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Direct and Indirect Photochemical Degradation of Two Polycyclic Musk Fragrances and Two Polycyclic Aromatic Hydrocarbons in Natural Waters

Thesis

Presented in Partial Fulfillment of the Requirements for the Degree Master of Science in the Graduate School of The Ohio State University

Collin Patrick Ward, B.S.

Environmental Science Graduate Program

The Ohio State University

2010

Thesis Committee:

Yu-Ping Chin, Advisor

Heather Allen

John Lenhart

Collin Patrick Ward

2010

Abstract

The direct and indirect photolysis of two polycyclic musk fragrances (HHCB and AHTN) was examined in natural water systems. Both HHCB and AHTN are susceptible to photolytic degradation, with HHCB being the more refractory species.

In the presence of DOM, reaction rates increased significantly for both compounds.

Both compounds displayed a photoreactive dependence on the source of DOM.

Quenching and competition experiments suggest that OH plays a significant role in the photolytic degradation of HHCB, but this reaction is highly dependent on environmental [OH]ss. Proposed reaction schemes for HHCB with OH are presented. Quenching and photosensitizing experiments using rose bengal and D2O

1 suggest that O2 plays a prominent role in the photo degradation of AHTN.

1 Additionally, proposed reaction schemes for AHTN with O2 are presented.

Direct photolysis of two polycyclic aromatic hydrocarbons (phrenanthrene and naphthalene) was examined in two unique solvents. Organic surface microlayers

(OSM) and underlying “bulk water” samples were collected on two separate occasions at Old Woman Creek, Ohio. The glass plate technique was used to collect

OSMs. Analytical and spectroscopic techniques demonstrated that these samples were compositionally unique, although the OSMs were believed to be highly diluted. Slight degradation rate increases were observed in OSM solutions compared to bulk water,

ii but the increases were not statistically significant. Significant sample dilution and experimental design flaws were offered as suggestions for inconclusive results.

iii

Dedication

To Kiffy, Kevin, and Tucker. Thanks for Everything.

See You on the Flip Side.

Acknowledgements

Many people must be acknowledged for without them, this work would have never been accomplished. First and foremost, my advisor, Yo Chin. Thank you for the endless amounts of opportunities and for encouraging me to pursue graduate studies. For his relentless support in almost every aspect of this project, I thank Ryan

Fimmen, a friend and colleague. Thanks to Kris McNeill for his insight on several aspects of the project. I thank the additional members of my committee for their intellectual support and recommendation letters, Heather Allen and John Lenhart.

Many thanks go out to past and present members of the Chin research group.

Notably, Ale, Laura, Jenn, Sheela, and Marcy. I am indebted to you for countless hours of training and support. I thank Dave Klarer at OWC for the incredible hospitality while using the facilities during sampling trips. Thanks go out to Marcy,

Reece, and Yo for help with sampling. To Sue Welsch, thank you for running the nitrate samples.

I would like to thank my family and friends for supporting me over the years.

Special mention goes out to my: parents, brothers, grandparents, Uncle Pat, Uncle

Dan, and Tio Jerry. None of this would be possible with out all of you. I truly believe that.

Additionally, thank you to the endless amount of resources that kept me entertained throughout the generally long days in and outside of lab, including:

v KEXP, KCRW, NPR, The Wexner Center, Promowest Productions, and most notably, daily conversations with Marcy. These outlets helped keep me grounded over the years.

This research was funded by NOAA’s National Estuarine Research Reserve

System Graduate Research Fellowship # NA09NOS4200030.

Vita

2008…………………………………...... B.S. Environmental Science, The Ohio

State University

2008-2009…………………………………Graduate Research Associate, The Ohio

State University

2009-2010………………………………….Graduate Research Fellow, The Ohio

State University

Field of Study

Major Field: Environmental Science Graduate Program

vii

Table of Contents

Page

Abstract ...... ii

Dedication ...... iv

Acknowledgments ...... v

Vita ...... vii

List of Tables ...... x

List of Figures ...... xi

Chapter 1: Introduction ...... 1 1.1. Nature and Scope of Research...... 1 1.2. Photochemistry in Natural Waters...... 2 1.3. DOM Source Continuum...... 3 1.4. Organic Surface Microlayers...... 3 1.5. Research Objectives...... 4 1.6. Figures ...... 6 1.7. References ...... 9

Chapter 2: The Direct and Indirect Photofate of Two Polycyclic Musk Fragrances: HHCB and AHTN...... 12 2.1. Introduction ...... 12 2.2. Materials and Methods...... 15 2.2.1. Chemicals...... 15 2.2.2. Experimental Design...... 15 2.2.3. HPLC Analysis...... 18 2.3. Results and Discussion...... 19 2.3.1. HHCB ...... 19 2.3.2. AHTN ...... 23 2.4. Conclusion...... 26 2.5. Tables ...... 28 2.6. Figures...... 32 2.7. References ...... 38

viii Chapter 3: Photodegradation of Two Polycyclic Aromatic Hydrocarbons in Organic Surface Microlayers ...... 41 3.1. Introduction ...... 41 3.2. Materials and Methods...... 42 3.2.1 Field Site and Sampling Technique...... 42 3.2.2 Materials...... 43 3.2.3 Photolysis Experiments...... 43 3.2.4 HPLC Analysis...... 45 3.3. Results and Discussion...... 46 3.4. Conclusion...... 49 3.5. Tables ...... 50 3.6. Figures...... 51 3.7. References ...... 57

Chapter 4: Conclusions and Future Research ...... 59

Bibliography ...... 62

List of Tables

Page

Table 2.1. Physical Properties of HHCB and AHTN...... 28

-1 Table 2.2(a). Observed degradation rate constants (kobs (hrs )) and half-lives (hrs) for HHCB. Screening factors and the contribution of direct and indirect photolysis are also reported...... 29

-1 Table 2.2(b). Observed degradation rate constants (kobs (min )) and half- lives (min) for AHTN. Screening factors and the contribution of direct and indirect photolysis are also reported ...... 29

Table 2.3. Half-life of HHCB corresponding to hydroxyl radical concentrations...... 30

-1 Table 2.4(a). Observed degradation rate constants (kobs (hrs )) and half-lives (hrs) for HHCB in solutions containing quenchers ...... 31

-1 Table 2.4(b). Observed degradation rate constants (kobs (min )) and half- lives (min) for AHTN in solutions containing quenchers and sensitizers ...... 31

Table 3.1. Physiochemical properties of collected OSMs and Bulk Waters...... 50

List of Figures

Figure 1.1. Reaction scheme showing direct and indirect photochemical degradation pathways...... 6

Figure 1.2. Chemical structures of the chosen contaminants of study...... 7

Figure 1.3. Map of Old Woman Creek in Huron, Ohio ...... 8

-1 Figure 2.1(a). Observed degradation of HHCB (kobs (hrs )) in the presence of fulvic acids...... 32

-1 Figure 2.1(b). Observed degradation of HHCB (kobs (hrs )) in the presence of fulvic acids and quenchers, isopropanol (ISP) and t-butanol (t-but).... 32

Figure 2.2(a). OH competition experiment results using Fenton’s Reagent, acetophenone as a reference, and HHCB as a substrate ...... 33

Figure 2.2(b). OH competition experiment results using Fenton’s Reagent, acetophenone as a reference, and HHCB as a substrate ...... 33

-1 Figure 2.3(a). Observed degradation of AHTN (kobs (min )) in the presence of fulvic acids...... 34

-1 Figure 2.3(b). Observed degradation of AHTN (kobs (min )) in the presence of fulvic acids, rose bengal (RB) and heavy water (D2O 43%(v/v))...... 34

Figure 2.4. Proposed reaction scheme of HHCB with OH via: (i) electrophilic addition and (ii) hydrogen abstraction...... 35

1 Figure 2.5. Proposed reaction scheme of AHTN with O2 via Diels-Alder mechanism ...... 36

Figure 3.1. Schematic showing compositional make-up of OSMs...... 51

Figure 3.2. Aerial photograph of Old Woman Creek in Huron, Ohio ...... 52

-1 Figure 3.3. Observed degradation (kobs (hrs )) of phenanthrene in Milli-Q, bulk water, and OSM...... 53

-1 Figure 3.4. Observed degradation (kobs (hrs )) of naphthalene in Milli-Q, bulk water, and OSM...... 54

Figure 3.5. UV-Vis absorbance spectrum of Bulk Water and OSM collected on 07.07.09...... 55

Figure 3.6. UV-Vis absorbance spectrum of Bulk Water and OSM collected on 10.22.09...... 56

xii Chapter 1: Introduction

1.1. Nature and Scope of Research

Non-point source (NPS) pollution is a worldwide problem believed to cause serious impairment of freshwater resources (Kolpin et al., 2002; Sodre et al., 2010;

Hong et al., 2009). Amongst the compounds detected, the most concerning are labeled “emerging contaminants,” as little is known about their fate in aquatic ecosystems. This classification includes pharmaceutical drugs, antibiotics, pesticides, flame-retardants, personal care products (PCPs), and derivatives from oil and gas products. The occurrence of these compounds in surface waters cannot be attributed to one specific source. Instead, multiple contamination pathways have been proposed, including: incomplete removal from wastewater treatment plants (WWTPs), leaching landfills and septic systems, agricultural and urban run-off, and atmospheric deposition (Jacobs, 2008; Slack et al., 2005; Wania and Mackay, 1996). Regardless of the source, it is imperative to employ mechanisms that will eliminate these contaminants.

