be attributed to

OH•. Isopropanol slows degradation in the presence of fulvic acids, confirming the findings of Buerge et al. (2003) and the importance of OH•. My observation coupled with estimates of hydroxyl radicals influence on the indirect photolytic rate indicate isopropanol could be shutting down other reactive transients generated by PLFA as well.

When present at concentrations reported in agricultural watersheds (100-1000 µM), nitrate generated OH• accelerates the rate of photolysis even thought DOM is an excellent OH• scavenger (Torrents et al., 1997); half-life up to 48 hours in natural sunlight. At current experimental conditions (pH 7, 31 µM iron) the photo-Fenton pathway and subsequent generation of OH• is unimportant. However, I suspect that this pathway will play a larger role in more acidic natural waters. Therefore, particular attention should be paid to OH• induced photochemical degradation of caffeine in agricultural watersheds and possibly slightly acidic natural waters.

Caffeine photolysis reactions at lower, more environmentally relevant concentrations

(0.1 µM) in the presence of each fulvic acid significantly increase degradation (vs 10 µM in the presence of each fulvic acid), with the fastest half-life of approximately 90 hours in natural sunlight. This acceleration demonstrates the importance of long lived reactive species. This acceleration is variable for each fulvic acid indicating that different long lived reactive transient(s) affect caffeine photodegradation in the presence of each fulvic

92 acid. Suboxic experiments at lower concentrations (0.1 µM) slow the reaction kinetics demonstrating that dioxygen in needed to generate the long lived species, which I suspect to be phenoxyl or peroxyl radicals. Nonetheless, suboxic conditions only partially slowed the degradation kinetics relative to direct photolysis, indicating the existance of another reactive transient responsible for the photodegradation of caffeine at low concentrations.

93 5.5 References

Brezonic P.L., Fulkerson-Brekken J. (1998) Nitrate-induced photolysis in natural waters: Controls on concentrations of hydroxyl radical photo-intermediates by natural scavenging agents. Environ. Sci. Technol., 32, 3004-3010.

Bunker, M.; McWilliams, M. (1979) Caffeine content of common beverages. J American Dietetic Association, 74, 28-32.

Buerge, I.J.; Poigner, T.; Müller, M.D.; Buser, H.R. (2003) Caffeine, and anthropogenic marker for wastewater contamination of surface waters. Environ. Sci. Technol., 37, 691- 700.

Canonica, S.; Freiburghaus, M. (2001) Electron-rich Phenols for probing the photochemical reactivity of freshwaters. Environ. Sci. Technol. 35, 690-695.

Canonica, S.; Hoigné J. (1995) Enhanced oxidation of methoxy phenols at micromolar concentration photosensitized by dissolved organic matter. Chemosphere. 30, 2365-2374.

Chin, Y.P.; Aiken, G.; O’Loughlin, E. (1994) Molecular weight, polydispersity, and spectroscopic properties of aquatic humic substances. Environ. Sci. Technol., 28, 1853- 1858.

Eganhouse, R.P. (1997) Molecular markers and environmental organic geochemistry: An overview. 671, 1-20.

Faust, B.C.; Hoigné, J. (1987) Sensitized photooxidation of phenols by fulvic acid and in natural waters. Environ. Sci. Technol. 21(10), 957-964.

Gardinali, P.R.; Zhao, X. (2002) Trace determination of caffeine in surface water samples by liquid chromatography-atmospheric pressure chemical ionization-mass spectrometry (LC-APCI-MS). Environ. Int., 28, 521-528.

Gossett, R.; Brown, D.A. (1983) Predicting the bioaccumulation of organic compounds in marine organisms using octanol/water partition coefficients. Mar. Pollut. Bull. 14 (10) 387-392.

Graham, D.M. (1978) Caffeine- Its identity, dietary sources, intake, and biological effects. Nutrition Reviews. 36 (4) 97-102.

94 Haag, W.R.; Hoigné, J. (1986) Singlet oxygen in surface waters. 3. Photochemical formation and steady-state concentrations in various types of waters. Environ. Sci. Technol., 20, 341-348.

Miller, P.L.; Chin, Y.P. (2002) Photoinduced degradation of carbaryl in a wetland surface water. J. Agric. Food Chem., 50, 6758-6765.

Miller, P.L.; Chin, Y.P. (2005) Indirect photolysis promoted by natural and engineered wetland water constituents: Processes leading to alachlor degradation. Environ. Sci. Technol., 39, 4454-4462.

Ogunseitan, O.A. (1996) Removal of caffeine in sewage by Pseudomonas putida: Implications for water pollution index. World J. Microb. Biot., 12, 251-256.

