Photochemistry and Toxicity of Triclosan, Triclocarban, and Their Photoproducts and Mixtures in Freshwater Systems

Photochemistry and Toxicity of Triclosan, Triclocarban, and their Photoproducts and Mixtures in Freshwater Systems

Dissertation

Presented in Partial Fulfillment of the Requirements for the Degree Doctor of Philosophy In the Graduate School of The Ohio State University

Katie A. Albanese, B.S. Graduate Program in Environmental Science The Ohio State University 2016

Dissertation Committee: Roman P. Lanno, Advisor Yu-Ping Chin, Co-Advisor Christopher M. Hadad John J. Lenhart

Abstract Triclosan (TCS) and triclocarban (TCC) are two antimicrobial compounds ubiquitously found in surface waters throughout the world. While an abundance of studies exist for the photolysis and toxicity of TCS, the environmental fate and toxicity of

TCC has not been described in nearly as much depth. Both compounds are known to undergo photolytic degradation, and while the photoproducts of TCS have been determined, the products of TCC photolysis are currently unknown. These compounds are both commonly found simultaneously in these surface waters, yet the effects of mixture exposures of these compounds to aquatic organisms is not known. The aims of these projects are to determine the relative toxicities of TCS, TCC and their photolysis products, and to determine the structure of any TCC photoproducts deemed important due to toxic effects. The effects of dissolved organic matter (DOM), which is found in all aquatic systems, on acute toxicities will be determined. Further, the effects of mixtures of these parent compounds and photoproducts will be studied to assess toxicities that could be exerted on organisms in natural aquatic systems due to their potential exposure to multiple compounds and photoproducts.

Acute toxicities were tested using Daphnia magna in 96-hour lethal concentration-50% mortality (LC50) assays. Daphnia magna are a commonly used freshwater test organism and many standard protocols exist for these tests (United States

Environmental Protection Agency (USEPA)). Tests were conducted for parent TCS, ii

parent TCC, and the photolyzed compounds with and without Suwannee River Natural

Organic Matter (SRNOM), the model DOM, in a standard synthetic moderately hard water (known herein as EPA water). LC50s were calculated to compare relative toxicities of each compound, its photoproducts, and the effects of water type on toxicity. DOM has no impact on the toxicity of TCS (LC50 of 1.62µM (95% confidence interval (CI) 0.920-

2.864) without DOM, 1.76 µM (95% CI 1.174-2.652) with DOM), but TCS was significantly detoxified after photolysis, although water composition also had no effect on the toxicities of the photolysis products (LC50s of 8.48µM (95% CI 6.681-10.77) without

DOM, 8.51µM (95% CI 2.700-26.84) with DOM). The toxicity of TCC was not affected by the presence of DOM (LC50 of 0.087µM (95% CI 0.040-0.191) without DOM, LC50 of

0.147 µM (95% CI 0.050-0.434) with DOM), and TCC photolyzed in EPA water alone was significantly less toxic (LC50 of 2.67µM, 95% CI 2.067-3.443). However, when TCC was photolyzed with DOM, the toxicity of the photoproducts did not decrease (LC50 of

0.032µM, 95% CI 0.017-0.060), indicating that DOM mediates the transformation of

TCC to toxic photoproducts. After analysis of photoproducts in various waters with various DOM types, a mechanism for this interaction was described, which states that

UV irradiation leads to triplet-state DOM, which then donates an electron to TCC, causing TCC to cleave at one of two bonds.

Using data from acute toxicity tests, appropriate binary mixture concentrations were calculated to determine the effects of these compounds and photoproducts simultaneously versus their individual effects. To conduct these tests, LC50s were used as the basis for toxic units (TUs), with a mixture of ½ the LC50 concentration of one compound and ½ the LC50 of another compound representing one TU, or LC50, at which

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50% mortality would be expected if toxicity was additive. Mixture toxicity tests were conducted in the same manner as single chemical tests, using a range of TUs as the different concentrations to estimate 96-h TU50s, or the number of TUs at which 50% mortality was observed after 96 h. These TU50s indicated the type, if any, of interaction occurring between D. magna and multiple compounds versus the effects of a single compound. Results indicated that there were synergistic and additive interactions among the compounds and/or photoproducts, with TU50s ranging from 0.025 to 1.45. While the exact reasons for these mixtures effects were not determined, the importance of understanding the complexity of real-life organism exposure to multiple chemicals was further exemplified.

During initial photolysis experiments aimed at determining the proper exposure time of the compounds to nearly fully degrade for toxicity tests, a difference in photolysis rates of TCS was observed in waters with different compositions (i.e., ultrapure “Milli-

Q” water versus EPA water). To assess the reasons for these differences, the kinetics of

TCS photolysis were determined for waters containing a single constituent of EPA water.

Results indicated that bicarbonate buffers in the EPA water enhanced the direct photolysis of TCS (kobs of 0.7109/min vs 0.0382/min, respectively) at pH values well below its pKa. Further, experiments were conducted to determine the interaction of TCS and carbonate that leads to these faster photolysis kinetics. The effects of pH and bicarbonate concentration on the TCS photolysis rate were tested, with results indicating that bicarbonate has a stronger influence on the photolysis rate at higher concentrations and at lower pH values below TCS pKa. To assess the role of deprotonation on the direct photolysis of TCS I used methyl-triclosan (MeTCS), a non-acidic analog of TCS.

MeTCS was not degraded in any experiments, regardless of solution pH or bicarbonate concentration, demonstrating that bicarbonate enhances TCS photolysis via an interaction with the ionizable hydroxyl group at pH values where it should be fully protonated. A mechanism, which suggests that excited-state TCS and bicarbonate interact to form an ion pair, leading to the deprotonation of TCS, was proposed to describe this interaction and how it affects TCS’ direct photolysis.

For V and O

Acknowledgements

I would first like to thank my advisors, Dr. Yo Chin and Dr. Roman Lanno, for help with the planning, support, and implementation of this research. Thank you to Dr.

Christopher Hadad and Dr. John Lenhart for their advice and recommendations and for serving on my committee.

There are five people who were especially helpful during my time in graduate school, and I would like to thank them for all of their advice, support, and mentoring over the past seven years. Thanks to John Phillips for being my emotional support system and for taking care of me during the last few years, Sarah Bowman and Maya Wei-Haas for being such wonderful friends, labmates, and mentors, and Lisa Falland for always being available to give advice late at night after a stressful day, and for the continuous support and encouragement all the way from New Zealand. Finally, thank you to Mitch Phelps, my first research advisor, who has always been there to give advice and act as an unofficial advisor for both my academic career and life in general.

There are also so many other people who have been helpful to me in making it through my PhD: thanks to my mom, Loraine Allison, and Blanche Luczyk for helping take care of my dogs so I could travel to conferences; Alan Jones for his support and advice on career decisions; Kyle Dzwigalski for his never ending encouragement and positivity during my last year as a graduate student; Chloe, Meg, Hannah, Jordyn, and

Bob for the decompression time in the dog park; Kelly and Natalie for always being there

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to talk and comfort; Heather Nutter, Michael Luczyk, and Tycho Luczyk for the never ending laughs and adventures that helped me keep everything in perspective; my dad Bill and stepmom Kelly S. for their support; the many patrons and employees at the Knotty

Pine who were always excited to hear about my research and always encouraged me to never give up; Ramone for the stress-relieving running workouts and the many heart to hearts; Bill Margiotta for the calming freediving sessions; Shelby for staying up late with me to work on assignments and data analysis; Cindy for allowing me to TA under her for seven terms, and for being willing to accommodate my schedule and for her words of wisdom on the PhD process; Sue, Corey, Chanelle, Lynn, and Teresa for all of their administrative help; Mitch Phelps for always providing advice on academics, careers, and life in general; Owen Rings for always making every problem in the world disappear for a while.

Special thanks to those who provided lab support: Mrinal Chakraborty, Molly

Semones, Chenyi Yuan, Sarah Bowman, Brandon McAdams, Maya Wei-Haas, Allison

Kreinberg, Kate Ziegelgruber, Marcy Card McCarty, Steven Nagel, Cody Holland, Smiti

Gupta, Katelyn Schockman, Arus Glover, Lisa Falland, and Kelly A. Albanese. Also thank you to George Keeney for supplying Daphnia and alfalfa pellets, and to Sandy

Jones for free-drying samples.

I would also like to acknowledge research funding and travel support from the following: The Ohio State University R.H. Edgerley Environmental Toxicology Fund;

The Ohio State University Environmental Science Graduate Program; The Ohio State

University Department of Evolution, Ecology, and Organismal Biology; The Ohio State

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University School of Earth Sciences; the Society of Environmental Toxicology and

Chemistry; and The Ohio State University Council of Graduate Students.

Finally, I want to thank Vincent and Otago for being the best support system I could have ever hoped for throughout this journey. I love you, and would never have been able to do this without you.

Vita

2004...... Grandview Heights High School

2008 ...... B.S. Biology, The University of North

Carolina Wilmington

2009 to present ...... Graduate Student, Environmental Science

Graduate Program, The Ohio State

University

Publications

William Blum, Mitch A. Phelps, Rebecca B. Klisovic, Darlene M. Rozewski, Wenjun Ni, Katie A. Albanese, Brad Rovin, Cheryl Kefauver, Steven M. Devine, David M. Lucas, Amy Johnson, Larry J. Schaaf, John C. Byrd, Guido Marcucci, and Michael R. Grever. 2010. Phase I clinical and Pharmacokinetic study of a novel schedule of flavopiridol in relapsed and refractory acute leukemias. Haemotologica 95(7): 1098-1105

Mitch A. Phelps, Darlene M. Rozewski, Jeffrey S. Johnston, Katherine L. Farley, Katie A. Albanese, John C. Byrd, Thomas S. Lin, Michael R. Grever, and James T. Dalton. 2008. Development and validation of a sensitive liquid chromatography/mass spectrometry method for quantitation of flavopiridol in plasma enables accurate estimation of pharmacokinetic parameters with a clinically active dosing schedule. Journal of Chromatography B 868: 110-115

Fields of Study Major Field: Environmental Science Area of Emphasis: Aquatic Ecotoxicology and Chemistry

Table of Contents

Abstract ...... ii

Dedication ...... vi

Acknowledgements ...... vii

Vita ...... x

List of Tables ...... xv

List of Figures ...... xvi

List of Abbreviations ...... xx

Chapter 1: Introduction ...... 1

1.1 References ...... 9

Chapter 2: Photolysis- and dissolved organic matter-induced toxicity of triclocarban to Daphnia magna ...... 14

2.1 Introduction ...... 14 2.2 Methods/Results and Discussion ...... 15 2.3 Supporting Information ...... 20 2.3.1 Materials ...... 20 2.3.2 Methods...... 20 2.3.2.1 EPA water ...... 20 2.3.2.2 Half-life determination...... 21 2.3.2.3 HPLC analysis ...... 21 2.3.2.4 Daphnia magna culture maintenance ...... 22 2.3.2.5 Acute toxicity testing ...... 22 2.3.2.6 WWTP effluent collection/photolysis...... 23 2.3.2.7 Effects of DOM composition ...... 23 2.3.2.8 GC/MS analysis ...... 24 2.3.2.9 Data analysis ...... 25 2.3.3 Supplemental experiments ...... 25 2.3.3.1 No mortality concentrations ...... 25 2.3.3.2 Determination of irradiation time ...... 25 xii

2.3.3.3 Toxicity of SRNOM ...... 26 2.3.3.4 Residual TCC ...... 26 2.4 Tables ...... 28 2.5 Figures...... 29 2.6 References ...... 32

Chapter 3: Mixture toxicities of triclosan, triclocarban, and their photolysis products to Daphnia magna ...... 34

3.1 Introduction ...... 34 3.2 Materials and Methods ...... 37 3.2.1 Materials ...... 37 3.2.2 Test water ...... 38 3.2.3 Maintenance of test organisms ...... 38 3.2.4 Acute toxicity testing with Daphnia magna ...... 38 3.2.5 Determination of photolysis times ...... 40 3.2.6 Calculation of toxic units ...... 41 3.2.7 Mixture toxicity testing ...... 42 3.2.8 Data analysis ...... 43 3.3 Results ...... 43 3.3.1 Effects of photolysis on triclosan toxicity ...... 43 3.3.2 Effects of mixtures ...... 44 3.4 Discussion ...... 45 3.5 Tables ...... 50 3.6 Figures...... 51 3.7 References ...... 53

Chapter 4: Enhancement of triclosan photolysis by carbonate ion-pair formation ...... 60

4.1 Introduction ...... 60 4.2 Materials and methods ...... 62 4.2.1 Chemicals ...... 62 4.2.2 Photolysis solution preparation ...... 62 4.2.3 Photolysis reactions ...... 63 4.2.4 Methyl triclosan synthesis...... 64 4.2.5 HPLC analysis ...... 64 4.2.6 Data analysis ...... 65 4.3 Results ...... 65 4.3.1 The effects of solution pH on direct photolysis ...... 65 4.3.2 Effects of MgSO4/CaSO4 ...... 66 4.3.3 Buffer effects ...... 66 4.3.4 Effects of radical scavengers ...... 67 4.3.5 Photolysis of methyl-triclosan ...... 69

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4.3.6 Effects of bicarbonate concentration ...... 69 4.4 Discussion ...... 70 4.4.1 Effects of ionization ...... 70 4.4.2 Mechanism of bicarbonate/carbonate enhancement ...... 71 4.4.3 Environmental Implications ...... 71 4.5 Figures...... 73 4.6 References ...... 80

Chapter 5: Conclusions ...... 85

5.1 References ...... 89

Appendix A: Chemical structures ...... 90

Appendix B: Mass spectra of TCC and photolysis products ...... 97

Appendix C: LC50 curves ...... 106

Appendix D: NMR of synthesized methyl triclosan ...... 140

Appendix E: Toxicity of dissolved organic matter ...... 144

Appendix F: LC50s and confidence intervals ...... 146

Appendix G: Compiled toxicity data ...... 150

Appendix H: Toxicity data in SI units ...... 152

Bibliography ...... 154

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List of Tables

Table 2.1 Photolysis products of TCC in different DOM solutions ...... 28

Table 3.1 Potential photolysis product exposure of aquatic organisms ...... 50

Table E.1 Toxicity of DOM ...... 145

Table F.1 LC50s and confidence intervals of single compounds ...... 147

Table F.2 TU50s and confidence intervals of mixtures ...... 148

Table F.3 Slopes and 95% confidence intervals of photolysis experiments ...... 149