One passive, cost-effective, and possibly efficient remediation mechanism that has gaining heightened awareness is the application of natural and constructed wetlands. Several studies have documented the removal of emerging contaminants in wetlands (Chen, T, et al. 2006; Kao et al., 2002; Miller and Chin, 2005; Jacobs,

2008). One prominent removal mechanism is photochemistry. Wetlands are an ideal ecosystem for photochemistry because of their shallowness and long residence times.

1 Thus, for those chemicals that are susceptible to degradation by photolysis (either through direct or indirect pathways), wetlands may prove to be an efficient means to threat these compounds prior to discharge to downstream waters.

1.2. Photochemistry in Natural Waters

Direct photolysis is the process through which a given compound absorbs light and is subsequently transformed. Compounds that are structurally composed of chromophoric moieties (i.e. conjugated C=C, aromatics, ketones) and are capable of absorbing light at wavelengths present in sunlight, are susceptible to direct photolysis

(Schwarzenbach et al., 2003). Indirect photolysis can also transform compounds and is a complex process involving the excitation of photosensitizers, such as dissolved

- organic matter (DOM) or nitrate (NO3 ), resulting in the formation of reactive transients, which can transform contaminants (Schwarzenbach et al., 2003)(Figure

1.1).

DOM is an extremely important naturally occurring photosensitizer. It is a mixture of biogenically derived compounds found ubiquitously in aquatic environments. Several studies have examined the role of DOM as a sensitizer in the photochemical degradation of organic contaminants (Boreen et al., 2004; Packer et al., 2003; Guerard et al., 2009). DOM is known to photochemically produce a series

1 of reactive oxygen species (ROS) such as (H2O2) singlet oxygen ( O2), superoxide

- 3 (O 2), hydroxyl radical (OH), and triplet DOM ( DOM) that can facilitate the degradation of anthropogenically produced compounds (Blough and Zepp, 1995;

Schwarzenbach et al., 2003). Alternatively, DOM can “light-screen” other compounds by absorbing photons and can act as a natural scavenger of reactive

2 species (Torrents et al., 1997; Brezonik and Fulkerson-Brekken, 1998).

1.3. DOM Source Continuum

The source of DOM, either terrestrial or microbial, has been shown to play a pivotal role in its photochemical interactions with contaminants (Guerard et al.,

2009). The chemical and structural make-up of DOM is reflective of its parent materials. Terrestrial DOM is derived from higher order plants, whereas microbial

DOM is derived from phytoplankton and bacteria (McKnight et al., 1997; Guerard,

2009). Previous studies have experimentally used fulvic acid as a representative fraction of DOM. Multiple fulvic acids were used in this study in an effort to probe the influence of source composition on photoreactivity. Suwannee River Fulvic Acid

(SRFA) was chosen to represent a terrestrially derived end member. Conversely,

Pony Lake Fulvic Acid (PLFA) was chosen to represent a microbially derived end member. Old Woman Creek Fulvic Acid (OWCFA) lies in the middle of the source continuum and is representative of both terrestrially and microbially derived members. Using fulvic acids from both ends of the source continuum in photochemical reactions will offer insight into the fate of organic contaminants in differing ecosystems around the world.

1.4. Organic Surface Micolayers

Typically photochemical reactions in natural waters occur at depths in which light can penetrate through the surface, defined as the photic zone. Previous studies have focused on the photic zone but have overlook the air and water interface, defined as organic surface microlayers (OSMs), as a microreactor for photolytic degradation of organic contaminants. Organic surface microlayers are defined as the 1

3 to 1000 µm thick boundary separating the water column from the overlying air. The composition of this layer is comprised of naturally derived hydrophobic organic substances (Liss and Duce, 1997). As a result, nonpolar contaminants will accumulate in OSMs relative to the aqueous phase (Liss and Duce, 1997). For example, Cross and co-workers (2004) compiled concentration and enrichment factors for a number of contaminants including heavy metals, pesticides, polycyclic aromatic hydrocarbons, and persistent organic pollutants from OSMs sampled globally.

In addition to accumulating pollutants, photochemical reactions within OSMs are significantly different than processes in the water column because of unique solvent-light interactions (Feigenbrugel et al., 2005; Eneida et al., 2008). For example, solutes in organic solvents will absorb light at different wavelengths, whereby many compounds exhibit “red shifts” i.e., absorbance at longer wavelengths in organic solvents relative to water (Feigenbrugel et al., 2005; Eneida et al., 2008).

These changes in solute photochemical behavior could result in quicker and more efficient direct photodegradation within microlayers relative to underlying waters.

Finally, OSMs attenuate very little light because of their thickness and direct photodegradation of contaminants will be enhanced. Therefore, determining the rate and extent of photochemical transformations within microlayers will provide vital information to the fate of organic pollutants in wetlands.

1.5. Research Objectives

There are two main objectives of this research. I will first probe the photoreactivity of two commonly used polycyclic musk fragrances, HHCB and

AHTN (Figure 1.2), and elucidate the dominant degradation pathway in sunlit waters.

4 Only one study has considered the role of photochemical degradation as an effective transformation process in aquatic systems (Buerge et al., 2003). Their findings were limited, stating that AHTN is susceptible and HHCB recalcitrant to photochemical loss. The role of DOM in their study, however, was not considered. This study will examine the efficacy of DOM in the indirect degradation of AHTN and HHCB. I will also use a series of quenchers and sensitizers to determine the importance of indirect pathways in the photofate of these chosen fragrances. Additionally, multiple fulvic acids will be used experimentally to determine the sensitivity of these compounds towards the compositional make-up of DOM.

The second objective is to probe the efficacy of OSMs to act as photoreactors towards the direct photochemical degradation of polycyclic aromatic hydrocarbons

(PAHs), specifically phenanthrene and naphthalene (Figure 1.2). The photoreactivity of PAHs in natural waters have been extensively studied and are known to accumulate in OSMs, so they serve as ideal compounds for this study (Jacobs et al.,

2008; Liu and Dickhut, 1997). Organic surface microlayer samples will be collected at Old Woman Creek in Huron, Ohio (Figure 1.3). This will be the first study to examine the potential kinetic variation in PAH photodegradation between OSMs and underlying bulk water. Results of this study could lead to a more comprehensive understanding of the fate of PAHs in natural waters.

1.6. Figures:

Figure 1.1. Reaction scheme showing direct and indirect photochemical degradation pathways. (Courtesy Jennifer Guerard)

HHCB AHTN

Phenanthrene

Figure 1.2. Chemical structures of the chosen contaminants of study.

Figure 1.3. Map of Old Woman Creek in Huron, Ohio.

1.7. References

Boreen, A. L., Arnold, W. A., McNeill, K., (2004). Photochemical fate of sulfa drugs in the aquatic environment: Sulfa drugs containing five-membered heterocyclic groups. Environ. Sci. Technol. 38, 3933 – 3940.

Blough, N. V. and Zepp, R.G., (1995). “Reactive oxygen species in natural waters.” In: Active Oxygen: Reactive Oxygen Species in Chemistry, C. S. Foote, J. S. Valentine, A. Greenberg and J. F. Liebman. NY, Chapman & Hill: 280 – 332.

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.

Buerge, I.J.B., Buser, H.R., Mueller, M.D., Spoige, T., (2003). Behavior of the Polycyclic Musks HHCB and AHTN in Lakes, Two Potential Anthropogenic Markers for Domestic Wastewater in Surface Waters. Environ. Sci. Technol. 37, 5636-5644

Chen, T et al., (2006). Application of a constructed wetland for industrial wastewater treatment: A pilot-scale study. CHEMOSPHERE. 64: 497-502.

Cross, J.N., Hardy, J.T., Hose, J.E., Hershelman, G.P., Antrim, L.D., Gossett, R.W., Crecelius, E.A., Wurl, O., Obbard, J.P., (2004). A review of pollutants in the sea- surface microlayer (SML): a unique habitat for marine organisms. Mar. Pollut. Bull. 48: 1016–1030.

Eneida, R.P., Calvé, S.L., Mirabel, P., (2008). Near-UV molar absorptivities of alachlor, mecroprop-p, pendimethalin, propanil and trifluralin in methanol. J. Photochemistry Photobiology A: Chemistry. 193: 237-244.