Seiler, R.L.; Zaugg, S.D.; Thomas, J.M.; Howcroft, D.L. (1999) Caffeine and pharmaceuticals as indicators of waste water contamination in wells. Ground Water. 37 (3), 405-410.

Shi, X.; Dalal, N.S. (1991) Antioxidant behavior of caffeine: efficient scavenging of hydroxyl radicals. Fd. Chem. Toxic. 29 (1), 1-6.

Southworth, B.A.; Voelker, B.M. (2003) Hydroxyl radical production via the photo- Fenton reaction in the presence of eachfulvic acid. Environ. Sci. Technol. 37, 1130-1136.

Strauch G.; Möder, M.; Wennrich, R.; Osenbrück, K.; Gläser, H.R.; Schladitz, T.; Müller, C.; Schirmer, K.; Reinstorf, F.; Schirmer, M. (2008) Indicators for assessing anthropogenic impact on urban surface and groundwater. J. Soils Sediments. 8 (1), 23-33.

Tang-Liu, D.D.; Williams, R.L.; Riegelman, S. (1982) Disposition of caffeine and its metabolites in man. J. Pharm. Exp. Ther. 224 (1), 180-185.

Telo, J.P.; Vieira, A.J.C. (1997) Mechanism of free radical oxidation of caffeine in aqueous solution. J. Chem. Soc. Perkin Trans. 2, 1755-1757.

Thurman, E.M.; Malcolm, R.L. (1981) Preparative isolation of aquatic humic substances. Environ. Sci. Technol., 15 (4), 463-466.

Torrents A, Anderson BG, Bilboulian S, Johnson WE, Hapeman CJ. 1997. Atrazine photolysis: Mechanistic investigations of direct and nitrate-mediated hydroxyl radical processes and the influence of dissolved organic carbon from the Chesapeake Bay. Environ Sci Technol 31: 1476-1482.

White, E. (2003) Determination of photochemical production of hydroxyl radical by dissolved organic matter and associated iron complexes in natural waters. M.S. Thesis. The Ohio State University, Columbus OH, USA. 95

Zepp, R.G.; Hoigné, J.; Bader, H. (1987) Nitrate-Induced Photooxidation of trace organic chemicals in water. Environ. Sci. Technol. 21: 443-450.

96 5.6 Figures

Figure 5.1 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) in Milli-Q (Direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek -1 Fulvic Acid (OWCFA) where kobs (h ) is the observed rate constant per hour.

Figure 5.2 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) in Milli-Q (Direct) and in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) with and without 20 mM isopropanol (Iso) and with visual trendlines.

Figure 5.3 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) photolysis in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid -1 (OWCFA) with and without various levels of nitrate (NO3) where kobs h is the observed rate constant per hour.

Figure 5.4 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM) photolysis in the presence of Old Woman Creek Fulvic -1 Acid (OWCFA) with and without 31 µM iron where kobs h is the observed rate constant per hour.

100

Figure 5.5 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (10 µM; 0.1 µM) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid -1 (OWCFA) where kobs h is the observed rate constant per hour.

101

Figure 5.6 Photoinduced pseudo first order (Ln(Concentration/Initial Concentration)) degradation of caffeine (0.1 µM) in the presence of Pony Lake Fulvic Acid (PLFA), Suwannee River Fulvic Acid (SRFA), and Old Woman Creek Fulvic Acid (OWCFA) -1 under oxic and suboxic conditions where kobs h is the observed rate constant per hour.

102

CHAPTER 6

CONCLUSIONS AND FUTURE RESEARCH

6.1 Conclusions

The photochemical degradation of the organic contaminants pyrene, phenanthrene, naphthalene, ibuprofen and caffeine occurs through direct photolysis. Direct photolysis is the dominant mechanism for pyrene’s photo-fate but for the other compounds it plays a minor role. In the presence of natural water samples, the phenanthrene degradation rate, when compared to its direct photolysis rate, increased in Gary, Indiana water (GIN), but decreased in the Wilmington, North Carolina water (WNC). The GIN water sample contains relatively higher nitrate levels (0.046 mM vs 0.205 mM) while the WNC water sample contains relatively higher dissolved organic carbon (9.29 mg/L vs 6.73 mg/L), and I attribute the lower rate of degradation in the presence of WNC water to light screening by DOC, while I believe that the faster reaction rate observed for the Indiana water is the result of hydroxyl radical (OH•) chemistry generated by nitrate photolysis.