Table G.1 Compiled toxicity data ...... 151

Table H.1 Toxicity data in SI units ...... 153

List of Figures

Figure 2.1 Toxicity of parent and photolyzed TCC with and without DOM ...... 29

Figure 2.2 Toxicity of TCC photolysis products ...... 30

Figure 2.3 Proposed mechanism of TCC photolysis with DOM ...... 31

Figure 3.1 Toxicity of parent and photolyzed TCS with and without DOM ...... 51

Figure 3.2 Mixture toxicities...... 52

Figure 4.1 Photolysis of buffered and unbuffered waters ...... 73

Figure 4.2 Effects of bicarbonate and sulfates ...... 74

Figure 4.3 Effects of different buffers ...... 75

Figure 4.4 Photolysis with DABCO ...... 76

Figure 4.5 Methyl triclosan photolysis ...... 77

Figure 4.6 Effects of bicarbonate concentration on photolysis rates ...... 78

Figure 4.7 Proposed TCS mechanism ...... 79

Figure A.1 Structure of triclosan ...... 91

Figure A.2 Structure of triclocarban ...... 92

Figure A.3 Structure of 4-chloroaniline ...... 93

Figure A.4 Structure of 3,4-dichloroaniline ...... 94 xvi

Figure A.5 Structure of 4-chlorophenyl isocyanate ...... 95

Figure A.6 Structure of 3,4-dichlorophenyl isocyanate ...... 96

Figure B.1 Mass spectra of photolyzed triclocarban in EPA-DOM water ...... 98

Figure B.2 Mass spectra of TCC photolyzed in EPA with DOM ...... 99

Figure B.3 Mass spectra of 4-chloroaniline standard ...... 100

Figure B.4 Mass spectra of 3,4-dichloroaniline standard ...... 101

Figure B.5 Mass spectra of 4-chlorophenyl isocyanate standard ...... 102

Figure B.6 Mass spectra of 3,4-dichlorophenyl isocyanate standard ...... 103

Figure B.7 Comparison of photolyzed TCC and product standards ...... 104

Figure B.8 Mass spectra of triclosan photolyzed in EPA without DOM...... 105

Figure C.1 LC50 curve of parent TCC in EPA water ...... 107

Figure C.2 LC50 curve of parent TCC in EPA-DOM water ...... 108

Figure C.3 LC50 curve of photolyzed TCC in EPA water ...... 109

Figure C.4 LC50 curve of photolyzed TCC in EPA water ...... 110

Figure C.5 LC50 curve of photolyzed TCC in EPA water ...... 111

Figure C.6 LC50 curve of photolyzed TCC in EPA-DOM water ...... 112

Figure C.7 LC50 curve of photolyzed TCC in EPA-DOM water ...... 113

Figure C.8 LC50 curve of photolyzed TCC in EPA-DOM water ...... 114

Figure C.9 LC50 curve of 4-CA in EPA water ...... 115

Figure C.10 LC50 curve of 4-CA in EPA water ...... 116

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Figure C.11 LC50 curve of 3,4-DCA in EPA water ...... 117

Figure C.12 LC50 curve of 3,4-DCA in EPA water ...... 118

Figure C.13 LC50 curve of 4-CPI in EPA water ...... 119

Figure C.14 LC50 curve of 4-CPI in EPA water ...... 120

Figure C.15 LC50 curve of 3,4-DCPI in EPA water ...... 121

Figure C.16 LC50 curve of 3,4-DCPI in EPA water ...... 122

Figure C.17 LC50 curve of parent TCS in EPA water...... 123

Figure C.18 LC50 curve of parent TCS in EPA-DOM water ...... 124

Figure C.19 LC50 curve of photolyzed TCS in EPA water ...... 125

Figure C.20 LC50 curve of photolyzed TCS in EPA-DOM water ...... 126

Figure C.21 TU50 curve of parent TCC and parent TCS mixture in EPA ...... 127

Figure C.22 TU50 curve of parent TCC and parent TCS mixture in EPA ...... 128

Figure C.23 TU50 curve of parent TCC and parent TCS mixture in EPA ...... 129

Figure C.24 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water ...... 130

Figure C.25 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water ...... 131

Figure C.26 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water ...... 132

Figure C.27 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water ...... 133

Figure C.28 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water ...... 134

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Figure C.29 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water ...... 135

Figure C.30 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water ...... 136

Figure C.31 TU50 curve of mixture of TCC photolyzed in EPA-DOM water and TCS photolyzed in EPA water...... 137

Figure C.32 TU50 curve of mixture of TCC photolyzed in EPA-DOM water and TCS photolyzed in EPA-DOM water ...... 138

Figure C.33 TU50 curve of mixture of TCC and TCS photolyzed in same solution of EPA-DOM water ...... 139

Figure D.1 1H NMR of synthesized Me-TCS ...... 141

Figure D.2 13C NMR of synthesized Me-TCS ...... 142

Figure D.3 HMQC NMR of synthesized Me-TCS ...... 143

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List of Abbreviations

1,4-diazabicyclo[2.2.2]octane……………………………………………………..DABCO 2-dimensional nuclear magnetic resonance spectroscopy……………………….2-D NMR 3,4-chloroaniline……………………………………………………………………3,4-CA 3,4-chlorophenyl isocyante…………..…………………………………………….3,4-CPI 4-chloroaniline………………………………………………………………………..4-CA 4-chlorophenyl isocyante…………………………………………………………….4-CPI 95% Confidence interval…………………………………………………………....95%CI Acetonitrile……………………………………………………………………………ACN

Acid dissociation constant……………………………………………………...………..Ka

Acid dissociation constant negative base-10 logarithm………………………..……….pKa

- Bicarbonate……………………………………………………..……………………HCO3

- Bisulfate…………………………………………………………………….………..HSO4

Calcium sulfate…………………………………………..………………………….CaSO4 Carbon-13 nuclear magnetic resonance spectroscopy……………………..…….13C-NMR

2- Carbonate……………………………………………………………………………..CO3

•- Carbonate radical………………………………………….………………………….CO3

Carbonic acid………………………………………………………………………..H2CO3 Celsius…………………………………………………………………………………….C Chloroaniline……………………………………………………………………………CA Chlorophenyl isocyante………………………………………………………………...CPI

Dichloromethane……………………………………………………………………CH2Cl2 Dissolved organic carbon……………………………………………………………..DOC

Dissolved organic matter……………………………………………………………..DOM Environmental Protection Agency…………………………………………………….EPA Fulvic acid……………………………………………………………………………….FA Gas chromatograph..……………………………………………………………….……GC Gas chromatography coupled with mass spectrometry………………...... ………..GC-MS

Half-life…………………………………………………………………………………t1/2 Heteronuclear multiple quantum coherence………………………………………..HMQC High-pressure liquid chromatography………………………………………………HPLC Humic acid……………………………………………………………………………...HA Hydrochloric acid………………………………………………………………………HCl Hydrogen ion molarity negative base-10 logarithm…………………………………….pH

Hydrogen peroxide……………………………………..……………………………..H2O2 Hydroxyl radical………………………………………….…………………………….OH•

Lethal concentration of 50% of population………………..…………………………..LC50

Lethal concentration of 50% of population after 96 hours………………..…….96-hr LC50 Little Miami River humic acid…………………………………………………....LMRHA Lower confidence limit……………………………………………………………..…LCL

Magnesium sulfate………………………………………………………………….MgSO4 Mass spectrometry……………………………………………………………………...MS Mass to charge ratio……………………………………………………………………m/z Methanol…………………………………………………………………………….MeOH Methyl triclosan………………………………………………………………….....MeTCS Methylmelhacrylate resin……………………………………………………………XAD8 Milli-Q water…………………………………………………………………………...MQ Natural organic matter……………………………………………………………….NOM

- Nitrate…………………………………………………………………………………NO3

- Nitrite……………………………………………………………………………….....NO2 No observed effect concentration……………………………………..…………….NOEC xxi

Nuclear magnetic resonance spectroscopy…………………………………………...NMR

Observed photolysis rate constant……………………...……………………………….kobs

Octanol-water partition coefficient………………………………………...……………kow Old Woman Creek fulvic acid…………………………………………………….OWCFA Photolyzed…………………………………………………………………………….….ph Pony Lake fulvic acid………………………………………………………………...PLFA Positive ion mode electrospray ionization………………………………….………….ES+ Potassium chloride……………………………….…………………………………….KCl Proton nuclear magnetic resonance spectroscopy……………………………..….1H-NMR Reactive oxygen species………………………………………………………………ROS Reverse osmosis………………………………………………………………………...RO

1 Singlet oxygen…………………………………………………………………….……. O2

Sodium bicarbonate………….……………………………………………………NaHCO3 Sodium hydroxide……………..…………………………………………………….NaOH

Sodium sulfate………………………………………………………………..…….Na2SO4

2- Sulfate…………………………………………………………………………………SO4

Sulfuric acid…………………………………………………………………………H2SO4

•- Superoxide……………………………………..…………………………….…………O2 Suwannee river fulvic acid…………………………………………………………...SRFA Suwannee river natural organic matter……………………………………………SRNOM Total organic carbon…………………………………………………………………..TOC Toxic unit………………………………………………………………………………..TU

Toxic units lethal to 50% of a population……………………………………………..TU50 Triclocarban…………………………………………………………………………...TCC Triclosan……………………………………………………………………………….TCS Triplet excited state dissolved organic matter………………………………………3DOM United States Environmental Protection Agency………………………………….USEPA Upper confidence limit……………………………………………………………..…UCL xxii

Wastewater treatment plant……..……………………………………….…………WWTP Wastewater treatment plant effluent……………………...………………………...WWTP

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Chapter 1: Introduction

Triclosan (TCS) and Triclocarban (TCC) are two antimicrobial compounds that are used in many common household products, including liquid and bar soaps, toothpaste, diapers, and various cleaning supplies. Their antimicrobial properties are due to the inhibition of fatty acid synthesis by bacteria caused by interrupting an enzyme not present in humans (McMurry et al., 1998; Levy et al., 1999). This prevents bacteria from producing cell membranes that are vital to survival and replication.

Due to the high volume usage of these compounds, TCS and TCC are often found in wastewater treatment plant (WWTP) influent. Despite the fact that WWTPs are very efficient at removing these substances they still enter rivers and streams in treated effluent at trace levels (Halden and Paul, 2005; Zhao et al., 2010; Zhao et al., 2013;

Guatam et al., 2013;), with concentrations as high as 0.02 µM TCC (Halden and Paull,

2005) and 0.008 µM TCS (Kolpin et al., 2002) reported for surface waters. Once in the receiving waters they can undergo a number of attenuation processes such as photolysis and sorption. With respect to sorption, both compounds can be highly particle-reactive due to their hydrophobicity as estimated by their high log Kow values (TCS: 4.8-5.4, Ying and Kookana, 2007; Son et al., 2009; TCC: 4.2, US EPA 2008). As a result they are often found associated with particulates in sediment core profiles which have been

used to describe usage trends of these compounds over time (Buth et al., 2010; Anger et al., 2013).

The environmental fate of TCS has been widely studied, with many peer- reviewed articles reporting its persistence in the environment (Kolpin et al., 2002; Singer at al., 2002; Buth et al., 2010; Buth et al., 2011; Venkatesan et al, 2012; Gautum et al.,

2014; Anger et al., 2015). Photolysis has been shown to be an important attenuation pathway leading to the formation of various potentially toxic photoproducts (e.g., dichlorodibenzodioxins; Latch et al., 2003; Mezcua et al., 2004; Latch et al., 2005;

Sanchez-Prado et al., 2006; Buth et al., 2009). Additionally, in WWTPs that use chlorination to disinfect the polished effluent, TCS can become further chlorinated, and undergo further reactions to form other chlorinated compounds that can be released into the aquatic environment (Buth et al., 2011). Despite the significant number of studies on the photolysis rate and degradation products of TCS, certain aspects of this process remain a mystery. In this dissertation I will demonstrate the effects of buffers such as bicarbonate on the direct photolysis of TCS and how it may affect the photofate of TCS in aquatic systems.

The toxicity of TCS has also been examined in a number of organisms. Orvos et al. (2002) found 48-h LC50 values (i.e., concentration of chemical that results in 50% mortality of a test population over a specified time period) of 0.069 µM for Ceriodaphnia dubia, and 1.347 µM for Daphnia magna, two species of water fleas. This study also noted a 96-h LC50 of 0.898 µM for fathead minnows. DeLorenzo et al. (2008) observed decreases in phytoplankton cell density (EC50 of 0.012 µM) with TCS exposure, and that bioluminescent bacteria Vibrio fischeri luminescence (15-min IC50 of 0.183 µM) and

grass shrimp Palaemonetes pugio mortality (96-h LC50s of 0.532-2.248 µM) were also significantly affected. In comparison, the acute toxicities of two well-known pesticides, atrazine and chlorpyrifos, to Vibrio fischeri (30 min bioluminescence EC50) and Daphnia magna (48-hr LC50) were measured. Chlorpyrifos was the most toxic of the two, with

EC50s of 8.243 µM and 2.111 µM, respectively, while atrazine was slightly less toxic, with EC50s of 321.758 µM and 164.588 µM, respectively (Palma et al., 2008). Chronic toxicity tests have examined estrogenic effects in aquatic organisms such as increased vitellogenin and vitellogenin mRNA in Japanese medaka larvae (2.079 µM water concentration for a 21-d exposure) and increased vitellogenin mRNA expression and decreased sperm counts in male western mosquitofish (exposure concentration of 0.350

µM for 35 d) (Ishibashi et al., 2004; Raut and Angus, 2010). Finally, TCS has also been shown to bioaccumulate, with bioaccumulation factors (BAFs) ranging from 500 in snails and up to 1,400 in algae (Coogan et al., 2007; Coogan et al., 2008).

In contrast to TCS, fewer studies have examined the toxicity and photochemistry of TCC. While BAFs have been reported in aquatic organisms, ranging from 1,600 in snails to 1,900 in algae (Coogan et al., 2007; Coogan et al., 2008), little is known about its toxicity to aquatic organisms. The photolysis of TCC has only recently been studied, and not extensively, with research focusing on the indirect photolysis rates of TCC in water (Lv et al., 2014) with different types of dissolved organic matter (DOM) (Guerard et al., 2009; Trouts and Chin, 2015). Research and GC-MS results of one group has suggested that TCC can potentially form toxic photoproducts (Ding et al., 2013; Ding et al., 2015).