Feigenbrugel, V., Loew, C., Le Calve, S., Mirabel, P., (2005). Near-UV molar absorptivities of acetone, alachlor, metolachlor, diazinon and dichlorvos in aqueous solution. J. Photochemistry and Photobiology A: Chemistry. 174: 76-81.

Guerard, J.L., Miller, P.L., Trouts, T.D., Chin, Y.P., (2009). The role of fulvic acid composition in the photosensitized degradation of aquatic contaminants. Aquat. Sci. 71: 160 – 169.

Guerard, J.L, (2009). The characterization of dissolved organic matter and its influence on the photochemical fate of antibiotics used in aquaculture. PhD Dissertation. The Ohio State University, Columbus OH, USA.

9 Hong, S, et al., (2009). PCDD/F, PBDE, and nonylphenol contamination in a semi- enclosed bay (Masan Bay, South Korea) and a Mediterranean lagoon (Thau, France).” CHEMOSPHERE. 77: 854-862.

Jacobs, L., Weavers, L., Chin, Y., (2008). Direct and indirect photolysis of polycyclic aromatic hydrocarbons in nitrate-rich surface waters. ENVIRONMENTAL TOXICOLOGY AND CHEMISTRY. 27: 1643-1648.

Jacobs, L.E., (2003). Photochemical transformation of three polycyclic aromatic hydrocarbons, ibuprofen, and caffeine in natural waters. PhD Dissertation. The Ohio State University, Columbus OH, USA.

Kao, C.M., Wang, J.Y., Chen, K.F., Lee, H.Y., Wu, M.J., (2002). Nonpoint source pesticide removal by a mountainous wetland. Water Sci.Technol. 46: 199–206.

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.

Liss, P.S. and Duce, R.S., (1997). The Sea Surface and Global Change, Cambridge Univ. Press, Cambridge, UK.

Liu, K., Dickhut, R., (1997). Surface microlayer enrichment of polycyclic aromatic hydrocarbons in Southern Chesapeake Bay. ENVIRONMENTAL SCIENCE & TECHNOLOGY. 31: 2777-2781.

McKnight, D. M., Harnish R., Wershaw R. L., Baron, J. S., Schiff, S., (1997). Chemical characteristics of particulate, colloidal and dissolved organic material in Loch Vale Watershed, Rocky Mountain National Park. Biogeochem. 36: 99-124.

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.

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

Schwarzenbach, R. P., Gschwend, P. M., and Imboden, D. M., (2003). Environmental Organic Chemistry. J. W. Wiley: New York, 1313.

Slack, R; Gronow, J; Voulvoulis, N., (2005). Household hazardous waste in municipal landfills: contaminants in leachate. SCIENCE OF THE TOTAL ENVIRONMENT. 337: 119-137.

Sodre, F; Locatelli, M; Jardim, W., (2010). Occurrence of Emerging Contaminants in Brazilian Drinking Waters: A Sewage-To-Tap Issue. WATER AIR AND SOIL POLLUTION. 206: 57-67.

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.

Wania F, Mackay D., (1996). Tracking the distribution of persistent organic pollutants. Environmental Science and Technology. 30: 390- 397.

11 Chapter 2: The Direct and Indirect Photofate of Two Polycyclic Musk Fragrances: HHCB and AHTN

2.1. Introduction

Synthetic polycyclic musk fragrances are emerging contaminants that have gaining heightened scrutiny over the past 15 years. Classified as pharmaceutical and personal care products (PPCPs), and as a high-production-volume chemical by the

EPA (EPA, 2003), polycyclic fragrances are used extensively in industrial and household cleaners, and most cosmetic products. Production volumes of polycyclic fragrances in 2000 were estimated at 6500 tons and 1800 tons in the United States and EU, respectively, and are expected to increase due to the growing demand and the lack of an economically feasible alternative (Somogyi and Kishi, 2001; Luckenbach and Epel, 2005; Peck et al., 2006). It has been estimated that 95% of the production volume contains 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-y-2- benzopyran (HHCB; Galaxolide) and 7-acetyl-1,1,3,4,4,6-hexamethyl-1,2,3,4- tetrahydronaphthalene (AHTN; Tonalide)(Balk and Ford, 1999) (Table 2.1). More importantly, both AHTN and HHCB have been shown to exhibit detrimental effects to aquatic biota at environmentally relevant concentrations (An et al., 2009;

Yamauchi et al., 2008). Therefore, understanding the fate of polycyclic fragrances in aquatic environments is imperative.

HHCB and AHTN are introduced into the environment through a combination

12 of direct (point source) and non-direct pathways. The primary direct source has been termed “down-the-drain,” and results from the use of consumer products containing polycyclic fragrances and subsequent disposal into waste drains (Sumner et al., 2010).

Indirect sources include wastewater treatment plant (WWTP) sludge that has been applied agriculturally or has been disposed of in a landfill (Horii et al., 2007; Slack et al., 2007). Previous studies have focused on direct sources, as it is relatively difficult to quantitatively assess the contribution of indirect sources.

Detection of HHCB and AHTN is worldwide in a variety of mediums: aquatic systems (Rimkus, 1999; Moldovan, 2006), wastewater treatment plant influent and effluent (Ricking, 2003; Simonich, 2002), and air (Peck et al., 2006; Xie et al., 2007).

Due to their hydrophobicity (log Kow= 5.7-5.9; Table 2.1), both polycyclic musks bioaccumulate and have been detected in aquatic species (Kannan et al., 2005;

O’Toole, 2006), human adipose tissue, and human milk (Mueller et al., 1996; Reiner et al., 2007). Low (µg/L) concentrations have been detected in WWTP effluents with dilution effects measured downstream (ng/L) (Sumner et al., 2010, Horii et al., 2007).

Common physiochemical processes in WWTPs to affect removal of polycyclic fragrances, i.e., absorption to sludge and biodegradation, have proven to be effective but are not comprehensive, with removal efficiency ranging between 72-

98% (Horii et al., 2007, Kupper et al., 2006). Additional biological sludge treatments, with goals of limiting the re-introduction of fragrances during agricultural application, have proven to be unsuccessful (Chen et al., 2009). Advanced oxidation processes have proven effective, specifically the use of UV/H2O2 tandem treatment

(Felis et al., 2008).

13 To date, very little is understood with regard to the fate of polycyclic fragrances in the natural environment. Only one study has considered the role of photochemical degradation as an effective remediation mechanism in aquatic systems (Buerge et al.,

2003). Their findings were limited, stating that AHTN is susceptible and HHCB recalcitrant to photochemical loss. The role of Dissolved Organic Matter (DOM) in this study, however, was not considered.

DOM is a mixture of biogenically derived compounds found ubiquitously in aquatic environments. Several studies have examined the role of DOM as a sensitizer in the photochemical degradation of organic contaminants (Boreen et al., 2004;

Packer et al., 2003; Guerard et al., 2009). DOM is known to photochemically produce a series of reactive oxygen species (ROS) such as hydrogen peroxide (H2O2), singlet

1 - oxygen ( O2), superoxide (O 2), and hydroxyl radical (OH), which can facilitate the degradation of anthropogenically produced compounds in a process known as indirect photochemical degradation (Blough and Zepp, 1995; Schwarzenbach et al., 2003).

Alternatively, DOM can “light-screen” other compounds by absorbing photons and can act as a natural scavenger of reactive species (Torrents et al., 1997; Brezonik and

Fulkerson-Brekken, 1998). Additionally, the source of DOM, either terrestrial or microbial, has been shown to play a pivotal role in its photochemical interactions with contaminants (Guerard et al., 2009).

The purpose of this study is to determine the role of DOM in the photofate of

HHCB and AHTN in freshwater ecosystems. The central hypothesis of this study is that DOM generated reactive transients will play a significant role in the degradation

14 of HHCB and AHTN. And finally, the conclusions will demonstrate that DOM composition will greatly affect the photochemical reactivity of HHCB and AHTN.

2.2 Materials and Methods

2.2.1. Chemicals

All solutions were prepared in 18 MΩ Milli-Q purified water (Millipore).

HPLC grade acetonitrile, hydrochloric acid (certified ACS), sodium hydroxide

(certified ACS), ferrous sulfate (FeSO47H2O), acetophenone (certified ACS) isopropanol, t-butanol, rose bengal, and deuterium oxide were all purchased from

Fisher Scientific. HHCB and AHTN were provided by the John D. Walsh Company.

Hydrogen peroxide (30%) was purchased from Mallinckrodt. All chemicals were used without additional purification. Suwannee River Fulvic Acid (SRFA), Suwannee River

Natural Organic Matter (SRNOM), and Pony Lake Fulvic Acid (PLFA) were purchased from the International Humic Substance Society (IHSS). Old Woman Creek

Fulvic Acid (OWCFA) was collected and isolated from Old Woman Creek (Huron,

Ohio) using XAD-8 methods described in Thurman and Malcolm (1981).