This is confirmed by 1) correcting the slightly lower phenanthrene degradation rate in the presence of WNC water for light screening effects and finding the rate statistically the same as the direct photolysis experiments, and 2) observing a statistical increase in

103 degradation of phenanthrene when nitrate (at 0.2 and 0.4 mM) was added to the WNC water sample.

Direct photolysis plays an important role in the degradation of naphthalene. However, calculations indicate that OH• chemistry will also react with the compound as well in a

- significant manner. Additionally, the irradiation of naphthalene and NO3 (in the presence of a phosphate buffer) increases the rate of degradation (× 2), further confirming the importance of the hydroxyl radical. Lack of water sample prevented me from further probing naphthalene degradation in the presence of OH• generating natural waters (GIN water). Overall, photo-reaction rates decrease for the lower molecular weight PAHs as the fastest naphthalene photolytic rate was roughly two orders of magnitude slower than the photolysis of pyrene. Therefore, direct photolysis is an important pathway for the degradation of larger ringed PAHs in sunlit waters due to their higher extinction coefficients at wavelengths present in sunlight, while waters rich in nitrate have increased photo-reactivity toward lower ring PAHs through reaction with OH•.

When ibuprofen and caffeine (10 µM) are irradiated in the presence of fulvic acids, degradation occurs more rapidly when compared to direct photolysis rates. Rate constants for ibuprofen degradation in the presence of Pony Lake (PLFA), Suwannee River

(SRFA), and Old Woman Creek (OWCFA) fulvic acids all increase (× 5). The introduction of isopropanol slows the rate of ibuprofen degradation in the presence of

OWCFA and SRFA (46% and 64%, respectively) but the indirect photolysis of ibuprofen in the presence of PLFA is unchanged with the introduction of isopropanol. These results are consistent with the level of [OH•]ss measured for each fulvic acid. Conversely,

104 caffeine degradation in the presence of Old Woman Creek Fulvic Acid (OWCFA) and

Suwannee River Fulvic Acid (SRFA) increases threefold while Pony Lake Fulvic Acid

(PLFA) solutions increase the rate by a factor of four, results that are not consistent with its [OH•]ss concentration. The presence of isopropanol in these experiments slows kinetics to the rate reported for direct photolysis, demonstrating the importance of OH• in the photodegradation of caffeine. The data, however, is still inconsistent with respect to

PLFA where I believe that a different reactive transient (in addition to OH•) is responsible for caffeine degradation that is also scavenged by isopropanol.

The effect of nitrate generated OH• chemistry was also studied whereby caffeine (10

µM) was irradaited in the presence of fulvic acids (PLFA, OWCFA, and SRFA) and nitrate (0.2 to 1 mM). I observed enhanced degradation rates to the most rapid rate observed in my experiments. Hydroxyl radical generated by the photo-Fenton pathway at neutral pH has no effect on the rate of caffeine degradation. Nonetheless, I suspect this pathway would become more relevant in slightly acidic waters because of caffeines reactivity towards OH•. Regardless, hydroxyl radical generated from organic matter and particularly nitrate plays a significant role in sunlight-induced degradation the organic contaminants caffeine and ibuprofen (10 µM). It is important to note, however, that

PLFA produces reactive transients in addition to OH• that can also be responsible for degradation at these concentrations.

When evaluating photochemical reactions contaminant concentration must be taken into account. I photolyzed both caffeine and ibuprofen at low, more environmentally relevant concentrations (0.1 µM) and in the presence of each fulvic acid. At the lower

105 concentration in the presence of each fulvic acid, all rates of degradation (caffeine and ibuprofen) increased when compared to experiments conducted at 10 µM photolyses.

Rates for ibuprofen increased slightly in the presence of OWCFA, × 2 in the presence of

SRFA, and × 5 in the presence of PLFA. Rates for caffeine increased relative to experiments at 10 µM, × 1.5 with OWCFA, × 1.8 with PLFA, and × 2.5 with SRFA. This demonstrates the importance of long-lived reactive species generated from fulvic acids that are capable of reacting with my compounds at more realistic environmental contaminant concentrations. Also, I speculate that caffeine is more susceptible to long- lived reactive transients produced by SRFA while ibuprofen reacts more rapidly with long-lived reactive transients produced by PLFA.