While the toxic effects of a single compound are commonly tested, this does not always represent the true toxicity experienced by organisms in the environment as they are often exposed to mixtures of compounds. Since chemical mixtures can be extremely complex, with the numbers of chemicals an organism is exposed to at one time potentially numbering in the hundreds, it is very difficult to assess the actual risk of chemicals to organisms in this type of environment. Thus, it would be useful to study simple binary mixtures of compounds that are commonly found concurrently in the environment, such as TCS and TCC, to better understand whether effects on organisms are additive, antagonistic, or synergistic.

Photochemistry in association with ecotoxicology is also widely understudied, with the exception of polycyclic aromatic hydrocarbons (PAHs). Studies of PAHs have shown an increase in toxicity with UV exposure, both in the lab and in-situ (Bowling et al., 1983; Ankley et al., 1995; Diamond et al., 1995; Ireland et al., 1996; Nikkilä et al.,

1999). While the toxicities of parent compounds and the photolysis and environmental fate of these compounds are widely studied in separate experiments, the results are rarely combined to provide a comprehensive overview of the fate and toxicity of these compounds in the aquatic environment.

The effects of DOM on photolysis and toxicity add another layer of complexity to the toxicology/photochemistry conundrum, as DOM can participate in the degradation of these compounds through indirect pathways such as the formation of intermediate, potentially toxic chemicals such as reactive oxygen (ROS), triplets, solvated electrons, and carbon-centered radicals (Zepp et al., 1992; Scully et al., 1996; Holder-Sandrik et al.,

2000; al Housari et al., 2010; Aiken et al., 2011; Wenk et al., 2011). The degree of

degradation that occurs is dependent upon the chemical structure of the target analyte.

The presence of DOM could potentially decrease the apparent toxicity of a compound due to the formation of toxicant-DOM complexes (Day 1991; Haitzer et al., 1998; Kim et al., 1999; Bejerano et al., 2005; Ra et al., 2008; Doig and Liber, 2006). While the exact structure of DOM is not fully understood due to its complexity, it is generally accepted by environmental scientists that DOM is derived from either microbial (autochthonous) sources that are produced within the waterbody, e.g., simple prokaryotic and eukaryotic single cell organisms like algae, bacteria, and archae, or terrestrial plant-derived

(allochthonous) organic matter that enters receiving waters from runoff or groundwater infiltration. These two geochemical endmembers result in DOM with different compositions, where allocthonously-derived organic matter is more aromatic in composition (from lignin), and autochthonous DOM is relatively more enriched in aliphatic (from fatty acids) and carbohydrate moieties. Understanding the effects of

DOM on toxicity and photochemistry is highly important and relevant because DOM is ubiquitous to the aquatic environment globally.

My study will examine the photofate of TCS and TCC and the acute toxicity of these compounds and their photoproducts to a common freshwater crustacean (Daphnia magna). I will also examine the role of DOM with respect to the photofate of TCC and

TCS, products formed, and their acute toxicity, as well as the effects of carbonate on TCS photolysis. My overarching hypothesis is that the indirect photolysis of TCC and TCS by

DOM creates photo-derivatives that differ with respect to composition and toxicity relative to those formed by direct photolysis.

From this hypothesis I have developed three sub-hypotheses that will be met with the following objectives:

1) Hypothesis: Direct and indirect photolysis pathways for TCC and TCS create

different photoproducts with different toxicities. Based on the knowledge that

DOM can interact with organic compounds to enhance their photodegradation

rates, I suspected that this interaction may also lead to the production of

different photoproducts than direct photolysis. Dissolved organic matter-

enhanced photolysis generally occurs via introduction of an excited

component, such as 3DOM, or a different excited molecule generated by

3DOM, which could potentially lead to photolysis of TCS and TCC via

different pathways.

Objectives: Determine the acute toxicities of parent and photolyzed TCC and

TCS and assess the impact of UV exposure and DOM source on toxicity.

2) Hypothesis: Mixtures of TCS, TCC, and photoproducts will exert higher

toxicity than any of these compounds alone. Triclosan and TCC exert their

toxicities via different mechanisms, and I suspected that exposure to one

compound would cause an organism to be more susceptible to the toxic effects

of another. While I do not necessarily know the toxic mechanisms of the TCS

and TCC photoproducts, they are likely not the same as the parent compounds

or each other due to their very different structures. I therefore hypothesized

that exposure to a photodegradate would make an organism more susceptible

to TCS, TCC, or another photodegradate.

Objectives: Determine the acute mixture toxicities of both parent and

daughter TCS and TCC. This will be done using toxic unit equivalents (TUs)

to create mixtures in order to account for differences in LC50s values between

compounds.

3) Hypothesis: Carbonate and phosphate will enhance the photolysis of TCS.

Triclosan is an ionizable compound with a pKa around that of natural waters,

indicating that it will be in both ionized and protonated forms in natural

waters. This potential extensive percent of ionization in natural waters could

allow for the interaction of negatively charged TCS ions with positively

charged or partially charged water constituents. As studies of other

compounds have shown that water constituents can enhance photolysis, I

suspected that TCS could be interacting with charged water constituents,

found in the EPA toxicity testing water, in a manner that would enhance TCS

photolysis. This hypothesis was formed after initial kinetics experiments

conducted for toxicity tests showed an increase in TCS photolysis rates in

EPA water compare to. Milli-Q (ultra pure) water.

Objectives: Assess the effects of these buffers on the photolysis rate of TCS

and the role they play in TCS photofate in waters with pH values much less

than its pKa.

This dissertation is organized as three stand-alone papers that investigate the objectives above. Chapter 2 was written with the intentions of submitting to Science

Magazine, and is formatted as appropriate for Science, while the other chapters follow a more standard format. The results in this study are important in understanding the fate of 7

TCS and TCC in the natural environment. The method of toxicity testing of photolysis products described in this dissertation is an important step in better understanding the true potential toxicity of a compound introduced into the environment, and can serve as a valuable tool for future studies.

1.1 References

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Buth, J.M., M.R. Ross, K. McNeill, & W.A. Arnold. 2011. Removal and formation of chlorinated triclosan derivatives in wastewater treatment plans using chlorine and UV disinfection. Chemosphere. 84: 1238-1243. Coogan, M.A., R.E. Edziyie, T.W. La Point, & B.J. Venables. 2007. Algal bioaccumulation of triclocarban, triclosan, and methyl-triclosan in a North Texas wastewater treatment plant receiving stream. Chemosphere 67: 1911-1918. Coogan, M.A., & T.W. La Point. 2008. Snail bioaccumulation of triclocarban, triclosan, and methyltriclosan in a North Texas, USA stream affected by wastewater treatment plan runoff. Environmental Toxicology and Chemistry 27: 1788-1793. Day, K.E. 1991. Effects of dissolved organic carbon on accumulation and acute toxicity of fenvalerate, deltamethrin and cyhalothrin to Daphnia magna (straus). Environmental Toxicology and Chemistry 10: 91-101. DeLorenzo, M.E., J.M. Keller, C.D. Arthur, M.C. Finnegan, H.E. Harper, V.L. Winder, & D.L. Zdankiewicz. 2008. Toxicity of the antimicrobial compound triclosan and the formation of the metabolite methyl-triclosan in estuarine systems. Environmental Toxicology 23: 224-232. Diamond, S.A., J.T. Oris, & S.I. Guttman. 1995. Adaptation of fluoroanthene exposure in a laboratory population of fathead minnows. Environmental Toxicology and Chemistry 14: 1393-1400. Ding, S.-L., X.-K. Wang, W.-Q. Jiang, X. Meng, R.-S. Zhao, C. Wang, & X. Wang. 2013. Photodegradation of the antimicrobial triclocarban in aqueous systems under ultraviolet radiation. Environmental Science and Pollution Research 20: 3195-3201. Ding, S.-L., X.-K. Wang, W.-Q. Jiang, R.-S. Zhao, T.-T. Shen, C. Wang, & X. Wang. 2015. Effect of pH, inorganic ions, and dissolved organic matter on the photolysis of antimicrobial triclocarban in aqueous systems under simulated sunlight irradiation. Environmental Science and Pollution Research 22: 5204-5211. Doig, L.E., K. Liber. 2006. Influence of dissolved organic matter on nickel bioavailability and toxicity to Hyalella azteca in water-only exposures. Aquatic Toxicology 76: 203-216. Gautum, P., J.S. Carsella, & C.A. Kinney. 2014. Presence and transport of the antimicrobials triclocarban and triclosan in a wastewater-dominated stream and freshwater environment. Water Research 48: 247-256. Guerard, J.J., P.L. Miller, T.D. Trouts, & Y.-P. Chin. 2009. The role of fulvic acid composition in the photosensitized degradation of aquatic contaminants. Aquatic Sciences 71: 160-169.

Haitzer, M., S. Hüss, W. Traunspurger, & C. Steinberg. 1998. Effects of dissolved organic matter (DOM) on the bioconcentration of organic chemicals in aquatic organisms – a review. Chemosphere 37: 1335-1362. Halden, R.U., & D.H. Paul. 2005. Co-occurrence of triclocarban and triclosan in U.S. water resources. Environmental Science and Technology 39: 1420-1426. Holder-Sandrik, S.L., P. Bilski, J.D. Pakulski, C.F. Chignell. 2000. Photogeneration of singlet oxygen and free radicals in dissolved organic matter isolated from the Mississippi and Atchafalaya River plumes. Marine Chemistry 69: 139-152. Ireland, D.S., G.A. Burton Jr., & G.G. Hess. 1996. In situ toxicity evaluations of turbidity and photoinduction of polycyclic aromatic hydrocarbons. Environmental Toxicology and Chemistry 15: 574-581. Ishibashi, H., N. Matsumura, M. Hirano, M. Matsouka, H. Shiratsuchi, Y. Ishibashi, Y. Takao, & K. Arizono. 2004. Effects of triclosan on the early life stages and reproduction of medaka Oryzias latipes and induction of hepatic vitellogenin. Aquatic Toxicology 67: 167-179. Kim, S.D., H. Ma, H.E. Allen, & D.K. Cha. 1999. Influence of dissolved organic matter on the toxicity of copper to Ceriodaphnia dubia: effect of complexation kinetics. Environmental Toxicology and Chemistry 18: 2433-2437. Kolpin, D.W., E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber, & H.T. Buxton. 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams 1999-2000: a national reconnaissance. Environmental Science and Technology 36: 1202-1211. Latch, D.E., J.L. Packer, W.A. Arnold, & K. McNeill. 2003. Photochemical conversion of triclosan to 2,8-dichlorodibenzo-p-dioxin in aqueous solution. Journal of Photochemistry and Photobiology A 158: 63-66. Latch, D.E., J.L. Packer, B.L. Stenden, J. VanOverberke, W.A. Arnold, & K. McNeill. 2005. Aqueous photochemistry of triclosan: formation of 2,4-dichlorophenol, 2,8- dichlorodibenzo-p-dioxin, and oligomerization products. Environmental Toxicology and Chemistry 24: 517-525. Levy, C.W., A. Roujeinikova, S. Sedelnikova, P.J. Baker, & A.R. Stuitje. 1999. Molecular basis of triclosan activity. Nature 398: 383-384. Lv, M., Q. Sun, H. Xu, L. Lin, M. Chen, & C.-P. Yu. 2015. Occurrence and fate of triclosan and triclocarban in a subtropical river and its estuary. Marine Pollution Bulletin 88: 383-388. McMurry, L.M., M. Oethinger, & S.B. Levy. 1998. Triclosan targets lipid synthesis. Nature 394: 531-532.

Mezcua, M., M.J. Gómez, I. Ferrer, A. Aquera, M.D. Hernando, A.R. Fernández-Alba. 2004. Evidence of 2,7/2,8-dichlorodibenzo-p-dioxin as a photodegradation product of triclosan in water and wastewater samples. Analytica Chimica Acta 524: 241-257. Nikkilä A., S. Penttinen, & J.V.K. Kukkonen. 1999. UV-B-induced acute toxicity of pyrene to the water flea Daphnia magna in natural freshwaters. Ecotoxicology and Environmental Safety 44: 271-279. Orvos, D.R., D.J. Versteeg, J. Inauen, M. Capdevielle, A. Rothenstein, & V. Cunningham. 2002. Aquatic toxicity of triclosan. Environmental Toxicology and Chemistry 21: 1338-1349. Palma, P., V.L. Palma, R.M. Fernandes, A.M.V.M. Soares, & I.R. Barbosa. 2008. Acute toxicity of atrazine, endosulfin sulfate and chlorpyrifos to Vibrio fischeri, Thamneocephalus platyurus, and Daphnia magna, relative to their concentrations in surface waters from the Alentejo Region of Portugal. Bulletin of Environmental Contamination and Toxicology 81: 485-489. Ra, J.S., S.-Y. Oh, B.C. Lee, & S.D. Kim. 2008. The effect of suspended particles coated by humic acid on the toxicity of pharmaceuticals, estrogens, and phenolic compounds. Environment International 34: 184-192. Raut, S.A., & R.A. Angus. 2010. Triclosan has endocrine-disrupting effects in male western mosquitofish, Gambusia affinis. Environmental Toxicology and Chemistry 29: 1287-1291. Sanchez-Prado, L., M. Llompart, M. Lores, C. García-Jares, J.M. Bayona & R. Cela. 2006. Monitoring the photochemical degradation of triclosan in wastewater by UV light and sunlight using solid-phase microextraction. Chemosphere 65: 1338- 1347. Scully, N.M., D.R.S. Lean, D.J. McQueen, & W.J. Cooper. 1996. Hydrogen peroxide formation: the interaction of ultraviolet radiation and dissolved organic carbon in lake waters along a 43-75°N gradient. Limnology and Oceanography 41: 540- 548. Singer, H., S. Mueller, C. Tixier, & L. Pillonel. 2002. Triclosan: occurrence and fate of a widely used biocide in the aqueous environment: field measures in wastewater treatment plants, surface waters, and lake sediments. Environmental Science and Technology 36: 4998-5004.