2.2.2. Experimental Design

HHCB and AHTN stock solution were prepared in acetonitrile. Working solutions were prepared in 120mL glass vials by pipetting an appropriate volume of stock followed by evaporation of acetonitrile and dilution in Milli-Q. In an effort to avoid loss of fragrance to volatilization, all working solutions were prepared with no headspace. All solutions were pH adjusted to 8.0 ± 0.1 with HCl and/or NaOH. Next, all solutions were transferred to Air-Tite Luer Lock 100mL syringes to further

15 avoid loss to volatilization. Equivalent weights of fulvic acids and organic matter were used for indirect photolysis experiments. Sorption of fragrances to fulvic acids and organic matter was not a significant loss process. Using known Kow values (Table

2.1) and relationships with fiw, the fraction in dissolved form, reported in

Schwarzenbach et al. (2003), it was estimated that greater than 99% of the initial concentration of fragrance remained in solution. Aliquots were taken for total organic carbon analysis (Shimadzu TOC-VCPN) and UV-Vis spectroscopy analysis (Varian

Cary 50). Isopropanol (25mM) and tert-butanol (25mM) were used as oxygen based radical scavengers (Packer et al., 2003; Buxton et al., 1988). Rose Bengal (40uM) and

1 D2O (43% (v/v)) were used as O2 sensitizers (Boreen et al., 2004). Anoxic experiments were prepared by sparging working solutions with argon (1min/mL) and transferring to a glovebox (95%N2/5%H2) for further preparation.

Quartz tubes (0.9 cm path length capped with Teflon® lined quartz caps) were filled with working solutions, screening wavelengths <290nm. Photolysis experiments were 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 equivalent to approximately two half-lives. Dark controls were wrapped in foil, run concurrently, and showed no degradation throughout all experiments. Temperature and radiometer readings of Suntest conditions were monitored and remained constant throughout all experiments.

16 of the light source throughout the course of our actinometry experiments. Using the data collected in the actinometry experiments, it was estimated by Guerard and co- workers (2009) that the light source is approximately 4.5 times more intense than average sunlight conditions measured at 40°N at noon in June. Therefore, conclusions from the observed degradation of HHCB and AHTN, have considered the measured intensity of the light source relative to natural sunlight.

All fulvic acid solutions were corrected for light-screening effects using the procedure described in (Miller and Chin, 2002) and equation (1) described below.

(1)

Sλ is the wavelength specific screening factor, aλ is the experimentally determined absorbance at a specific wavelength, and l is the path length for the photolysis tubes used (0.9cm for quartz tubes). Wavelength specific screening factors, Sλ, were determined from 290nm – 750nm and then plotted vs. wavelength to determine a 4th order polynomial least squares regression for each solution (R2’s = .99). The regression line was integrated over the area in which the absorbance of HHCB and

AHTN overlap with the absorbance of the fulvic acid solutions (290nm – 370nm for both compounds). These values were then normalized to the area of a solution in which no screening was anticipated (i.e., Sλ = 1) to yield an overall screen factor,

(SΣλ).

Using the calculated SΣλ, overall reaction rates, kobs, can be interpreted in

17 terms of their contribution of direct and indirect photolysis, kdp and kip, respectfully.

Direct photolysis rates for solutions containing fulvic acids, kdp, were calculated by multiplying the determined overall screening factor by kmilliq, where kmilliq is the direct photolysis observed rate experimentally determined in Milli-Q. Assuming an independent reaction with direct pathways, indirect photolysis rates for solutions containing fulvic acids, kip, were calculated by subtracting kdp from kobs. The contribution towards direct and indirect pathways was calculated with the equations:

%DP = (kdp / kobs)*100 and % IP = (kip / kobs)*100.

Hydroxyl radical competition experiments were set-up, in a dark room, using

Fenton’s Reagent, similar to as described in Haag and Yao (1992) and Packer et al.

(2003). A 100mL solution (pH 3.5) containing 100µM acetophenone, 100nM HHCB, and 0.2mM ferrous sulfate was spiked with 5mM H2O2. The reaction took place in an

Air-Tite Luer Lock syringe. Following the addition of the H2O2, samples were taken consecutively and immediately added to HPLC vials pre-loaded with an equal amount of methanol, to quench the reaction. The equal amounts (v/v) of reaction solution and methanol totaled the capacity of the HPLC vial, 5.2 mL, allowing for no headspace. The experiment was performed twice to account for potential variability of the experimental design.

2.2.3. HPLC Analysis

The concentrations of HHCB and AHTN were determined using high pressure liquid chromatography (HPLC). Analytes were separated with a Restek C-18 Reverse

Phase Column (5um x 100mm x 2.1 mm). A 60% acetonitrile (ACN) and 40% water

(v/v) mobile phase at a flow rate of 0.5ml/min was used for all experiments except for

18 competition kinetic experiments. A gradient method was used for competition experiments: 1-7 min 80/20 (v/v) H2O/ACN; 9-23 min 30/70 (v/v) H2O/ACN; 25-30 min 80/20 (v/v) H2O/ACN. 100µL samples of HHCB were injected and quantified using fluorescence spectroscopy detection at λex = 275nm / λem = 295nm (Waters

Corporation 2475, Breeze 3.3 software). 75µL samples of AHTN were injected and quantified using UV-Vis spectroscopy detection at λ = 257 nm (Waters Corporation

2487, Breeze 3.3 software). 100uL samples of acetophenone were injected and quantified using UV-Vis spectroscopy detection at λ = 255 nm. All HPLC sample vials were filled with no headspace in an effort to avoid loss to volatilization. Because of inconsistency of multiple injects, due to rapid volatilization of compound, replicate photolysis tubes were prepared for select time points.

2.3 Results and Discussion

2.3.1. HHCB

HHCB (Co = 45 - 65 nM) degrades by direct photolysis, with a pseudo first

-1 order rate constant of .0267 ± .00491 Hrs (t1/2 = 26.0 Hours)(Table 2.2(a)). These results are faster than previously reported, (Buerge et al., 2003 (t1/2 = 138.6 Hours, Co

= 1ug/L :: 3.87nM)) who ran his experiments under a low-pressure mercury-vapor fluorescent lamp, but are comparable when considering the differing intensities of the light sources used in these studies. In natural sunlight, June at 40°N and assuming 12 hrs sunlight/day, it can be estimated that the environmental half-life of HHCB is 9.75 days (Table 2.3).

The addition of fulvic acids to working solutions (8.22 ± 0.01 mg-C/L SRFA,

7.25 ± 0.01 mg-C/L PLFA, 7.66 ± 0.14 mg-C/L OWCFA, and 6.00 ± 0.02 mg-C/L

19 SRNOM) resulted in changes in the photoreactivity, inferring a dependence on DOM composition (Figure 2.1(a)). When adding SRFA, PLFA and OWCFA, the rate constant increased by a factor of 1.81, 1.29, and 1.62, respectively. Solutions containing SRFA, SRNOM, and OWCFA displayed significant contribution, >40%, from indirect pathways with SRFA having the most significant contribution at 52%

(Table 2.2(a)). The rates remained statistically equivalent when comparing the enhancements of SRFA vs. SRNOM solutions, confirming reports that fulvic acid is the primary reactive portion of DOM (McKnight et al., 2001). In an effort to isolate which indirect pathway was responsible for enhanced degradation, a series of quenchers and sensitizers were added to the working solutions.

Photolysis experiments were conducted in the presence of isopropanol, known

9 -1 -1 to scavenge oxygen-based radicals (e.g., OH, ROO, RO (kOH = 4.3 x 10 M s ;

Buxton et al., 1988; Lloyd et al., 1976), and t-butanol, known to scavenge only

8 -1 -1 hydroxyl radical (kOH = 6.0 x 10 M s ; Buxton et al., 1988). In the presence of isopropanol, complete quenching was observed in SRFA solutions (kip/kmq = 0.99)

(Table 2.4(a), Figure 2.1(b)). In the presence of t-butanol, significant but incomplete quenching was observed (ktb/kmq = 1.22) (Table 2.4(a), Figure 2.1(b)). Incomplete quenching with the addition of t-butanol can be explained by its slower reaction rate with OH compared to that of isopropanol. Isopropanol is reported to react with OH

~7x faster than t-butanol. These results suggest that a reaction with OH is an important pathway responsible for enhanced indirect HHCB degradation.