To probe these long-lived reactive transients, I performed suboxic photolysis experiments with [Ibuprofen, Caffeine]o of 0.1 µM. Ibuprofen rates of degradation slowed in the presence of PL and SR fulvic acids, but remained statistically the same in the presence of OWCFA relative to direct photolysis rates. In contrast, caffeine rates of degradation slowed in the presence of SR and OWC fulvic acids, while in the presence of

PLFA remained statistically the same. This indicates that both ibuprofen and caffeine are susceptible to long-lived reactive species whose formation is dependent upon dioxygen levels (peroxyl or phenoxyl radicals). Additionally, the lack of ibuprofen rate change in the presence of OWCFA and the same for caffeine in the presence of PLFA, indicates the reactive transients responsible for degradation are not the same for each compound.

The presence of isopropanol in experiments conducted at low concentrations of ibuprofen slows the degradation rates. Specifically, in the presence of OWCFA, isopropanol slows the reaction by ~half while SRFA and PLFA respectively decreased by 106 33% and 86%. Considering the lower [OH•]ss values for each PLFA, this indicates isopropanol is scavenging reactive transients other than hydroxyl radical a conclusion consistent with the caffeine photolysis data. However, further probing of the role of isopropanol is beyond the scope of this project.

Irradiation of ibuprofen in DOM solutions produces multiple photo-derivates in all photolyses experiments. I have assigned chemical structures to two of these photo- derivatives, isobutylacetophenone and ‘4IBPE’. A third structural assignment is the addition of a more electronegative moiety to the former ibuprofen benzene ring. NMR spectra of the dominant HPLC byproduct peak indicates the co-elution of two or more compounds, one of which is 4IBPE. During irradiation at higher concentrations (10 µM) this peak area is significantly larger in presence of PLFA when compared to direct photolysis, SRFA, and OWCFA demonstrating the importance of DOM composition on the fate of this compound production. At environmentally relevant concentrations of ibuprofen, the same trend is not observed demonstrating that the DOM produced short and long-lived excited state species complicates the mechanism of transformation.

6.2 Future Research

The identity of the reactive transients responsible for ibuprofen and caffeine degradation at low concentrations remains unclear. I would recommend to future research the use of molecule probes, such as 2,4,6-trimethylphenol, to determine the specific chemical species responsible (ie. phenoxyl and peroxyl radical scavengers). Once established I would probe the interaction of isopropanol and DOM sourced species to

107 assess the role of isopropanol as a specific OH• scavenger. Additionally, steady state concentrations for these transients would be made valuable from these experiments.

Based upon GC-MS data, additional photodegradation byproducts of ibuprofen photolysis probably exist. Both these compounds and the degradation products of caffeine photolysis would need to be identified. Given the ubiquity of both drugs, I believe quantifying byproduct concentrations in wastewater and environmental matrices worthy. Additionally, I would undertake a synthesis of 4EMDS to further confirm this structural assignment and probe the compounds toxicology.

As I did with ibuprofen, I would perform caffeine natural sunlight photolyses at Old

Woman Creek National Estuarine Research Reserve. I would investigate the production and yield of photoproducts under natural sunlight conditions as well as conditions produced in a solar simulator. I would employ state-of-the-art spectroscopic techniques

(NMR and LC-MS/MS; GC-MS/MS) to determine the identity of these derivatives. I would also further probe the role of caffeine as a wastewater marker by photolyzing nitrate in the presence of 0.1 µM caffeine (the lowest value that I can detect with confidence.

108

BIBLIOGRAPHY

Bertilsson S, Widenfalk A. 2002. Photochemical degredation of PAHs in freshwater and their impact on bacterial growth- Influence of water chemistry. Hydrobiologia 469: 23- 32.

Boreen, A. L.; Arnold W. A.; McNeill K. (2003) Photodegradation of pharmaceuticals in the aquatic environment: A review. Aquat. Sci. 65, 320-341.

Boscá, F.; Marín M. L.; Miranda M. A. (2001) Photoreactivity of the non-steroidal anti- inlammatory 2-arylpropionic acids with photosensitizing side effects. Photochem. Photobiol. 74, 637-655.

Brown, A.; McKnight, D.M.; Chin, Y.P.; Roberts, E.C.; Uhle, M. (2004) Chemical characterization of dissolved organic matter in Pony Lake, a saline coastal pond in Antarctica. Mar Chem. 89, 327-337.

Budac, D.; Wan, P. (1992) Photodecarboxylation: mechanism and synthetic utility. J. Photochem. Photobio. A: Chem. 67, 135-166.

Buerge, I.J.; Poigner, T.; Müller, M.D.; Buser, H.R. (2003) Caffeine, and anthropogenic marker for wastewater contamination of surface waters. Environ. Sci. Technol., 37, 691- 700.

109 Bunker, M.; McWilliams, M. (1979) Caffeine content of common beverages. J American Dietetic Association, 74, 28-32.