Son, H.S., G. Ko, & K.D. Zoh. 2009. Kinetics and mechanism of photolysis and TiO2 photocatalysis of triclosan. Journal of Hazardous Materials 166: 954-960. Trouts, T., & Y.-P. Chin. 2015. Direct and indirect photolysis of triclocarban in the presence of dissolved organic matter. Elementa: Science of the Anthropocene. 12

Venkatesan, A.K., B.F.G. Pycke, R.U. Halden, L.B. Berber, & K.E. Lee. 2012. Occurrence of triclosan, triclocarban, and its lesser chlorinated congeners in Minnesota freshwater sediments collected near wastewater treatment plants. Journal of Hazardous Materials 229-230: 29-35. Wenk, J., U. von Gunten, & S. Canonica. 2011. Effect of dissolved organic matter on the transformation of contaminants induced by excited triplet states and the hydroxyl radical. Environmental Science and Technology 45: 1334-1340. Ying, G.G., & R.S. Kookana. 2007. Triclosan in wastewaters and biosolids from Australian wastewater treatment plants. Environment International 33: 199-205. Zepp, R.G., B.C. Faust, & J. Hoigne. 1992. Hydroxyl radical formation in aqueous reactions (pH 3-8) or iron (II) with hydrogen peroxide: the photo-Fenton reaction. Environmental Science and Technology 26: 313-319. Zhao, J.-L., G.-G. Ying, Y.-S. Liu, F. Chen, J.-F. Yang, & L. Wang. 2010. Occurrence and risks of triclosan and triclocarban in the Pearl River system, South China: From source to the receiving environment. Journal of Hazardous Materials 179: 215-222. Zhao, J.-L., Q.-Q. Zhang, F. Chen, L. Wang, G.-G. Ying, Y.-S. Liu, L.-J. Zhou, S. Liu, H.-C. Su, & R.-Q. Zhang. 2013. Evaluation of triclosan and triclocarban at river basin scale using monitoring and modeling tools: implications for controlling of urban domestic sewage discharge. Water Research 47: 395-405.

Chapter 2: Photolysis- and Dissolved Organic Matter-Induced Toxicity of Triclocarban to

Daphnia magna*

*formatted for Science Magazine

2.1 Introduction

Pharmaceuticals and personal care products (PPCPs) continue to be an emerging concern because their usage has increased dramatically over recent decades and little is known about their impact on the environment. Because many of these PPCPs enter receiving waters through wastewater treatment plant (WWTP) effluents and to a lesser degree from septic systems, understanding their environmental fate has gained interest due to their potential endocrine disrupting properties (Brooks and Huggett, 2009).

Triclocarban (TCC; 3,4,4’-trichlorocarbanilide) is an anti-microbial compound that is commonly found in many household products, including bar soaps, deodorants, detergents, and cosmetics (Chu and Metcalfe 2007). As use of these TCC-containing products has increased, so has its concentration in the environment, increasing the likelihood of organism exposure. Triclocarban is highly lipophilic, with a log Kow value of 4.2 (US EPA 2008), and has the potential to bioaccumulate in organisms. It has indeed been detected (i.e., found in the tissue or blood) in a variety of organisms including algae

(Coogan 2007), New Zealand mudsnails (Giuduce 2010), and fish (Schebb 2011), and its lethal and sublethal toxicity in fathead minnows (Pimephales promelas) has been noted

(Schultz 2012). 14

While the photolysis of triclosan (TCS), a similar antimicrobial compound, has been extensively studied because it can photodegrade into potentially toxic photoproducts, including chlorophenols and dioxin congeners (Latch 2003; Mezcua

2003; Latch 2005; Sanchez-Prado 2006; Chau 2008; Buth 2009; Buth 2010; Buth 2011;

Anger 2013), TCC has garnered much less attention. Guerard et al. (2009) and Trouts and

Chin (2015) reported significant indirect photolysis mediated by dissolved organic matter

(DOM), and only one study has identified photochemically-derived products after direct photolysis (Ding et al., 2013). Nonetheless, to date, there have been no studies that have examined the toxicity of TCC photoproducts on aquatic organisms. For these reasons, we wanted to investigate the photoinduced toxicity of TCC and the influence of DOM on the formation of photoproducts.

2.2 Methods/Results and Discussion

We investigated the lethal toxicity to Daphnia magna (US EPA Method 821-R-

12-012; US EPA 2002) of TCC and TCC photodegradation products formed in the presence and absence of DOM in US EPA moderately hard water. Initial experiments were conducted only with parent TCC to establish a baseline toxicity level. We then photolyzed solutions of TCC in the presence and absence of DOM until the parent compound was below the observed no mortality concentrations in the test solution by liquid chromatography-tandem mass spectrometry. The 96-hLC50 values (the concentration of chemical at which 50% mortality of the test population occurs over a specified time period) for photo-derivatives formed from the direct photolysis (no DOM)

15 of TCC increased; i.e., they became significantly less toxic relative to the parent compound (LC50 of 2.67 µM - 95% CL (confidence limit) 2.07-3.44) photolyzed vs.

0.087 µM (95% CL 0.04-0.19) for TCC) (Figure 2.1). In contrast, no difference in toxicity was observed between TCC solutions subjected to irradiation in the presence of

12 mg/L dissolved Suwannee River Natural Organic Matter (SRNOM, which is typically

50% carbon (giving 6 mg/L carbon, an environmentally relevant concentration) and represents the terrestrial endmember, i.e. the terrestrially-derived end of the range of

DOM compositions; LC50 values of 0.032 µM - 95% CL 0.017-0.060) and the TCC in the presence of SRNOM (0.146 µM 95% CL 0.050-0.434) (Figure 2.1).

Thus, the only significant effect on toxicity was observed when TCC was photolyzed in EPA water and toxicity decreased significantly. We believe that interaction between TCC and SRNOM during photolysis coupled with the formation of excited state

DOM species lead to the formation of degradation products that differ from those formed in the absence of DOM. Indeed we observed persistent toxicity even after the parent compound was below detection limits in the SRNOM solutions, while TCC solutions photolyzed in water were significantly less toxic. We believe that DOM in this case is capable of shielding the photoproducts through a combination of processes such as light screening and possible interactions with DOM moieties. Given that DOM is ubiquitous to all aquatic systems, it is likely that in sunlit natural surface waters, TCC photodegrades into these more toxic products that may be more persistent. There was no statistical difference between the toxicities of parent TCC in EPA water and parent TCC in EPA-

DOM water, indicating that there is not enough interaction between parent TCC and

DOM to significantly reduce the exposure of an organism to TCC, or that if there are strong interactions and an organism comes into contact with this TCC-DOM complex the

TCC can still exert its normal toxicity on the organism.

Ding et al (2013) reported four photolysis products for TCC: 4-chloroaniline, 3,4- dichloroaniline, 4-chlorophenyl isocyanate, and 3,4-dichlorophenyl isocyanate (log Kow values of 1.83, 2.69, 3.24, and 3.88, respectively (EPI Suite estimate)). Photolyzed TCC solutions, with and without DOM, were analyzed by gas chromatography-mass spectrometry (GC-MS) and all four compounds in the photolyzed solution containing

SRNOM were observed. However, none of the products were detected in the solutions which were photolyzed without DOM, thereby suggesting that DOM plays a role in both the formation and shielding of these toxic photoproducts while also making them more persistent in the environment. Based upon GC-MS analysis results it can be concluded that TCC is either rapidly degraded to unknown, less toxic photoproducts or that the characteristic photoproducts of TCC are further degraded to less toxic compounds.

To assess whether the formation of these toxic photoproducts is unique only to

SRNOM, the TCC photolysis experiments were repeated using DOM originating from different WWTP effluent waters collected from four different sites around central Ohio as well as a variety of DOM isolated by chromatographic methods, e.g., using XAD-8

(methylmelhacrylate) resins. The chloroaniline photoproducts were detected in the

SRNOM and wastewater effluent DOM, but not in the organic matter fractions isolated by chromatography. We attribute this to the XAD-8 isolation method, which likely removed the fraction of DOM responsible for the chloroaniline products being formed or

17 not further broken down, although this fraction is not known. In contrast, the chlorophenyl isocyanate products were found in all of the non-WWTP DOM samples

(Table 2.1). Due to the more complex composition of the WWTP effluent, as it contains

DOM, ions, and likely many other chemicals that we did not attempt to measure, we attribute the lack of chlorophenyl isocyante products to another effluent component that is likely either preventing the formation of these products or mediating their further breakdown.

To determine which of the four photoproducts are most likely responsible for the observed toxicity, we created de novo solutions with each of the four photoproducts and performed 96-h lethality toxicity tests on D. magna. Assuming additivity of toxic effects, the results indicated that 4-chloroaniline likely exerts the majority of the toxicity in tests with the photolyzed TCC (LC50 value of 0.082 µM (95% CL 0.034- 0.236), while 3,4- dichloroaniline (LC50 of 1.029 µM (95% CL 0.413-3.505) and the two chlorophenyl isocyanate compounds (4-CPI LC50 of >10 µM (calculated LC50 was extrapolated above highest test concentration); 3,4-DCPI LC50 of 2.030 µM (95% CL 1.337-2.979)) contributed less to toxicity (Figure 2.2). The low toxicities of 4-CPI and 3,4-CPI could be due to their lower solubilities. The two chloroanilines are more water soluble (4-CA

20.4mM (EPI Suite estimate); 3,4-DCA 2.08mM (EPI Suite estimate)) while the chlorophenyl isocyanates are often characterized as not very soluble (4-CPI 0.819 mM

(EPI Suite estimate); 3,4-DCPI 0.129 mM (EPI Suite estimate)), although at all concentrations used in the test, the compounds appeared to have fully dissolved, i.e. there was no visible particulates in the solution. This indicates that regardless of the true

18 toxicities of the chlorophenyl isocyanate products, they will likely be of less concern for organisms because of the low likelihood of organism exposure. 3,4-DCA will also be of less concern because of its low toxicity, even though it is more water soluble than the CPI compounds..

Finally, a mechanism was proposed to explain the mediation of DOM on the formation of toxic photoproducts from photolyzed TCC. The hypothesis is that DOM undergoes excitation under UV exposure, and then transfers an electron to TCC, forming a TCC radical anion. This radical anion can then lead to the rearrangement of the molecule and then cleavage at either of the bonds between the NH and C=O groups. This in turn will form either a chloroaniline radical or chlorophenyl isocyanate, depending on bond cleavage, and these radicals will undergo rearrangement to form the stable photolysis products (Figure 2.3).

The overarching conclusions of this study are that TCC is acutely toxic to D. magna and that the presence of DOM does not affect the toxicity of the parent compound.

When exposed to UV irradiation TCC in solution without DOM is detoxified due to the photodegradation of TCC to unknown, less toxic products. When photolyzed in solution with DOM, TCC breaks down to form 4-CA, 3,4-DCA, 4-CPI, and 3,4-DCPI, and together these products exert a toxicity that is statistically the same as parent TCC itself.

It is plausible that once these compounds are formed, DOM also then slightly shields the photoproducts from further photodegradation. While four photoproducts were detected, the LC50 of 4-CA is at least an order of magnitude lower than the three other

19 photoproducts and likely exerts the majority of the toxicity when TCC is photolyzed in the presence of DOM.

2.3 Supporting Information

2.3.1 Materials

Triclocarban (99%) was purchased from Sigma-Aldrich (St. Louis, MO).

NaHCO3 (99-100%), KCl (99.4%), MgSO4 (99%), and CaSO4 (98%) were purchased from Fisher Scientific (Waltham, MA). HPLC grade acetonitrile was purchased from

Fisher Scientific (Waltham, MA). Suwannee River Natural Organic Matter (SRNOM)

(RO Isolation), and Suwannee River Fulvic Acid (FA) (XAD8 isolation) were purchased from the International Humic Substances Society (St. Paul, MN). Two Old Woman Creek

FAs (XAD8 isolation), two Pony Lake, Antarctica FAs (XAD8 isolation), and Little

Miami River Humic Acid (HA) (XAD8 isolation) were collected by our lab.

2.3.2 Methods

2.3.2.1 EPA water

EPA water (US EPA 2002) was prepared by dissolving NaHCO3, KCl, and

MgSO4 in Milli-Q water (obtained from a Millipore Elix 10 Reverse Osmosis system connected to a Millipore Milli-Q UV Plus Ultra Pure Water System run at 18.2mOhms).

This water was shaken and allowed to aerate overnight. Into a separate container of Milli-

Q water we added CaSO4, which was allowed to mix overnight on a stirplate. The next

20 day these waters were then combined to achieve a final nominal concentration of 96 mg/L NaHCO3, 4 mg/L KCl, 60 mg/L MgSO4, and 60 mg/L CaSO4.

2.3.2.2 Half-life Determination

To determine the amount of time needed to degrade compounds for toxicity tests, degradation kinetics tests were used to determine the photolysis rates and half-lives of our compounds. We formulated 1µM solutions of the desired compound by dissolving a high concentration of the compound in methanol in a beaker, evaporating off the solvent, and adding EPA water to achieve the desired nominal concentration. The solutions were thoroughly mixed and then pipetted into quartz phototubes and exposed in the SunTest at

250 watts. Phototubes, representing each time point, were removed at various times and left in the dark until all tubes were removed. Irradiation was monitored with a VWR UV

Light Meter/Ultra-violet Radiometer. The solutions were then placed into HPLC vials and analyzed via HPLC/UV-vis. From the calculated concentrations of each time point we plotted ln C/Co vs. time and determined a degradation rate constant (kobs) for each compound. The kobs was then inserted into the formula:

-kt C=Coe (2.1)

Where C was the concentration at time t, Co was the initial concentration, and k was the degradation rate constant. To determine one half-life based on the degradation

1 rate constant, we set C= /2, Co=1 and solved for t.

2.3.2.3 HPLC Analysis

Analyses of water samples containing or TCC were performed using a Waters

1515 Isocratic HPLC pump and 717 Plus Autosampler with a Waters 2487 Dual Beam

Absorbance UV Detector. Water samples (50µL) were autoinjected then separated using a Restek Pinnacle DB C18 column (5µm particle size, 250 mm x 4.6 mm) at a flow rate of 0.8 mL/min. Mobile phase consisted of 70:30 Acetonitrile:Milli-Q water (v:v) adjusted to pH 3-4. Absorption was measured at 257nm (RT=9 min).

2.3.2.4 Daphnia magna Culture Maintenance

D. magna were purchased from Aquatic Biosystems, Inc. and maintained in EPA water until testing time. Organisms were fed a yeast/trout chow/algae (YTC/algae)

(personal communication with Aquatic Biosystems, Inc., Fort Collins, CO) and ground alfalfa (obtained from George Keeney, The Ohio State University). D. magna culture tanks were kept at 20 degrees C and 12hr:12hr light:dark photoperiod while under constant aeration with aeration stones and air pumps to ensure proper oxygenation of culture water.