In order to elucidate the relative importance of OH, competition kinetics were performed to determine the 2nd order reaction rate of HHCB with OH. Using

20 the Fenton pathway and a competitive probe molecule, acetophenone, the data was fitted to equation 2,

(2)

where S is the substrate HHCB, R is the reference compound acetophenone, and kR OH is the known 2nd order rate constant of acetophenone with OH (5.9 x 109

M-1s-1; Buxton et al., 1988). Plotting pseudo first order degradation kinetics of HHCB and acetophenone yields the line ln (S/So) vs. ln (R/Ro) with an averaged slope of

0.7995 ± 0.0091 (R2’s > .99, n=2)(Figure 2.3). Multiplying the slope by the known

9 -1 -1 kOH for acetophenone yields 4.717 ± .05 x 10 M s , kOH for HHCB. The experimentally determined kOH for HHCB is most likely undervalued as the initial concentrations of acetophenone and HHCB (100 µM and 100nM, respectively) were not equivalent because the limits of detection for each analyte were different.

Therefore, the actual in situ reaction rate is most likely slightly higher than reported in this study.

The observed dependence on DOM composition corroborates previous studies from our group. Fulvic acids from allocthonous sources (SRFA) produce OH three times faster than autochthonous sources (Lake Fryxell Fulvic Acid, Antarctica), 6.0 x

10-11 and 2.3 x 10-11 M/s per mg-C/L, respectively (White, 2000). Environmental

-18 -14 hydroxyl radical concentrations are typically low and range from 10 M to 10 M,

- depending on the photosensitizers present at the sampling site, i.e. NO3 and DOM

(Mabury and Crosby, 1994; Lam et al., 2003). Our group experimentally determined

21 steady state hydroxyl radical concentrations, [OH]ss, for Old Woman Creek to range

-16 -16 from 2.1 x 10 M to 5.9 x 10 M for solutions containing 4 and 6 mg/L C, respectively (Houtz, 2008). Table 2.3 shows the calculated half-life for HHCB at environmentally relevant [OH]ss. Clearly the environmental photofate of HHCB is highly dependent on [OH]ss. The experimentally determined [OH]ss were derived under constant irradiance and do not consider production limitations such as diurnal effects and cloud cover. Regardless, it is evident that the reaction of HHCB with

[OH] is a pertinent and environmentally relevant photochemical loss mechanism that has not been examined in previous studies.

Hydroxyl radical is a highly reactive non-specific radical found at low concentrations in the environment. It reacts readily with many organic pollutants at nearly diffusion controlled rates, ~ 1010 M-1s-1 (Haag and Yao, 1992). Two primary reaction pathways for OH with organic contaminants include: (i) electrophilic addition to a double bond or an aromatic system, and (ii) hydrogen abstraction from a carbon atom (Schwarzenbach et al., 2003; Mazellier et al., 2007; Haag and Yao,

1992). Electrophilic addition will occur in systems containing unsaturated electron rich double bonds. When considering the structure of HHCB, two unsaturated carbon atoms in the aromatic ring, positions 1 and 4, offer a potential site for the electrophilic addition of OH. The carbon atom in position 4 serves as the most likely candidate for electrophilic addition due to its proximity to and subsequent radical stability provided by the highly electron donating ether oxygen on the alicyclic ring (Figure

2.4). The potential for hydroxylation to occur via hydrogen abstraction is strictly dependent on the susceptibility of hydrogen atoms to leave and the stability of the

22 newly formed radical. Accordingly, there are three sites that are preferential for hydrogen abstraction: (i) the tertiary carbon on the 6-membered alicyclic ring, (ii) the tertiary carbon on the 5-membered alicyclic ring, and (iii) the secondary carbon adjacent to the ether oxygen and in conjunction with the aromatic ring (Figure 2.44).

The susceptibility for the hydrogen to leave is likely greatest in the 5-membered ring due to its close proximity to multiple (5) electron donating methyl groups. Yet, the most stable radical would be formed from the abstraction of either of the 6-membered hydrogens, as resonance structures with the neighboring aromatic ring would allow for delocalization of the electron charge. Therefore, hydroxylation via hydrogen abstraction can occur at multiple sites, but would favor the abstraction of the hydrogen atoms on the tertiary carbon and secondary carbon of the 6-membered alicyclic ring, in conjunction with the aromatic ring.

2.3.2. AHTN

AHTN, Co = 5± 0.3µM, degrades by direct photolysis, with a pseudo first

-1 order rate constant of .009 ± 2e-4 min (t1/2 = 77 mins)(Table 2.2(b); Figure 2.2(a)).

These results are faster than reported in previous studies, (Buerge et al., 2003 (t1/2 =

227 minutes, Co = 1µg/L or 3.87nM)) who ran his experiments under a low-pressure mercury-vapor fluorescent lamp, but are comparable when considering the differing intensities of the light sources used in these studies. In natural sunlight, June at 40°N and assuming 12 Hrs sunlight/day, it can be estimated that the environmental half-life of AHTN is 693 minutes.

The addition of fulvic acids to working solutions (8.22 ± 0.01 mg-C/L SRFA,

7.25 ± 0.01 mg-C/L PLFA, 7.66 ± .14 mg-C/L OWCFA, 6.00 ± .02 mg-C/L

23 SRNOM) increased rate constants by a factor of ~1.5 for SRFA and PLFA, and ~2.4 for OWCFA, inferring a dependence on DOM composition (Figure 2.2(a)). All solutions had significant contribution (> 40%) from indirect pathways with OWCFA having the most pronounced contribution (63%)(Table 2.2(b)). Reaction rates remained statistically equivalent when comparing the effects of SRFA vs. SRNOM in solution, which again confirms the role of the fulvic acid fraction in the photoreactivity of DOM. It can be inferred that the acceleration of degradation results from the development and subsequent reaction of DOM generated species with

AHTN. In an effort to isolate which reactive species was responsible for degradation, a series of quenchers and sensitizers were added to the working solutions.

In the presence of isopropanol some quenching of the reaction occurs for

SRFA solutions (kip/kmq = 1.22), but not for PLFA solutions (kip/kmq = 1.55) (Table

2.4(b)). This observation suggests that the importance of oxygen containing radicals is slightly dependent upon DOM composition but overall plays an insignificant role towards AHTN degradation. To test the sensitivity of AHTN toward photo-excited

DOM (3DOM) as a potential oxidant, experiments were performed anoxically in the presence of SRFA and PLFA. If the reaction of AHTN with 3DOM were significant, it would be expected that reaction rates of these solutions would increase with the removal of oxygen, a 3DOM scavenger (Guerard et al. 2009). Results in Table 2.4(b) show that AHTN degradation actually decreased when oxygen was removed which suggests that the formation of photo-stimulated 3DOM is not an important reactive transient. Additionally, it has been shown that indirect degradation pathways involving 3DOM is highly dependent upon DOM composition (Cawley et al., 2009).

24 This was not the case for AHTN as all sources of DOM were highly reactive, providing further evidence that 3DOM is not an important pathway. The observed decrease in kinetics with the absence of oxygen suggests the involvement of an oxygen dependent reactive transient, other than 3DOM.

1 To probe the potential reactivity of singlet oxygen ( O2), rose bengal, a known

1 O2 photo-sensitizer, was added to working solutions. Reaction rates in all rose bengal solutions, containing SRFA and PLFA, showed enhanced rates by a factor of ~2.5 compared to that of direct photolysis rates (Table 2.4(b); Figure 2.2(b)). Rose bengal experiments were also performed in the absence of fulvic acids to ensure that no unwarranted reactions were occurring between rose bengal and the fulvic acids. These experiments showed slightly higher degradation rates, inferring that the fulvic acids were screening the light absorption of either rose bengal or AHTN. Next, AHTN working solutions were prepared in H2O and D2O (43% (v/v)), again in the presence

1 of rose bengal. H2O is a known O2 quencher and it has been calculated that the

1 lifetime of O2 is 14 times greater in D2O than H2O (Wilkinson et al., 1995). The reaction rate increased significantly in D2O (kD2O/kH2O = 1.66)(Table 2.4(b); Figure

1 2.2(b)), again showing the importance of O2 as a reactive transient in the degradation of AHTN.

Singlet oxygen is a highly reactive species formed through the reaction of dissolved oxygen with the excited triplet states of DOM (Latch et al., 2006).

Moreover, there is significant evidence demonstrating the formation of intra-DOM

1 microenvironments with elevated O2 levels (Grandbois et al., 2008). Therefore,

1 photochemically produced O2 from DOM is expected to be a prominent loss process

25 for deleterious contaminants. Three primary reaction pathways for organic

1 contaminants susceptible to reaction with O2 include compounds containing: (i) moieties that are susceptible to Diels-Alder reaction, (ii) electron donating substituted double bonds, and (iii) easily oxidized functional groups (e.g., aninlines and phenols)(Schwarzenbach et al., 2003). When considering the structure of AHTN, it

1 becomes evident that the only feasible reaction pathway is via Diels-Alder, with O2 serving as the dienophile and the aromatic ring as the diene (Figure 2.5). This reaction is favorable because of the highly electron withdrawing ketone functional group that is in conjugation with an aromatic double bond. Moreover, the distribution of Diels-

Alder product isomers is complex and will be a function of electronic and steric properties. Therefore, Diels-Alder pathways will be the dominant transformation

1 mechanism for the reaction of AHTN with O2.