2.3.2.5 Acute Toxicity Testing

Acute (96-h) toxicity tests were performed using six concentrations, spanning five orders of magnitude (i.e., 10µM – 0.0001µM), of the compounds dissolved in EPA water or EPA water containing SRNOM. Test solution (25g) was added to 150 mL beakers, and five D. magna less than 24 h old were gently pipetted from the culture tank into each beaker. Test beakers containing D. magna were left in the dark during the duration of the

22 tests due to the photosensitivity of the compounds. After 48 h the D. magna were fed for two hours, then live organisms were gently removed from the beakers and allowed to remain in a small amount of solution while waters were renewed. After renewal, D. magna were replaced into the test chambers and monitored for another 48 h (96 h total).

Mortality was monitored and recorded every 24 h, and 96 hr-LC50 values were calculated using Solver in Microsoft Excel.

2.3.2.6 Wastewater Treatment Plant Effluent Collection/Photolysis

Wastewater treatment plant (WWTP) effluent was collected from four central

Ohio WWTPs: two Columbus urban WWTPS, Jackson Pike WWTP and Southerly

WWTP, and from two rural WWTPS, London WWTP and Plain City WWTP. Effluent was collected directly from the plant, prior to release into surface waters. Dissolved organic carbon (DOC) concentrations were measured using a Shimadzu TOC-Vcpn total organic carbon analyzer. These WWTP waters were spiked with 10µM TCC, placed in quartz phototubes, and irradiated for 150 h, or about five half-lives. Photolyzed samples were then analyzed via GC-MS, as described below, to determine which photoproducts were present.

2.3.2.7 Effects of DOM Composition

To determine which photoproducts were formed with various NOMs, FAs, and

HAs, seven different DOMs (Little Miami River FA, Suwannee River FA, two Old

Woman Creek FAs, two Pony Lake FAs, Old Woman Creek NOM) were added to EPA

23 water at concentrations of 12 mg DOM/L, giving DOC concentrations of around 4-6 mg/L, similar to what would be expected in natural aquatic environments. Actual DOC concentrations were measured using a Shimadzu TOC-Vcpn total organic carbon analyzer. These solutions were spiked with 10µM TCC, placed in quartz phototubes, and irradiated for 150 h, or about five half-lives, to ensure TCC was degraded to below a concentration that would exert acute toxicity. Photolyzed samples were then analyzed via

GC-MS to determine which photoproducts were present.

2.3.2.8 GC/MS Analysis

GC-MS analysis was performed to confirm the presence of the four reported photoproducts and to confirm that there was no residual parent TCC remaining.

Chloroaniline and chlorophenyl isocyanate standards (10 uM) were prepared by directly adding the solid compound to Milli-Q water and vortexing until dissolved. Standard solutions were then diluted 10-fold with Milli-Q water and 500 uL of this solution was then added to 500 uL CH2Cl2 in a glass tube, vortexed for 20 s, left to sit in the dark for

10 min. The vortexing/sitting in the dark process was repeated, and then 1 uL of the bottom layer of this solution was directly injected into the GC-MS. Triclocarban samples

(500 uL) were added to 500 uL of CH2Cl2, vortexed for 20 s, then left to sit in the dark for 20 s. This process was repeated, then 350 uL of the bottom layer of this solution was transferred to a new glass tube, dried completely under N2, and reconstituted in 35 uL of

CH2Cl2. One uL of this solution was then directly injected for GC-MS analysis on a

Thermo Trace GC Ultra gas-chromatograph and Thermo DSQII mass spectrometer. The

GC-MS needle was washed with CH2Cl2 followed by MeOH. The initial hold was at

40°C for 2 min, then increased to 300°C at a rate of 10°C/min, and then held at 300°C for

2 min. The ion source temperature was 250°C, using an ES+ ionization method and a full scan m/z of 10-750. Split injection mode was used, with a split flow of 15 mL/min, a split ratio of 10, inlet temperature of 225°C, and a transfer line temperature of 300°C.

2.3.2.9 Data analysis

Dose-response curves were analyzed using TCS, TCC, and photoproduct concentrations and were fitted by the log-logistic model (Doelman and Haanstra, 1989) using the Newton optimization method (Solver, Microsoft Excel, 2007). LC50 values were considered different if there was no 95% confidence interval overlap and if confidence limits did overlap, differences between LC50 values in different treatments were calculated using a single-sided student-t-test (p=0.05).

2.3.3 Supplemental Experiments

2.3.3.1 No Mortality Concentrations

To ensure that the photolyzed samples no longer contained enough parent compound to exert toxicity, we used data from the parent compound acute toxicity tests to determine which of the test concentrations resulted in no mortality to D. magna. This concentration for TCC was 0.001 µM.

2.3.3.2 Determination of Irradiation Time

The half-life of TCC was <60 h at 250 W. At 500 W we would expect the half-life to be <30 h. Using the no observed mortality concentration for TCC (0.001 µM) and

Formula 2.1, we calculated that TCC should be irradiated for at least 10 half-lives to reduce the initial concentration of 10 µM to a concentration where not enough residual parent compound would remain to exert toxic effects and skew the toxicity data of the photolysis product tests. This required TCC to be irradiated for at least 300 hours.

2.3.3.3 Toxicity of SRNOM

To ensure that the SRNOM and irradiated SRNOM were not inducing toxicity, a control study was performed in EPA water without the addition of TCC. The three controls consisted of (1) EPA water alone, (2) EPA water containing 12 mg/L SRNOM, and (3) Irradiated EPA water containing 12 mg/L SRNOM. Each control test consisted of four beakers, with each beaker containing five D. magna. Test procedures were followed as previously described, following EPA method 821-R-12-012. There was no mortality observed in any of the control tests (Table E.1). This demonstrates that any toxic effects seen in the acute toxicity tests were the result of parent compounds and/or photolysis products, and not the SRNOM itself.

2.3.3.4 Residual Parent Triclocarban

Mass spectra showed that there was no TCC remaining in the photolyzed solution

(mw = 315.58 g/mole), indicating that any toxic effects seen with these irradiated

26 solutions were solely due to the photolysis products, and not any residual parent TCC in the solutions.

2.4 Tables

Table 2.1 Photolysis products of TCC in different DOM solutions. Detection of the four photoproducts of triclocarban photolyzed in the presence of DOM in four wastewater treatment plant waters and seven different DOM solutions. Either 4- chloraniline (4-CA) and 3,4-dichloroaniline (3,4-DCA), 4-chlorophenyl isocyanate (4- CPI) and 3,4-dichlorophenyl isocyanate (3,4-DCPI), or all four, were detected in all of the solutions, regardless of DOM type (WWTPE=wastewater treatment plant effluent, HA=humic acid, FA=fulvic acid, NOM=natural organic matter) or DOM derivation. XAD8=a methylmelhacrylate resin.

DOM TOC 4- 3,4- 4- 3,4- Water DOM Type Derivation Isolation (mg/L) CA DCA CPI DCPI Method Allochthanous/ Jackson Pike WWTPE None 5.34 Y Y N N Authochthanous Allochthanous/ Southerly WWTPE None 4.81 Y Y N N Authochthanous Allochthanous/ London WWTPE None 4.76 Y Y N N Authochthanous Allochthanous/ Plain City WWTPE None 3.25 Y Y N N Authochthanous Little Miami Allochthanous/ HA XAD8 0.29 N N Y Y River Authochthanous Old Woman Allochthanous/ FA XAD8 7.28 N N Y Y Creek Authochthanous Old Woman Allochthanous/ FA XAD8 6.14 N N Y Y Creek Authochthanous Pony Lake FA Authochthanous XAD8 5.54 N N Y Y

Pony Lake FA Authochthanous XAD8 5.73 N N Y Y Suwannee FA Allochthanous XAD8 4.79 N N Y Y River Suwannee NOM Allochthanous RO 6.00 Y Y Y Y River Suwannee NOM Allochthanous RO 6.00 Y Y Y Y River Suwannee NOM Allochthanous RO 6.00 Y Y Y Y River EPA None N/A N/A 0.00 N N N N

2.5 Figures

3.5 *

2.5 (µM)

50 2 EPA

hr LC DOM

- 1.5 96 1

0.5

0 TCC Photolyzed TCC

Figure 2.1 Toxicity of parent and photolyzed TCC with and without DOM. 96-hr LC50 values (shown with 95% CI error bars) for triclocarban, both parent and photolyzed, with and without DOM. Triclocarban without DOM was less toxic after photolysis than TCC in US EPA water without photolysis (LC50s of 0.087 µM and 2.67 µM, respectively). In the presence of SRNOM, toxicity was the same as TCC in US EPA water regardless of the presence of DOM or photolysis (LC50s of 0.146 µM and 0.032 µM, respectively). Asterisk (*) indicates significant difference from all other LC50 values (p ≤ 0.05).

Figure 2.2 Toxicity of TCC photolysis products. 96-hr LC50 values (shown with 95% CI error bars) for 4-chloraniline (4-CA), 3,4-dichloroaniline (3,4-DCA), and 3,4- dichlorophenyl isocyanate (3,4-DCPI) in EPA water. The LC50 values for each compound were 0.082 µM, 1.029 µM, and 2.030 µM, respectively. The LC50 value for 4- chlorophenyl isocyanate was greater than the highest concentration tested (> 10 uM) is not included in the figure. Asterisk (*) indicates significant difference from all other LC50 values (p ≤ 0.05).

Figure 2.3 Proposed mechanism of TCC photolysis with DOM. Proposed mechanism of the DOM-mediated photolysis of triclocarban to 4-chloroaniline, 3,4-dichloroaniline, 4-chlorophenyl isocyanate, and 3,4-dichlorophenyl isocyanate. TCC is transferred an electron from UV-irradiation-excited DOM, forming a radical anion that cleaves between the C=O and either NH group, forming a set of chloroaniline and chlorophenyl isocyante radicals, which will undergo rearrangement to form stable molecules.

2.6 References

Anger, C.T., C. Sueper, D.J. Blumentritt, K. McNeill, D.R. Engstrom, & W.A. Arnold. 2013. Quantification of triclosan, chlorinated triclosan derivatives, and their dioxin photoproducts in lacustrine sediment cores. Environmental Science and Technology 47: 1833-1843. Brooks, B.W., & D.B. Huggett. 2009. Pharmaceuticals and personal care products: Research needs for the next decade. Environmental Science and Technology 28: 2469-2472. Buth, J.M., M. Granbois, P.J. Vikesland, K. McNeill, & W.A. Arnold. 2009. Aquatic photochemistry of chlorinated triclosan derivatives: potential source of pholychlorodibenzo-p-dioxins. Environmental Toxicology and Chemistry 28: 2555-2563. Buth, J.M., P.O. Steen, C. Sueper, D. Blumentritt, P.J. Vikesland, W.A. Arnold, & K. McNeill. 2010. Dioxin photoproducts of triclosan and its chlorinated derivatives in sediment cores. Environmental Science and Technology 44: 4545-4551. Buth, J.M., M.R. Ross, K. McNeill, & W.A. Arnold. 2011. Removal and formation of chlorinated triclosan derivatives in wastewater treatment plans using chlorine and UV disinfection. Chemosphere. 84: 1238-1243. Chau, W.C., J. Wu, & Z. Cai. 2008. Investigation of levels and fate of triclosan in environmental waters from the analysis of gas chromatography coupled with ion trap mass spectrometry. Chemosphere 73: S13-S17. Chu, S., & C.D. Metcalfe. 2007. Simultaneous determination of triclocarban and triclosan in municipal biosolids by liquid chromatography tandem mass spectrometry. Journal of Chromatography A 1164: 212-218. Coogan, M.A., R.E. Edziyie, T.W. La Point, & B.J. Venables. 2007. Algal bioaccumulation of triclocarban, triclosan, and methyl-triclosan in a North Texas wastewater treatment plant receiving stream. Chemosphere 67: 1911-1918. Ding, S.-L., X.-K. Wang, W.-Q. Jiang, X. Meng, R.-S. Zhao, C. Wang, & X. Wang. 2013. Photodegradation of the antimicrobial triclocarban in aqueous systems under ultraviolet radiation. Environmental Science and Pollution Research 20: 3195-3201.

Doelman, P., & L. Haanstra. 1989. Short- and long-term effects of heavy metals on phosphatase activity in soils: an ecological dose-response model approach. Biology and Fertility of Soils 8: 235-241. Giuduce, B.D. & T.M. Young. 2010. The antimicrobial triclocarban stimulates embryo production in the freshwater mudsnail Potamopyrgus antipodarum. Environmental Toxicology and Chemistry 29: 966-970. Guerard, J.J., P.L. Miller, T.D. Trouts, & Y.-P. Chin. 2009. The role of fulvic acid composition in the photosensitized degradation of aquatic contaminants. Aquatic Sciences 71: 160-169. Latch, D.E., J.L. Packer, W.A. Arnold, & K. McNeill. 2003. Photochemical conversion of triclosan to 2,8-dichlorodibenzo-p-dioxin in aqueous solution. Journal of Photochemistry and Photobiology A 158: 63-66. Latch, D.E., J.L. Packer, B.L. Stenden, J. VanOverberke, W.A. Arnold, & K. McNeill. 2005. Aqueous photochemistry of triclosan: formation of 2,4-dichlorophenol, 2,8- dichlorodibenzo-p-dioxin, and oligomerization products. Environmental Toxicology and Chemistry 24: 517-525. Mezcua, M., M.J. Gómez, I. Ferrer, A. Aquera, M.D. Hernando, A.R. Fernández-Alba. 2004. Evidence of 2,7/2,8-dichlorodibenzo-p-dioxin as a photodegradation product of triclosan in water and wastewater samples. Analytica Chimica Acta 524: 241-257. Sanchez-Prado, L., M. Llompart, M. Lores, C. García-Jares, J.M. Bayona & R. Cela. 2006. Monitoring the photochemical degradation of triclosan in wastewater by UV light and sunlight using solid-phase microextraction. Chemosphere 65: 1338- 1347. Schebb, N.H., I. Flores, T. Kurobe, B. Franze, A. Ranganathan, B.D. Hammock & S.J. Teh. 2011. Bioconcentration, metabolism and excretion of triclocarban in larval Qurt medaka (Oryzias latipes). Aquatic Toxicology 105: 448-454. Schultz, M.M., S.E. Bartell, & H.L. Schoenfuss. 2012. Effects of triclosan and triclocarban, two ubiquitous environmental contaminants, on anatomy, physiology, and behavior of the Fathead Minnow (Pimephales promelas). Archives of Environmental Contaminants and Toxicology 63: 114-124. Trouts, T., & Y.-P. Chin. 2015. Direct and indirect photolysis of triclocarban in the presence of dissolved organic matter. Elementa: Science of the Anthropocene. US EPA United States Environmental Protection Agency. 2002. Methods for measuring the acute toxicity of effluents and receiving waters to freshwater and marine organisms. Fifth Edition. EPA-821-R-02-012. 33