2.4. Conclusion

This study provides a comprehensive look at the photofate of polycyclic musk fragrances HHCB and AHTN in freshwater ecosystems. Results show that both

HHCB and AHTN react through direct photolytic pathways, with HHCB being significantly more refractory. Indirect photolysis proved to be important for both compounds. Competition experiments show that the fate of HHCB is highly dependent on environmental [OH]ss, as reaction with OH is an efficient degradation mechanism. Although competition experiments were not performed for AHTN,

1 results of multiple quencher and sensitizer experiments show O2 to be an important

DOM generated reactive transient. The predicted influence of DOM composition on reactivity was seen for both HHCB and AHTN. Photoproducts were not examined but

26 potential reaction pathways were proposed. Future research should include isolating

1 photoproducts as well as determining a second order reaction rate of AHTN with O2.

Finally, this is the first study to identify that DOM generated reactive transients play a significant role in the photo-fate of HHCB and AHTN in natural aquatic systems. The conclusions of this study did not account for effects retarding photodegradation in natural systems such as turbidity, water depth (represented as ~1cm), cloud cover, and diurnal cycles. Therefore, these results correspond to conditions considered to be optimal for photochemical degradation.

2.5. Tables

Table 2.1. Physical Properties of HHCB and AHTN. (From Artola-Garicano et. al., 2003)

Sample DOC SΣ290- kobs *100 Half-Life kdp *100 kip %DP %IP -1 -1 (mg C/L) 370nm (hrs ) (hrs) (hrs ) *100 (hrs-1) HHCB NA NA 2.67 26.0 2.67 NA 100 0.0 SRFA 8.2 0.86 4.82 14.4 2.30 2.52 47.6 52.3 PLFA 7.3 0.91 3.44 20.2 2.44 1.00 70.9 29.1 SRNOM 6.0 0.90 4.44 15.6 2.41 2.03 54.2 45.8 OWCFA 7.7 0.91 4.32 16.1 2.44 1.88 56.5 43.5

-1 Table 2.2(a). Observed degradation rate constants (kobs (hrs )) and half-lives (hrs) for HHCB. Screening factors and the contribution of direct and indirect photolysis are also reported.

Sample DOC SΣ290- kobs *100 Half-Life kdp *100 kip %DP %IP -1 -1 (mg C/L) 370nm (min ) (min) (min ) *100 (min-1) AHTN NA NA 0.90 77.0 0.90 NA 100 0.0 SRFA 8.2 0.86 1.30 53.3 0.77 0.53 60.0 40.4 PLFA 7.3 0.91 1.40 50.0 0.82 0.58 58.9 41.3 SRNOM 6.0 0.90 1.50 46.2 0.81 0.69 54.1 45.9 OWCFA 7.7 0.91 2.20 32.0 0.82 1.38 37.4 62.6

-1 Table 2.2(b). Observed degradation rate constants (kobs (min )) and half-lives (min) for AHTN. Screening factors and the contribution of direct and indirect photolysis are also reported.

Half-Life of HHCB Corresponding to Hydroxyl Radical Concentrations (hrs-1) -1 -1 9 1 2 Compound kOH (M s ) x 10 Hydroxyl Radical Concentrations Environmental 10-17M 5.9*10-16M 10-15M HHCB 4.717 4082 69 41 234

Table 2.3. Half-life of HHCB corresponding to hydroxyl radical concentrations 1Value corresponds to concentrations reported for Old Woman Creek, Ohio (Houtz, 2008). 2Environmental half-life reflects diurnal effects and light source intensity effects.

HHCB Indirect Photolysis Quencher Experiments -1 Sample kobs (hrs ) Half-life (hrs) 25mM Isopropanol SRFA 0.0263 ± 3.6e-3 26.35 25mM Tert-butanol SRFA 0.0325 ± 5.2 e-3 21.33

-1 Table 2.4(a). Observed degradation rate constants (kobs (hrs )) and half-lives (hrs) for HHCB in solutions containing quenchers.

AHTN Indirect Photolysis Quencher and Sensitizer Experiments -1 Sample kobs (min ) Half-life (min) 25mM Isopropanol Addition SRFA 0.011 ± 4e-4 63.0 PLFA 0.014 ± 6e-4 49.5 Anoxic Photolysis SRFA 0.009 ± 12e-4 77.0 PLFA 0.006 ± 6e-4 115.5 40uM Rose Bengal Addition SRFA 0.021 ± 2e-3 33.0 PLFA 0.024 ± 2e-3 28.9 MQ 0.0279 ± 2e-3 24.8 D2O 43% (v/v) 0.0464 ± 4e-3 14.9

-1 Table 2.4(b). Observed degradation rate constants (kobs (min )) and half-lives (min) for AHTN in solutions containing quenchers and sensitizers.

31 2.6. Figures

-1 Figure 2.1(a). Observed degradation of HHCB (kobs (hrs )) in the presence of fulvic acids. Errors represent 95% confidence interval.

-1 Figure 2.1(b). Observed degradation of HHCB (kobs (hrs )) in the presence of fulvic acids and quenchers, isopropanol (ISP) and t-butanol (t-but). Errors represent 95% confidence interval.

Figure 2.2(a). OH competition experiment results using Fenton’s Reagent, acetophenone as reference compound, and HHCB as a substrate.

Figure 2.2(b). OH competition experiment results using Fenton’s Reagent, acetophenone as reference compound, and HHCB as a substrate.

-1 Figure 2.3(a). Observed degradation of AHTN (kobs (min )) in the presence of fulvic acids. Errors represent 95% confidence interval.

-1 Figure 2.3(b). Observed degradation of AHTN (kobs (min )) in the presence of fulvic acids, rose bengal (RB) and heavy water (D2O 43%(v/v)). Errors represent 95% confidence interval.

Figure 2.4. Proposed reaction scheme of HHCB with OH via: (i) electrophilic addition and (ii) hydrogen abstraction.

1 Figure 2.5. Proposed reaction scheme of AHTN with O2 via Diels-Alder mechanism.

36 2.7 References

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40 Chapter 3: Photodegradation of Two Polycyclic Aromatic Hydrocarbons in Organic Surface Microlayers

3.1. Introduction

Organic surface microlayers (OSMs) are defined as the 1 to 1000 µm thick boundary separating the water column from the overlying air. The composition of this layer is comprised of naturally derived hydrophobic organic substances (Liss and

Duce, 1997)(Figure 3.1). As a result, nonpolar contaminants will accumulate in

OSMs relative to the aqueous phase (Liss and Duce, 1997). For example, Cross and co-workers (2004) compiled concentration and enrichment factors for a number of contaminants including heavy metals, pesticides, polycyclic aromatic hydrocarbons

(PAHs), and persistent organic pollutants (POPs) from OSMs sampled globally. In all cases these compounds accumulated in the OSMs relative to the underlying water.

In addition to accumulating pollutants, photochemical reactions within OSMs are significantly different than processes in the water column because of unique solvent- light interactions (Feigenbrugel et al., 2005; Eneida et al., 2008). For example, solutes in organic solvents will absorb light at different wavelengths, whereby many compounds exhibit “red shifts” i.e., absorbance at longer wavelengths in organic solvents relative to water (Feigenbrugel et al., 2005; Eneida et al., 2008). These changes in solute photochemical behavior could result in quicker and more efficient direct photodegradation within microlayers relative to underlying waters. Finally,

OSMs attenuate very little light because of their thickness and direct

41 photodegradation of contaminants will be enhanced. Indeed, the chemical properties of microlayers are so unique that the term “microreactor” has long been associated with them (Liss and Duce, 1997). Therefore, determining the rate and extent of photochemical transformations within microlayers will provide vital information to the fate of organic pollutants in wetlands such as OWC.

Polycyclic aromatic hydrocarbons (PAHs) are released into the environment primarily through the combustion of fossil fuels, but other sources include forest fires and the use of creosotes as wood preservatives (Schwarzenbach et al., 2003). PAHs were chosen as analytes for this study based upon four criteria: (i) their ubiquity in the environment, especially in the Lake Erie region (Sun et al., 2006; Kelly et al.,

1991) (ii) their known accumulation in OSMs (Liu and Dickhut, 1997; Lim et al.,

2007) (iii) their well-documented susceptibility to photodegradation (Jacobs et al.,

2008; Fasnacht and Blough, 2002; Chen et al., 2000), and (iv) their known deleterious effect on aquatic ecosystems and human health (Mearns et al., 2009). Based upon these criteria, both phenanthrene and naphthalene, two well-studied PAHs, are ideal candidates for probing the efficacy of OSMs as efficient photoreactors.

To date, no research has explored the photochemical transformation pathways of PAHs within OSMs. Thus, the central hypothesis of this study is that the unique chemical properties of OSMs coupled with the propensity for PAHs to accumulate in this layer will result in more efficient direct photodegradation, relative to processes in the underlying water column.

3.2 Materials and Methods

3.2.1 Field Site and Sampling Technique

42 Old Woman Creek National Estuarine Research Reserve (OWC) (Figure 3.2) is a freshwater wetland adjacent to Lake Erie in Huron, Ohio. Characteristics such as long residence times, littoral plant communities, high pelagic zone primary productivity, reduced winds, and subsequent lack of waves (Dr. David Klarer personal communication) make OWC an ideal environment for the establishment of a well-defined organic surface microlayer (Liss and Duce, 1997).

Multiple devices have been developed for microlayer sampling. Suitable samplers must selectively extract samples from only the uppermost portion of the air- water interface, avoiding significant dilution by the water column below the OSM.

Although high dilution factors were measured by Gever and co-workers (1996) with the glass plate technique (Harvey and Burzell, 1972), this still proved to be the most appropriate apparatus to use for sampling at OWC due to constricted boat space and limited mobility on the water.

Sampling occurred at two unique times to determine the effects of ecological changes at OWC on OSM composition and reactivity. The first OSM sample was retrieved on July 7, 2009, coinciding with peak primary production (OWC annual report, 2003). The second was retrieved on October 22, 2009, capturing the time when microbial degradation of littoral plants and algal biomass was most active

(OWC annual report, 2003). Plant exudates and the degradation of algal biomass is the source of OSMs in surface waters, therefore the thickness of these microlayers was theoretically maximized with this sampling schedule (Liss and Duce, 1997).

3.2.2 Materials

All solutions were prepared in 18 MΩ Milli-Q (MQ) purified water

43 (Millipore). HPLC grade acetonitrile, hydrochloric acid (certified ACS), and sodium hydroxide (certified ACS) were purchased from Fisher Scientific. Naphthalene (Sigma

99%) and phenanthrene (Eastman Kodak fluorescent grade) were chosen as analytes for this study.

Glass plates, 12” x 16”, were purchased from Ace Hardware (Columbus,

Ohio), ashed at 450°C for 4 hours and then wrapped in aluminum foil for transport to

OWC. Squeegees were wrapped in Teflon tape in order to avoid contamination from squeegee to samples. Collection of OSMs was performed in accordance with the method described by Harvey and Burzell (1972). Plates were dipped into the estuary at a rate of 20cm/sec, drained for ½ minute, and then squeegeed into a stainless steel funnel connected to a 500mL ashed glass jar. Bulk water samples were taken simultaneously in ashed glass jars at a depth >10cm. The samples were immediately transported to Columbus, filtered to .45um (47mm polycarbonate membrane filters purchased from Poretecs Corporation), and then refrigerated in the dark until further use. Aliquots were taken for pH analysis, total organic carbon analysis

(Shimadzu TOC-VCPN), UV-Vis spectroscopy analysis (Varian Cary 50), and Nitrate

- (NO3 ; Courtesy of Dr. Anne Carey). Blanks were taken on all sampling equipment and were negligible (<0.4 mg-C/L). Additionally, PAHs were not detected in samples retrieved from OWC.

3.2.3 Photolysis Experiments

Phrenanthrene (PN) and Naphthalene (NAP) stock solution were prepared in acetonitrile. Working solutions were prepared in 100mL volumetric flasks at desired concentrations, 25 nM PN and 155nM NAP, by pipetting an appropriate volume of

44 stock followed by evaporation of acetonitrile and dilution with either, MQ, bulk water or OSM. In an effort to avoid loss of PAHs to volatilization, all working solutions were prepared with no headspace. All solutions were pH adjusted to 8.0 ± .1 with HCl and/or NaOH.

Quartz tubes (0.9cm path length capped with Teflon lined quartz caps) were filled with working solutions, screening wavelengths <290nm. Photolysis 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 ~ two half-lives.

Dark controls were wrapped in foil, run concurrently, and showed no degradation throughout all experiments. Temperature and radiometer readings of Suntest conditions were monitored and remained constant throughout all experiments.

Actinometry, a chemical light meter, was performed to measure the photon flux of our light source. We used the p-nitroanisole (PNA)/pyridine system as described in Dulin and Mill, 1982. No significant changes were seen in the intensity of our light source throughout the course of our actinometry experiments. Using the data collected in our actinometry experiments, it was estimated by Guerard and co- workers (2009) that the solar simulator used is approximately 4.5 times more intense than average sunlight conditions measured at 40°N at noon in June. Therefore, conclusions from observed degradation of HHCB and AHTN should consider the intensity of the light source.

3.2.4 HPLC Analysis

Both PN and NAP were analyzed using high pressure liquid chromatography

(HPLC) (Waters Corporation 2475, Breeze 3.3 software) with fluorescence

45 spectroscopy detection at λex = 252 nm/λem = 352 nm for PN, and λex = 276 nm/λem =

332 nm for NAP. 100µL of solutions were injected directly and separated with a

Restek C-18 Reverse Phase Column (5um x 100mm x 2.1mm). A 60% methanol and

40% water mobile phase (v/v) at a flow rate of .5ml/min was used for all experiments.

All HPLC sample vials were filled with no headspace in an effort to avoid loss to volatilization. Because of inconsistency of multiple injects, due to rapid volatilization of compound, replicate photolysis tubes were prepared for select time points.

3.3. Results and Discussion

Photolysis experiments were ran simultaneously in MQ, bulk water, and

OSMs, to compare the photoreactivity of the selected solutes in the different solvents.

Samples collected on 7.7.10 were used to test the photoreactivity of PN while samples collected on 10.22.10 were utilized to probe the photoreactivity of NAP. In all experiments, the observed photodegradation of PN and NAP was significantly faster when using bulk water and OSMs as solvents compared to that observed using

Milli-Q as a solvent (Figure 3.3 & 3.4). These results were expected, as surface

- waters contain photosensitizers, i.e. NO3 , DOM, that are known to enhance degradation of PN and NAP (Jacobs et al., 2008, Fasnacht and Blough, 2002).

Additionally, rates observed for PN and NAP degradation were similar to those reported previously by our group using the same experimental set-up (Jacobs et al.,

2008). Conversely, the observed photodegradation of PN and NAP in bulk water and

OSM samples were statistically equivalent in all experiments (Figure 3.3 & 3.4).

Slight rate increases were observed in OSM solutions compared to bulk water, but when considering the calculated error, the increases were not statistically significant.

46 Additional analytical chemical parameters measured in this study included:

- pH, nitrate (NO3 ), dissolved organic carbon (DOC) (mg-C/L), specific UV absorbance (SUVA, which is essentially an extinction coefficient) at 280nm (L/mol-

- C), and UV-Vis spectroscopy. Results for pH, NO3 , DOC, and SUVA can be found in Table 3.1. Measured pH values were consistent between sampling dates. Nitrate levels were consistent between samples for the July event but showed an ~1.5 factor

- increase in OSMs for the October event. The variance in NO3 values between sampling events can likely be explained by precipitation that occurred in the days prior to sampling. Old Woman Creek is situated in an agricultural watershed known

- for its application of NO3 containing fertilizers. In both sampling events, measured

DOC values were higher in OSM than bulk water samples. These results corroborate findings in previous studies (Liu and Dickhut, 1998). Additionally, higher DOC values were detected in the July samples. Figures 3.5 and 3.6 show UV-Vis spectrums for their respected sampling date. The spectrums show that collected bulk water and

OSM samples were optically similar for individual sampling events. Significant differences were seen between sampling dates, with July samples showing higher absorbance. SUVA280 has been used as a simple proxy of DOM aromaticity. The larger SUVA values calculated for bulk water samples suggest that there was higher aromatic content in the bulk water DOM compared to the OSMs. These results were expected as OSMs accumulate organic molecules that do not absorb radiation in the

UV-Vis spectra (e.g., fatty acids, hydrocarbons) (Liss and Duce, 1997). Most importantly, the variation in calculated SUVA280 indicates that the collected bulk water and OSM samples were chemically unique. Yet, as previously reported,

47 fluctuations in the chemical photoreactivity of PN and NAP in these unique solvents were not observed.

Two theories have been proposed explaining statistically similar solute photoreactivity when dissolved in the unique solvents: (i) dilution effects and (ii) path length effects. Previous studies have measured high sample dilution of OSMs when using the glass plate method (Gever et al., 1999). OSMs are largely composed of organic molecules that absorb radiation at a much broader wavelength range, resulting in unique light to properties not observed in the underlying bulk water. The accumulation of hydrophobic organic molecules in the collected OSMs was not significant enough to exhibit this unique photochemical property of OSMs. Moreover, a more efficient sampling mechanism must be employed that can collect undiluted microlayer samples, without compromising the convenience of the glass plate method.