Chapter 3: Mixture toxicities of triclosan, triclocarban, and their photolysis products to Daphnia magna

3.1 Introduction

Triclosan (TCS; 5-chloro-2-(2,4-dichlorophenoxy)phenol) and triclocarban (TCC;

3-(4-chlorophenyl)-1-(3,4-dichlorophenyl)urea) are two ubiquitous antimicrobial compounds that are found in many common household items, including soaps, detergents, and toothpastes. The environmental fate of TCS has been studied widely and extensively and it is known to photolyze to dioxins and chlorophenols, among other products, in surfaces of freshwater systems (Tixier et al., 2002; Latch et al., 2003;

Mezcua et al., 2004; Lores et al., 2005; Latch et al., 2005; Buth et al., 2009; Buth et al.,

2010; Buth et al., 2011) and in seawater (Aranami et al., 2007). If TCS is further chlorinated in wastewater treatment plants (WWTPs) before photolysis, it can also form more highly chlorinated congeners of dioxin, such as 2,3,7-trichlorodibenzodioxin and

1,2,3,8-tetrachlorodibenzodioxin (Buth et al., 2011). As it is commonly believed that a higher degree of chlorination of dioxins, up to a certain level, makes the dioxin more toxic, these congeners are likely more toxic than the 2,8-DCDD species. Since TCS is recalcitrant to degradation in WWTPs, it has been found in many WWTP effluents entering freshwater systems, and due to its high log Kow of 4.8-5.4 (Ying et al., 2007; Son et al., 2009), it has a tendency to accumulate in sediments (Kolpin et al., 2002; Singer et al., 2002; Buth et al., 2011; Venkatesan et al., 2012; Gautum et al., 2014; Anger et al., 34

2015). Further, triclosan has also been found to accumulate in the tissues of marine mammals, including dolphins (Fair et al., 2009), orcas (Bennett et al., 2009), and polar bears (Sacco and James, 2005).

The environmental fate of TCC has not been studied in nearly as much detail as

TCS, but it has also been found in numerous freshwater systems (Halden et al., 2004), with a few studies noting the photolysis of TCC in water (Guerard et al., 2009; Ding et al., 2013; Ding et al., 2015; Trouts and Chin, 2015) and the formation of toxic photoproducts (Chapter 2). Triclocarban also has a moderately high log Kow (4.2, US

EPA 2008) and has the potential to accumulate in sediments or organisms.

In aquatic organisms, the acute toxicity of TCS has been determined (Orvos et al.,

2002; Smith et al., 2005; DeLorenzo et al., 2008; Harada et al., 2008; Oliveira et al.,

2009; Kim et al., 2009; Tamura et al., 2013) and chronic toxicity has been observed in the form of endocrine disrupting effects (Ishibashi et al., 2004; Raut et al., 2010). Although the toxicity of TCC has not been studied as extensively, it has been observed to exert toxic effects on aquatic and terrestrial organisms (Snyder et al., 2011; Tamura et al.,

2013; Gao et al., 2015). In studies comparing the toxicities of TCS and TCC, it was found that depending on the effect studied, TCC can be either more toxic or less toxic than

TCS. When studying the effects of TCS and TCC on the growth inhibition of the algae T. thermophila, Gao et al. (2015) found that TCC is about three times as toxic as TCS, with

24-h EC50s of 0.935 µM and 3.671 µM, respectively. Another study found that TCS was more toxic than TCC on the inhibition of the algae P. subcapitata growth, with 72-h EC50 values of 0.018 µM and 0.092 µM, respectively (Tamura et al., 2013). However, this

group also reported that TCC was more toxic than TCS when studying D. magna immobilization (48-h EC50 values of 0.032 µM and 0.622 µM, respectively) and Japanese medaka fish O. latipes juvenile mortality (96-h LC50 values of 0.269 µM and 0.725 µM, respectively) (Tamura et al., 2013). Chronic toxicity tests also reported different relative toxicities of TCS and TCC depending on test organism species used. The no observed effect concentrations (NOECs) of TCS and TCC to 72-h P. subcapitata growth were

0.002 µM and 0.018 µM, respectively (Tamura et al., 2013). Ceriodaphnia dubia reproduction was also monitored with 8-day NOECs of 0.104 µM TCS and 0.005 µM

TCC, as was hatching and hatchling survival of the zebrafish D. rerio, with 9-day

NOECs of 0.090 µM TCS and 0.076 µM TCC (Tamura et al., 2013).

Since TCS and TCC toxicity has been observed in ecological receptors and they are often found concurrently in aquatic receiving waters, WWTP effluent, sediments, suspended sediments, sewage sludge, and biosolids (Coogan et al., 2005; Halden and

Paul, 2005; Coogan et al., 2008; Miller et al., 2008; Zhao et al., 2010; Xia et al., 2010;

Yu et al., 2011; Zhao et al., 2013; Lozano et al., 2013; Pycke et al., 2014; Gautam et al.,

2014; Lv et al., 2015), it is important to understand the nature of joint toxicity of these compounds. Toxicity of organic chemical mixtures to aquatic organisms may be enhanced or synergistic (Walter et al., 2002; Cleuvers et al., 2004), including some mixtures of TCS and other organic compounds (DeLorenzo and Fleming, 2008). More commonly, mixture toxicity is simply additive (Hermes et al., 1985; Backhaus et al.,

2000; Fraker and Smith, 2004; Christensen et al., 2007; Knauert et al., 2008). Several studies have examined the mixture toxicity of TCS and TCC, with most results indicating

36 additive or enhanced effects (Yang et al., 2008; Schultz et al., 2012; Villa et al., 2014;

Zenobio et al., 2014). To date the effects of various mixtures of the TCS and TCC and their photolysis products have not been studied. Our previous research has indicated that

TCC remains just as toxic after photolysis in the presence of dissolved organic matter

(DOM), and as such it is important to understand the nature of effects in mixtures of these compounds.

In this study we investigated the toxicity of parent and photolyzed TCS with and without DOM present to determine baseline toxicities for comparison to the mixtures.

The main goal of this study was to then determine the nature of mixture effects of these compounds and their photoproducts to Daphnia magna in order to better understand the hazard posed by both compounds in surface waters. We were primarily interested in mixtures including TCC that had been photolyzed in solution with DOM due to the observed formation of highly toxic photoproducts as discussed in Chapter 2.

3.2 Materials and Methods

3.2.1 Materials

Triclocarban (99%) and Triclosan (≥97%, as Irgasan) were purchased from

Sigma-Aldrich (St. Louis, MO). Suwannee River Natural Organic Matter (RO isolation) was purchased from the International Humic Substances Society (St. Paul, MN).

NaHCO3 (99-100%), KCl (99.4%), CaSO4 (98%), and MgSO4 (99%) were purchased from Fisher Scientific (Hampton, NH). Daphnia magna and yeast/trout chow/algae mix

(YTC) were purchased from Aquatic Biosystems, Inc. (Ft. Collins, CO). HPLC grade acetonitrile was purchased from Fisher Scientific (Waltham, MA).

3.2.2 Test water

EPA moderately hard standard synthetic water was prepared according to standard methods (US EPA 2002). Briefly, NaHCO3, KCl, and MgSO4 were dissolved in

18 L of Milli-Q water (Millipore Elix 10 Reverse Osmosis system connected to a

Millipore Milli-Q UV Plus Ultra Pure Water System, 18.2 MΩs). This water was shaken and allowed to aerate overnight. Into a separate 1-L container of Milli-Q water we dissolved CaSO4which was allowed to mix overnight. The next day these waters were then combined to achieve a final nominal concentration of 96 mg/L NaHCO3, 4 mg/L

KCl, 60 mg/L MgSO4, and 60 mg/L CaSO4.

3.2.3 Maintenance of test organisms

D. magna were acclimated to laboratory conditions for at least 48 hours in EPA moderately hard water prior to toxicity tests and were fed a YTC/algae food and ground alfalfa (received from George Keeney, The Ohio State University, Columbus, OH). D. magna culture tanks were kept at 20 degrees C and 12hr:12hr light:dark photoperiod while under constant aeration with aeration stones and air pumps to ensure proper oxygenation of culture water.

3.2.4 Acute toxicity testing with Daphnia magna

Parent TCS tests were conducted initially as described in this paragraph, and from those results we observed a concentration at which there was no mortality. This concentration was used to determine how long to photolyze the TCS solutions (with and without DOM) so that residual parent TCS would not exert toxic effects. The photolysis procedures for this are described in the following section. Acute (96-h) toxicity tests were conducted using US EPA moderately hard standard synthetic water (US EPA Method

821-R-12-012; US EPA 2002). Triclosan was dissolved in EPA water or EPA water containing 12mg/L SRNOM (50% carbon, giving an environmentally relevant dissolved organic carbon concentration) and either left as the parent compound in solution (no photolysis) or photolyzed before beginning toxicity tests. Test solutions (25 g) were then added to 150-mL test chamber beakers. For initial TCS and photolysis product testing we used six concentrations spanning five orders of magnitude (i.e., 10 µM – 0.0001 µM). At each concentration we used four replicates, each containing five Daphnia magna less than 24 hours old (20 D. magna per concentration). D. magna were gently pipetted from the culture tank into each beaker. Test beakers containing D. magna were left in the dark during the duration of the tests due to the photosensitivity of the compounds. D. magna were monitored for 96 h; feeding occurred at 48 hours and lasted two hours (US EPA

Method 821-R-12-012; US EPA 2002). Feed solution was pipetted into test beakers, and after the feeding period live organisms were gently removed from the beakers and allowed to remain in a small amount of solution while waters were renewed. After renewal, D. magna were replaced into the test chambers and monitored for another 48 hours (96 hours total). Mortality was recorded at each concentration after 24, 48, 72, and

96 hours. We used this data to determine the concentrations at which we observed no mortality and combined this with the degradation rate data described below to determine the appropriate photolysis time for TCS.

3.2.5 Determination of photolysis times

Photolysis degradation rate constants (kobs) and half-lives (t1/2) of TCS were determined so as to calculate the appropriate amount of photolysis time required to degrade TCS to residual concentrations below the no observed mortality concentrations.

A nominal 1 µM solution of TCS was prepared by adding 17.5 µL of a high- concentration methanol stock solution to a beaker, evaporating off the solvent, and adding 70 mLs of EPA water with or without 12 mg/L SRNOM. Solutions were thoroughly mixed and then placed into quartz phototubes with a 1-cm path length, capped, and exposed at 250 watts in an Atlas SunTest CPS+ (Atlas Material Testing

Technology, Illinois, USA), which uses a Xenon light source with a filter (Filter Type H,

Atlas Material Testing Technology) to simulate natural sunlight. The temperature inside the solar simulator was held constant at room temperature (25° C) by a cooling system, and irradiation was monitored with a VWR UV light meter/ultra violet radiometer to confirm the consistency of irradiation levels. Phototubes, representing each time point, were removed at various times, with sequences ranging from one minute to one hour, to one hour to over six days, and left in the dark until all tubes were removed. A time zero sample was placed into an amber vial and stored in the dark while the other samples were irradiated, and a dark control was placed in an amber vial, covered in foil, and left in the

40 solar simulator until the last phototube was removed. Irradiation was monitored with a

VWR UV Light Meter/Ultra-violet Radiometer. The solutions were then placed into

HPLC vials and analyzed via HPLC/UV-vis on a Waters 1515 Isocratic HPLC pump and

717 Plus Autosampler with a Waters 2487 Dual Beam Absorbance UV Detector. A 50

µL aliquot of each water sample was autoinjected then separated using a Restek Pinnacle

DB C18 column (5 µm particle size, 250 mm x 4.6 mm) at a flow rate of 0.8 mL/min.

The mobile phase consisted of 70:30 Acetonitrile:Milli-Q water (v:v) adjusted to pH 3-4.

Absorption was measured at 280nm (RT=10.5min).

From the calculated concentrations of each time point we plotted ln C/Co vs. time and determined a degradation rate constant (kobs) for each compound. The kobs was then inserted into the formula:

-kt C=Coe (3.1)

Where C was the concentration at time t, Co was the initial concentration, and k was the degradation rate constant. To determine one half-life based on the degradation

1 rate constant, we set C= /2, Co=1, and solved for t.

3.2.6 Calculation of toxic units

Mixture concentrations of each compound were calculated using toxic unit equivalents. LC50 values of various compounds were estimated from 96-h lethality tests and were subsequently used to develop the mixture dosing ranges. At 1 toxic unit (TU), we would expect to see 50% mortality of a test population of organisms, so this would be the LC50 concentration. A range of proportional TU concentrations was established as the

41 exposure concentration series using appropriate concentrations of each compound or photolysis product. As an example for TCS,

LC50 = 1 TU (i.e., the LC50 of parent TCS in EPA water was 1.80 µM, so 1 TU of TCS in

EPA water is 1.80 µM, 2 TUs is 3.60 µM)

Therefore:

0.5 LC50 compound A + 0.5 LC50 compound B = 1 LC50 = 1 TU (3.2)

Where 50% mortality would be expected if toxicity is additive

1 LC50 compound A + 1 LC50 compound B = 2 LC50s = 2 TUs (3.3)

0.25 LC50 compound A + 0.25 LC50 compound B = 0.5 LC50s = 0.5 TUs (3.4)

0.05 LC50 compounds A + 0.05 LC50 compound B = 0.1 LC50 = 0.1 TUs (3.5)

0.025 LC50 compound A + LC50 compound B = 0.05 LC50 = 0.05 TUs (3.6)

Solutions were prepared with two-fold the desired final concentration of individual compounds then added together to achieve final concentrations of appropriate

TUs, with the exception of the solution in which TCS and TCC were photolyzed in the same solution.

3.2.7 Mixture Toxicity Testing

Mixture toxicity tests were performed as described above using binary mixtures of TCS, TCC, and their photolysis products (exposed in either 0 or 12 mg/L SRNOM.

Tests were conducted using 0.05, 0.1, 0.5, 1, and 2 TU concentrations and a control of

US EPA moderately hard water, and a TU50 was calculated for each mixture after exposing D. magna to the solution for 96 h. TU50s of the various mixtures were then compared to 1 TU50 to determine any synergistic/antagonistic effects. Mixtures tests were repeated as many times as possible based on the amount of test solution available with three tests conducted with the mixture of both TCC and TCS, three tests with for TCS and photolyzed TCC in DOM, and four tests with TCC and photolyzed TCC in DOM.

Due to low amounts of photolyzed TCS in EPA water, photolyzed TCS in DOM, and the solution photolyzed containing both compounds there was only one test performed for each of the relevant mixtures. It should be noted that there were still four replicates containing five D. magna at each concentration in all of these tests. One set of results was selected for each test to use in statistical analysis and discussion.

3.2.8 Data analysis

Dose-response curves were analyzed using TCS, TCC, and photoproduct concentrations and were fitted by the log-logistic model (Doelman and Haanstra, 1989) using the Newton optimization method (Solver, Microsoft Excel, 2007). If there was no overlap between the 95% CL and a TU50 of 1, mixtures were considered synergistic or antagonistic. If 95% CL overlapped with a TU50 of 1 mixtures were considered additive.

3.3 Results

3.3.1 Effects of photolysis on triclosan toxicity

The toxicity of TCS to D. magna in treatments without and with SRNOM was not significantly different (1.620 µM (95% CL 0.920-2.864) and 1.760 µM (95% CL 1.174-

2.652), respectively). The toxicities of TCS when photolyzed without or with SRNOM were also not significantly different from one another (8.482 µM (95% CL 6.681-10.77) and 8.513 µM (95% CL 2.700-26.84), respectively). However, the apparent toxicity of

TCS was reduced by photolysis regardless of presence of organic matter (Figure 3.1).

While both parent tests were not significantly different from one another and both photolyzed tests were not statistically different from one another, TCS photolyzed in

EPA water was statistically different from both parent tests. However, TCS photolyzed in

EPA water containing DOM was not statistically different from parent TCS in EPA water

(Figure 3.1), but was different from parent TCS in EPA water containing DOM (Figure

3.1).

3.3.2 Effects of mixtures

Of the six binary mixtures investigated, three appeared to have synergistic effects and the other three additive effects (Figure 3.2). A TU50 below 1 indicates that it is taking less than a 1 TU to kill 50% of the population, indicating that the two compounds are more toxic to D. magna in mixtures than when exposed independently. The mixture of

TCS+TCC (TU50 of 0.0246, 95% CL 0.0064- 0.6909) was about an order of magnitude more toxic than the mixtures of photolyzed TCC in DOM + TCC (TU50 of 0.4750, 95%

CL 0.3401- 0.6641), and photolyzed TCC in DOM + TCS (TU50 of 0.5298, 95% CL

0.4063- 0.6909), all of which displayed synergistic toxicities. These three mixtures were

44 all significantly different from a TU50 of 1 due to their 95% confidence intervals not overlapping with a TU50 of 1 (Figure 3.2). The remaining three mixtures of photolyzed

TCC in DOM + photolyzed TCS in DOM (TU50 of 1.450, 95% CL 0.5344- 3.924), both

TCS and TCC photolyzed simultaneously in the same solution (TU50 of 0.9870, 95% CL

0.3832- 2.541), and photolyzed TCC in DOM + photolyzed TCS in EPA (TU50 of 0.6790,

95% CL 0.2330- 1.981) were all statistically the same as a TU50 of 1, as their 95% confidence intervals overlapped with a TU50 of 1 (Figure 3.2).

3.4 Discussion

Toxicity tests with TCS and photolyzed TCS indicated that TCS not subjected to photolysis, regardless of the presence of SRNOM, was significantly more toxic than any photo-derivatives. Although TCS forms 2,8-dichlorodibenzo-p-dioxin and 2,4- chlorophenol after photolysis (Latch et al., 2003; Latch et al., 2005), these products are likely not present at high enough concentrations to exert as much acute toxicity as the parent compounds, but may still possibly impart sub-lethal effects. This would suggest that TCS, not its photolysis products, are of greater environmental concern when considering acute toxicity. However, in the context of chronic toxicity, TCS can accumulate in sediments, especially near wastewater treatment plants (Buth et al., 2010;

Zhao et al., 2013). We suspect that any “legacy” TCS desorbed from the sediment organic matter could accumulate in pore waters and pose a threat to benthic organisms.

The toxic effects, chronic exposure, and bioaccumulation potentials in a case like this are unknown, and should be investigated. In an opposing case, however, where TCS did not

45 sorb to sediments and remained in surface waters, we would expect to see toxic effects on aquatic populations due to the demonstrated acute toxicity (Figure 3.1), bioaccumulation potential (Coogan et al., 2007; Coogan et al., 2008), and estrogenicity (Ishibashi et al.,

2004; Raut et al., 2010) of parent TCS.

In natural freshwater systems, both TCS and TCC are found ubiquitously and concurrently (Halden and Paul, 2005; Coogan et al., 2007; Coogan et al., 2008; Miller et al., 2008; Zhao et al., 2010; Xia et al., 2010; Yu et al., 2011; Lozano et al., 2013; Zhao et al., 2013; Pycke et al., 2014; Gautam et al., 2014; Lv et al., 2015), increasing the likelihood of exposure of aquatic organisms to mixtures of these compounds. This is also of concern due to the synergisms seen with the mixtures of TCC+TCS, photolyzed TCC in DOM + TCS, and photolyzed TCC in DOM +TCC. The most toxic mixture of

TCS+TCC, unphotolyzed and without DOM, was approximately an order of magnitude more toxic than any other mixtures, suggesting that when combined, this mixture could be much more detrimental than previously thought. It cannot yet be determined if TCS and TCC act via the same mechanism that causes toxicity to organisms, or whether they act through different pathways, whereby one compound makes the organism more susceptible to the effects of the other. Understanding the reasons for increased toxicities seen with the photolyzed compounds becomes even more complex. As previously mentioned, both TCC (Chapter 2) and TCS (Tixier et al., 2002; Latch et al., 2003;

Mezcua et al., 2004; Latch et al., 2005; Lores et al., 2005; Buth et al., 2009; Buth et al.,

2010; Buth et al., 2011) photolyzed to multiple degradation products in freshwater, making it extremely difficult to determine which of the products are active participants in

46 the mixture toxicity enhancement (Table 1), although the individual toxic effects of the photolysis products of TCC were reported previously (Chapter 2).

To understand how compounds interact with each other and with organisms in mixtures, i.e. to understand why synergistic and additive effects were seen with the various mixtures, it is valuable to know the mechanisms of action of each of the individual compounds. The mechanism of action of TCS, and possibly TCC, although less is known about this, is the inhibition of fatty acid synthesis in bacteria (McMurry et al., 1998; Levy et al., 1999). Dioxins and dioxin-like compounds (2,8-DCDD is a photoproduct of TCS) interact with the aryl hydrocarbon (AH) receptor in organisms, which is involved in transcription of DNA (Poellinger 2000; Mandal 2005; Linden et al.,

2010). 4-Chloroanline exerts toxicity by causing hemolytic processes, leading to destruction of erythrocytes, or red blood cells, and causing anemia, among other issues

(Beard and Noe, 1981; Beutler 1985; Chhabra et al., 1990). Numerous studies have investigated the physiology of the toxicity of chlorophenols (photoproduct of TCS), suggesting that chlorophenols exert toxicity by inducing oxidative stress and DNA damage (Bukowska 2003; Truffin et al., 2003; Bukowska 2004; Dai et al., 2005), reducing the number of cytochrome P450 enzymes (Bukowska 2003; Bukowska 2004), and altering androgen induced transcriptional activity (Kim et al., 2005). Chlorophenol toxicity also appears to increase with the degree of chlorination (Chen et al, 2004). These various mechanisms of action demonstrate the complexity in determining precisely why synergistic or additive effects were seen in our results, as the physiology of each individual compound is so complex that determining the physiological effects of mixtures

47 would be difficult without running specific physiological experiments. At best we can hypothesize that the toxic effects of one chemical make D. magna more susceptible to the effects of another chemical, resulting in synergism, or that the chemicals do not make D. magna more susceptible to a second chemical, resulting in additive toxicity. This discussion becomes even more difficult because the specific mechanisms of action for all of the involved chemicals, i.e. chlorophenyl isocyanates, are not known. Without fully understanding the physiological responses of D. magna to each individual compound, parent or photoproduct, it is difficult to hypothesize as to why these results were seen.

The mechanisms of action are not known for all of the chemicals the D. magna were exposed to. This is a common issue in mixture toxicity testing. While there have been many studies investigating the toxicities of chemical mixtures, the discussions of these studies have focused on trends of mixture composition and environmental implications, as opposed to the specific mechanism of toxicity.

The general conclusion reached by this study is that since TCC and TCS can both be photolyzed, and both have multiple degradation products, determining the mechanisms for mixture toxicities can be difficult. It is therefore more important to be aware of potential synergisms of these mixtures that could lead to higher toxic effects in aquatic organisms as a way to help when performing risk assessments. This is especially relevant to TCS-TCC mixtures because of their frequent co-occurrences in the surfaces of freshwaters. There is a high likelihood that aquatic organisms are being exposed to both parent compounds and mixtures of their photolysis products, implying that studies of

48 ternary or higher mixtures would also be relevant and important in understanding the implications of complex TCS, TCC, and photolysis product mixtures.

3.5 Tables

Table 3.1 Potential photolysis product exposure of aquatic organisms. Potential involved compounds in each mixture based on the known photolysis products of TCS and TCC. The two compounds of the binary mixtures are listed under the Compound 1 and Compound 2 categories with “ph” indicating that the compound was photolyzed. The likely compounds in each solution due to parent compounds and/or known photolysis products are highlighted in each row.

Compound 1 Compound 2 Potential Exposure Compounds ph TCC in DOM ph TCS in DOM TCS TCC 2,8-DCDD 2,7-DCDD 2,4-DCP 4-CP 3,4-DCP 4-CPI 3,4-CPI ph TCC in DOM ph TCS in EPA TCS TCC 2,8-DCDD 2,7-DCDD 2,4-DCP 4-CP 3,4-DCP 4-CPI 3,4-CPI ph TCC in DOM TCC TCS TCC 2,8-DCDD 2,7-DCDD 2,4-DCP 4-CP 3,4-DCP 4-CPI 3,4-CPI ph TCC in DOM TCS TCS TCC 2,8-DCDD 2,7-DCDD 2,4-DCP 4-CP 3,4-DCP 4-CPI 3,4-CPI TCC TCS TCS TCC 2,8-DCDD 2,7-DCDD 2,4-DCP 4-CP 3,4-DCP 4-CPI 3,4-CPI

3.6 Figures

Figure 3.1 Toxicity of parent and photolyzed TCS with and without DOM. Acute toxicity of triclosan (TCS) and photolyzed TCS in EPA moderately hard water with and without Suwanee River Natural Organic Matter (SRNOM). LC50s (shown with 95% CI error bars) for parent TCS without and with DOM were 1.624 µM and 1.765 µM, respectively, while LC50s for photolyzed TCS without and with DOM were 8.482 µM and 8.513 µM, respectively. Each LC50 was assigned a letter (a-d), and statistical similarities were therefore indicated by placing the corresponding letter above LC50 bars that were not statistically different (i.e., parent TCS without DOM is “a” and is statistically not different from itself, “b,” and “d”) (p ≤ 0.05).

Figure 3.2 Mixture toxicities. Acute toxicities, represented as TU50 values with 95% CI error bars, of binary mixtures of triclosan (TCS), triclocarban (TCC), and the products of photolysis with and without SRNOM (DOM). “ph” indicates that the compound was photolyzed. Mixtures with 95% CI levels that overlapped with a TU50 of 1 were considered additive, while mixture with CI levels that did not overlap with a TU50 of 1 were considered either antagonistic (above 1) or synergistic (below 1). Asterisk (*) indicates significant difference from TU50=1 (p ≤ 0.05).

3.7 References

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Chapter 4: Enhancement of triclosan photolysis by bicarbonate ion-pair formation

4.1 Introduction Triclosan (TCS; 5-chloro-2-(2,4-dichlorophenoxy)phenol) is a commonly used antimicrobial compound found in many household items, including soaps and toothpastes. Even though TCS is efficiently removed in wastewater treatment plants

(WWTPs) (Thomas and Foster, 2005), it is still ubiquitously present in natural aquatic systems. Triclosan has been detected in WWTP effluents, rivers, lake waters, sediment cores, and suspended sediments (Singer et al., 2002; Thomas and Foster, 2005; Buth et al., 2011; Venkatesan et al., 2012; Anger et al., 2013; Gautam et al., 2014; plus many others). In addition TCS has the potential to become further chlorinated in WWTPs, and these chlorinated derivatives have also been found in surface waters and sediments near

WWTP output sites (Buth et al., 2011).

Numerous studies have investigated the photolysis of TCS in surface waters, and many phototransformation products have been identified, including numerous chlorinated dioxins from both TCS and its chlorinated derivatives (Latch et al., 2003; Mezcua et al.,

2004; Latch et al., 2005; Buth et al., 2009; Buth et al., 2011; Buth et al., 2010) and 2,4- dichlorophenol (Latch et al., 2005). The toxicity of TCS has also been widely studied,

and results show toxicity to aquatic organisms, including Daphnia spp. and various fish species (Orvos et al., 2002).

Numerous studies have investigated the effects of carbonate and bicarbonate on the photolysis of different compounds (Canonica et al., 2003), with results suggesting that carbonate/bicarbonate both enhances (Huang et al., 2000a; Huang et al., 2000b;

Tercero et al., 2007; Canonica et al., 2008; Chiron et al., 2009; Vione and Khanra, 2009;

Wallace et al., 2010; Mao et al., 2011; Ahmad 2014) and inhibits phototransformations

•- (Lam et al., 2003; Wallace et al., 2010). In all cases the carbonate radical anion (CO3 ) formed from the interaction of hydroxyl radicals and carbonate appears to be the responsible reactive intermediate. Since the carbonate radical anion is strongly electrophilic, it can selectively react via oxidation with compounds that are electron rich

(Chen et al., 1975; Huang et al., 2000b; Wallace 2010; Mao et al., 2011). Conversely, at relevant pH values (> 8) carbonate can act as a photolysis inhibitor by scavenging hydroxyl radicals (through the formation of carbonate radical anions). Since hydroxyl radicals are much more reactive to a wider class of organic contaminants, they are crucial to the phototransformation of these substances in sunlit natural waters (Zepp et al., 1987;

Larson et al., 1988; Brezonik et al., 1998; Miller and Chin., 2002; Lam et al., 2003; Halle et al., 2006; Vione and Khanra, 2009).