While dilution effects likely contributed to the inconclusive results of this study, the experimental design of this study contains geometric flaws that need to be addressed for future studies. The path length of quartz photolysis tubes is 0.9 cm.

Organic surface microlayers have reported thicknesses of 1 – 1000µm. Another unique property of OSMs is that they attenuate very little light because of their thickness. Using quartz photolysis tubes eliminates the effects of this unique physical property of OSMs. In order to accurately mimic the degradation of pollutants in

OSMs, the path length of the apparatus used in photochemistry experiments must be relative to the measured thickness of the OSM sample.

48 3.4 Conclusion

This study examined the potential of OSMs to facilitate the photodegradation of PAHs (phenanthrene and naphthalene) in freshwater ecosystems. Organic surface microlayers were collected on two separated sampling dates using the glass plate method at Old Woman Creek, Ohio. Observed variation in photoreactivity between the different solvents, bulk water and OSMs, was not evident. These observations were attributed to high dilution factors associated with the sampling technique, and an experimental apparatus that utilized an inappropriately path length. SUVA280 calculations determined that there were significant compositional differences between the two collected fractions. The study did not account for natural conditions that are believed to retard photodegradation such as: light screening, turbidity, water depth, cloud cover, and diurnal cycles. Future research should include developing a sampling mechanism that can collect accurate and precise OSMs. Additionally, a sampling apparatus should be developed that can accurately mimic the physical parameters of OSMs outlined in this study.

3.5. Tables

- Date Sample pH NO3 DOC mg-C/L Abs SUVA280 (ppb) 280nm L/mol-C 7.7.10 Bulk 8.03 23.70 10.37 ± 0.018 0.1737 201 OSM 8.09 21.96 12.52 ± 0.194 0.1762 169 10.22.10 Bulk 8.04 69.19 6.04 ± 0.069 0.1309 260 OSM 8.01 104.11 7.38 ± 0.009 0.1382 225

Table 3.1. Physiochemical properties of collected OSMs and Bulk Waters.

50 3.6. Figures

Figure 3.1. Schematic showing compositional make-up of OSMs. (From Liss and Duce, 1997)

Figure 3.2. Aerial photograph of Old Woman Creek in Huron, Ohio. (Courtesy of Dave Klarer)

-1 Figure 3.3. Observed degradation (kobs (hrs )) of phenanthrene in Milli-Q, bulk water, and OSM.

-1 Figure 3.4. Observed degradation (kobs (hrs )) of naphthalene in Milli-Q, bulk water, and OSM.

Figure 3.5. UV-Vis absorbance spectrum of Bulk Water and OSM collected on 07.07.09.

Figure 3.6. UV-Vis absorbance spectrum of Bulk Water and OSM Collected on 10.22.09.

56 3.7. References

Chen, J., Peijnenburg, W.J.G.M., 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.

Fasnacht M.P., Blough N.V., (2002). Aqueous photodegradation of polycyclic aromatic hydrocarbons. Environ Sci Technol 36: 4364-4369.

Gever, J. R., Mabury, S. A., Crosby, D. G., (1996). Rice field surface microlayers: collection, composition, and pesticide enrichment. Environ. Toxicol. Chem. 15: 1676- 1682.

HARVEY, G., BURZELL, L., (1972). “SIMPLE MICROLAYER METHOD FOR SMALL SAMPLES.” LIMNOLOGY AND OCEANOGRAPHY. 17:156-157.

Jacobs, L., Weavers, L., Chin, Y., (2008). Direct and indirect photolysis of polycyclic aromatic hydrocarbons in nitrate-rich surface waters. ENVIRONMENTAL TOXICOLOGY AND CHEMISTRY. 27: 1643-1648.

Kelly, T, et al. (1991). ATMOSPHERIC AND TRIBUTARY INPUTS OF TOXIC- SUBSTANCES TO LAKE ERIE. JOURNAL OF GREAT LAKES RESEARCH. 17: 504-516.

Lim, L., Wurl, O., Karuppiah, S., Obbard, J., (2007). Atmospheric wet deposition of PAHs to the sea-surface microlayer. MARINE POLLUTION BULLETIN. 54:1212- 1219.

Liss, P.S. and Duce, R.S., (1997). The Sea Surface and Global Change, Cambridge Univ. Press, Cambridge, UK.

57 Liu, K., Dickhut, R., (1997). Surface microlayer enrichment of polycyclic aromatic hydrocarbons in Southern Chesapeake Bay. ENVIRONMENTAL SCIENCE & TECHNOLOGY. 31: 2777-2781.

Liu, K., Dickhut, R.M., (1998). Effects of wind speed and particulate matter sources on surface microlayer characteristics and enrichment of organic matter in southern Chesapeake Bay. J. Geophys. Res. 103: 10571-10577.

Mearns, A, et al. (2009). Effects of Pollution on Marine Organisms. WATER ENVIRONMENT RESEARCH. 81: 2070-2125.

Old Woman Creek National Estuarine Research Reserve, (2003). Annual Report.

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Chapter 4: Conclusions and Future Research

The fate of NPS pollutants in freshwater ecosystems is complex. The combination of multiple contamination sources, hundreds of compounds, and several transformation mechanisms create a convoluted web of potential removal pathways.

Examining the fate of NPS pollutants in wetland ecosystems provides insight into the efficacy of natural attenuation towards the remediation of anthropogenically- produced contaminants. Photochemistry has been shown to be an important degradation mechanism in these systems.

This study had two central purposes. The first was, to elucidate the prominent photo degradation pathways (i.e., direct vs. indirect) of synthetic musk fragrances in natural systems. This was done by monitoring observed kinetic degradation changes following the addition of photosensitizers and quenchers to photolysis solutions.

Indirect photochemical pathways were once thought to be negligible, however, results from this study show that they are prominent and environmentally relevant. HHCB was shown to readily react with photochemically produced OH, yet this reaction is

- highly dependent on the concentration of photosensitizers (i.e., DOM, NO3 ) present in natural waters. Nonetheless, hydroxylation of HHCB is appears to be a significant remediation mechanism in natural systems. The addition of photosensitizers and

59 quenchers to AHTN photolysis solutions indicates strong evidence for the

1 involvement of photochemically generated O2 in its photofate. While these results suggest that natural and created wetlands indeed serve as viable ecosystems for contaminant remediation, additional work needs to be done to fully assess the fate of polycyclic musk fragrances.

Future research should include, but not be limited to: (i) the determination of a

nd 1 2 order rate constant for the reaction of AHTN with O2 , and (ii) the isolation and identification of photoproducts. In order to accurately estimate the role of

1 nd photochemically derived O2, a 2 order reaction rate must be determined. Results showed that AHTN is susceptible to efficient direct photolysis so it is quite possible that indirect pathways could prove to be less important than direct pathways. Yet, without an experimentally determined 2nd order reaction rate, this statement remains unproven. The isolation and identification of photoproducts through the means of advanced analytical techniques (i.e., mass spectrometry, NMR) is the obvious next step in fully understanding the fate of these compounds. Proposed photoproducts were provided, but it remains to be seen whether these products are stable. Once isolated and identified, toxicology experiments should determine whether or not these photoproducts exhibit detrimental effects to aquatic biota that are similar to their parent compounds.

The second central purpose of this study was to determine the efficacy of

OSMs as photoreactors that could potentially enhance the direct photodegradation of accumulated contaminants in this matrix. Results were limited, but did offer significant insight into changed that need to be considered for future work. Two key

60 changes that should be made in future experiments include: (i) the use of multiple sampling techniques, and (ii) the development of an experimental apparatus that accurately mimics the depth of OSMs. The use of multiple sampling techniques will offer another mechanism to evaluate physical properties (i.e., thickness, composition, extent of dilution) of collected microlayers. There is no doubt that the photodegradation of accumulated contaminants occurs in OSMs. But the ability to quantify the potential kinetic enhancement of photolysis between underlying bulk waters and OSMs is entirely dependent on the use of an experimental apparatus that mimics the properties of this unique matrix.

Partitioning, and subsequent accumulation, of organic contaminants into

OSMs is a function of a particular compound’s hydrophobicity. Previous studies have detected several groups of compounds to accumulate in OSMs including: PAHs,

PCBs, pesticides, and heavy metals. Yet, no studies have investigated the potential for hydrophobic pharmaceutical and personal care products (PPCPs) to accumulate in this matrix. Solely based on physical properties (i.e. Kow, aqueous solubility), several

PPCPs, including polycyclic musk fragrances, could accumulate and therefore be susceptible to enhanced direct photochemical degradation.

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