During our own initial photolysis experiments for the toxicity-related aspect of this study we noted a difference in degradation rates of TCS in waters differing in ion composition. The initial experiments conducted in Milli-Q and bicarbonate buffered water (based upon the composition of the EPA water) revealed significant differences in

degradation kinetics beyond those expected between the acidic and deprotonated forms of

TCS (pKa 7.9). For this paper we wanted to investigate the mechanism for this enhanced photolysis.

Degradation rates of many compounds are pH dependent (Canonica et al., 2008;

Ahmad et al., 2014), and the phototransformation of TCS is believed to be dependent on the degree of TCS deprotonation (Latch et al., 2005). We investigated the effects of buffers on TCS direct photolysis at pHs above and below the pKa to assess whether buffers can enhance the photolysis of unprotonated TCS through the creation of an ion- pair complex. The formation of this ion-pair below the pKa of TCS would greatly enhance its phototransformation at low pH values.

4.2 Materials and Methods

4.2.1 Chemicals

Triclosan (≥97%, as Irgasan) and DABCO (1,4-diazabicyclo[2.2.2]octane)

(>99%) were purchased from Sigma-Aldrich (St. Louis, MO). NaHCO3 (99-100%), KCl

(99.4%), MgSO4 (99%), and CaSO4 (98%) were purchased from Fisher Scientific

(Waltham, MA). HPLC grade acetonitrile and isopropanol were also purchased from

Fisher Scientific (Waltham, MA). Methyl triclosan (MeTCS) was synthesized in the lab.

4.2.2 Photolysis solution preparation

All solutions were prepared using 18.2 MΩ Milli-Q water (EMD-Milli-Pore,

Billerica, MA). Water constituents included NaHCO3 (96 mg/L=1.14 mM, or 9.6 mg/L,

62 or 960 mg/L), CaSO4 (60 mg/L), MgSO4 (60 mg/L), KCl (4 mg/L), DABCO (1 mM), isopropanol (25 mM), and KH2PO4/K2HPO4 (1.14 mM). Buffered solutions were prepared based upon concentrations used in EPA water, which is commonly used in toxicity testing (US EPA, 2002). Only the bicarbonate concentrations were ever modified from the suggested EPA recipe. In a dark room, TCS or Me-TCS was added by pipetting a 4,000 µM stock solution in MeOH into a beaker, evaporating the MeOH, and diluting to working solution concentrations using Milli-Q or buffered water and the pH adjusted using HCl and NaOH.

4.2.3 Photolysis Reactions

Solutions were placed into quartz phototubes with a 1-cm pathlength, capped, and placed in an Atlas SunTest CPS+ (Atlas Material Testing Technology, Illinois, USA), which uses a Xenon light source with the appropriate sunlight filter (Filter H). The temperature inside the solar simulator was held constant at room temperature (25° C) by a cooling system, and irradiation was monitored with a VWR UV Light Meter/Ultra-violet

Radiometer to confirm the consistency of irradiation levels. Phototubes were removed from the solar simulator at predetermined time points, placed in an amber vial, and stored in a dark room until all samples were removed from the solar simulator. A time zero sample was placed into an amber vial and stored in the dark while the other samples were irradiated, and a dark control was placed in an amber vial, covered in foil, and left in the solar simulator until the last phototube was removed. To compare irradiated versus non- irradiated samples, we also ran dark experiments in which solutions were covered and

63 placed in a cupboard in a dark room to prevent light exposure. Samples were then treated the same as with samples in the solar simulator.

4.2.4 Methyl-triclosan synthesis

Methyl-triclosan was synthesized in the Hadad Lab at The Ohio State University from TCS powder. A 1 mM TCS solution was prepared in methanol and 350 µL of 40%

NaOH solution and 110 µL (1.14 mmol) of dimethyl sulfate were added and allowed to reflux overnight. The solution was then cooled to room temperature and extracted three times with 5 mL of diethyl ether. The ether layer was then dried over Na2SO4 followed by rotary evaporation to yield a colorless oil. Column chromatography coupled with proton and carbon 13 nuclear magnetic resonance (1H-NMR, 13C-NMR) spectroscopy, and heteronuclear multiple quantum coherence (HMQC)-NMR, a form of 2-dimensional

NMR, were used to confirm the existence of pure methylated TCS.

4.2.5 HPLC Analysis

Samples were analyzed via high-pressure liquid chromatography (HPLC) using a

Waters 1515 Isocratic Pump, with Waters 717plus autosampler and Waters 2487 Dual

Wavelength Absorbance UV-detector. 50 µL of solutions were injected onto a Restek

Pinnacle DB C18 column, (5 µM particle size, 150x4.6 mm), at a flow rate of

0.8mL/minute and with an isocratic mobile phase of 70:30 (v:v) of acetonitrile:water.

Triclosan was detected at 280 nm at a retention time of 10.7 minutes, with a detection limit of 0.01 µM.

4.2.6 Data Analysis

Linear regression, i.e. plotting ln C/CO, was used to determine degradation rate constants, or kobs, which were equal to the slopes of the lines. Degradation rate constants were compared by calculating 95% confidence limits of the slope values (Table F.3) If

95% CI values did not overlap, slopes were considered different from each other.

4.3 Results

4.3.1 The effects of solution pH on direct photolysis

We determined the photolysis rate constants of TCS in the absence of any buffers at varying pH values to determine the effects of deprotonation on photolysis rates. With a pKa of 7.9, TCS can exist as both ionized and neutral forms under the pH regimes present in a variety of natural waters. Photodegradation of the conjugate base of TCS is significantly faster than its acidic form, and the irradiation of the ionized form of TCS that undergo rearrangement and cleavage to form photolysis products such as the chlorinated dioxins (Latch et al., 2005; Wong-Wah-Chung et al., 2007; Kliegman et al.,

2013). At a pH well above the pKa of TCS (pH 9.12), the observed rate constants were orders of magnitude faster (1.9285 min-1) than observed at lower pH values of 6.55 and

6.84 (0.0382 min-1 and 0.0039 min-1, respectively), which are well below the pKa of

TCS. These values are similar to values reported by others (Tixier et al., 2002) (Figures

4.1a-c). While some degradation rate constants were more similar in value than others, all of the kobs calculated were significantly different from each other (Table F.3)

4.3.2 Effects of MgSO4/CaSO4

In an effort to elucidate the reaction mechanism we conducted experiments in

EPA water, which contains separately and in various combinations the following salts;

NaHCO3, MgSO4, CaSO4, and KCl, i.e., we wanted to assess the role of each salt individually in enhancing TCS photolysis. When comparing the degradation kinetics of a

60 mg/L MgSO4 + 60 mg/L CaSO4 solution to a 96 mg/L NaHCO3 solution (all concentrations based on the EPA water) at pH 6.00, we saw that the kobs of TCS in the sulfate water was an order of magnitude slower, and statistically different, than that of the

-1 -1 NaHCO3 water (0.0394 min and 0.3288 min , respectively), indicating that the sulfate salts in EPA water are not participating in any TCS photolysis enhancement (Figure 4.2).

2+ 2+ As salts, MgSO4 and CaSO4 will completely dissociate in water, leaving Mg , Ca , and

2- SO4 . Since sulfate is the conjugate base of the mineral acid H2SO4, it has no propensity to take on any protons. This is similar for KCl, which is the salt form of the mineral acid

- HCl. In contrast, when NaHCO3 dissociates, it will form the conjugate bases (HCO3 , and

-2 - 2- CO3 ) of carbonic acid, H2CO3 with pKas of 6.3, and 10.3. Since HCO3 and CO3 bases are conjugates of a weak acid there is potential that they can interact with other proton donors such as TCS to form ion-pair complexes.

4.3.3 Buffer effects

To compare the effects of the carbonate/bicarbonate to commonly used phosphate buffers, TCS was photolyzed in 1.14 mM potassium phosphate dihydrogen and NaHCO3

66 solutions (96 mg/L NaHCO3 = 1.14 mM). We found nearly a three-fold significant

- difference between photolysis rates in the phosphate buffered solution (kobs=0.3129 min

1 -1 ) and NaHCO3 buffered solution (kobs=1.0383 min ) at a pH of 6.73, where TCS is predominantly in the acidic form (Figure 4.3), but these differences were not of the same magnitude as the difference seen with and without buffers. One study reported a kobs of

0.0228 min-1 in a 5 mM phosphate buffer at pH 5.9 (Tixier et al., 2002), which is a

-1 similar value to the kobs of 0.0494 min in NaHCO3 at pH 5.74 (Figure 4.4a).

These values, however, are significantly faster than the results in unbuffered water at similar pHs. At pHs of 6.55 and 6.84 the kobs of TCS in Milli-Q water (0.0039-

0.0382 min-1) was one to two orders of magnitude slower than in the buffered solutions

-1 (kobs 0.4781-0.7109 min ) (Figure 4.1a-b). At a higher pH of 9.12, nearly 1.5 pH units

-1 above the pKa, the kobs values were more similar in value at 1.9285 min for direct

-1 photolysis in Milli-Q and 2.8605 min for photolysis in NaHCO3 buffered water, showing only a 1.5-fold difference (Figures 4.1c). While some degradation rate constants were more similar in value than others, all of the kobs calculated were significantly different from each other (Table F.3)

4.3.4 Effects of radical scavengers

Photochemically formed radicals have been found to influence degradation of organic compounds (Kochany and Bolton, 1992; Oturan, 2000; Zhao et al., 2004), and

1 based on the constituents of our solutions we suspected that singlet oxygen ( O2),

• •- hydroxyl radicals (OH ), and carbonate radicals (CO3 ) could be potentially formed from

67 irradiated TCS i.e., self-sensitized reaction. Singlet oxygen is formed via scavenging of triplet state sensitizers, which absorb UV irradiation to their singlet and then triplet excited states. The scavenging of these triplets by oxygen leads to the formation of singlet oxygen, a reactive oxygen species (ROS) (DeRosa and Crutchley, 2002). Singlet

•- oxygen can then be reduced to form superoxide (O2 ) ion, another ROS (Saito and

Matsuura, 1981). Hydroxyl radicals can be formed in water due to Fenton reactions

- - (interaction of Fe II and hydrogen peroxide to form OH• and OH ), nitrate (NO3 )/nitrite

- (NO2 ) irradiation, or irradiation of hydrogen peroxide (H2O2) (Legrini et al., 1993;

Richard and Canonica, 2005; Vione et al., 2006). These hydroxyl radicals can then form carbonate radicals via scavenging of an electron from a carbonate molecule (Richard and

Canonica, 2005). To determine if radicals were involved in mediating TCS photolysis, we ran experiments at pH 5.74 with 1,4-diazabicyclo[2.2.2]octane (DABCO), a singlet oxygen quencher (Ouannes and Wilson, 1968; Packer et al., 1981). In a 1 mM solution of

NaHCO3, DABCO exhibited a two-fold significant enhancement on observed photolysis

-1 -1 rates compared to a solution without DABCO (kobs of 0.0875 min , 0.0494 min , respectively) (Figure 4a). There was also a significant difference between 1 mM

-1 -1 phosphate solutions with and without DABCO (kobs of 0.0595 min , 0.0759 min , respectively) (Figure 4a), indicating that DABCO was possibly acting as a sensitizer in these reactions due to its nucleophilic nature (the reasoning for this is discussed in relation to carbonate in the explanation of our proposed mechanism), and not participating in the quenching of potential radicals in these specific solutions. These results do show that singlet oxygen is not a likely player in the TCS photolysis pathway.

4.3.5 Photolysis of methyl-triclosan

Methyl-triclosan was synthesized because it is structurally the same as TCS, with the exception of a methoxy group in place of the hydroxyl group of TCS. Due to the presence of this methoxy group, Me-TCS is unionizable, and cannot form ion-pair complexes with the conjugate bases of weak acids. Methyl-triclosan was photolyzed in

Milli-Q (pH 6.71) and EPA water at various pH values (pHs 3.74, 6.80, 10.24) and we observed no degradation regardless of pH (Figure 4.5), whereas fully deprotonated TCS

-1 at pH 10.22 degraded quite rapidly (kobs=2.6361 min ) (Figure 4.5). These results demonstrate that the observed photolysis enhancement is dependent on the ionization of

TCS and interaction of the hydroxyl group with carbonate and bicarbonate given that Me-

TCS did not degrade under any conditions.

4.3.6 Effects of bicarbonate concentration

Three pH 6.55 solutions in Milli-Q water containing TCS and either 9.6 mg/L, 96 mg/L, or 960 mg/L (nominal concentrations) sodium bicarbonate (~0.1 mM, 1 mM, or 10 mM), were photolyzed to determine the effects of varying the bicarbonate concentrations on the enhancement of TCS photolysis. This study resulted in faster degradation with increasing bicarbonate concentration, further supporting our hypothesis that bicarbonate enhances the degradation rate of TCS at pHs below its pKa. At a pH of 6.55 the kobs were

0.0789, 0.1435, and 0.1687 min-1, respectively (Figure 4.6), with a statistical difference between the 9.6 mg/L treatment and the other two treatments. We attribute the lack of

69 statistical difference between 96 mg/L and 960 mg/L to the saturation of TCS by bicarbonate, meaning that regardless of continued increase in bicarbonate concentrations, all of the TCS molecules are associating with bicarbonate so there will be no further enhancement.

4.4 Discussion

4.4.1 Effects of ionization

The direct photolysis experiments indicated that TCS photodegrades faster with increasing pH, corroborating studies that suggest that the mechanism for TCS photolysis begins with the deprotonation of TCS, which in the presence of UV irradiation undergoes rearrangement and degradation (Latch et al., 2005; Wong-Wah-Chung et al., 2007;

Kliegman et al., 2013). We observed degradation rate constants with a general trend of increasing with increasing pH, which corroborates observations reported by others

(Tixier et al, 2002). Photolysis experiments with Me-TCS, an unionizable form of TCS, further supported the premise that the conjugate base of TCS is the more reactive form.

Unlike Me-TCS, however, TCS photodegradation was observed at all pHs. We believe that the slower reaction rates at pH values well below the pKa could be attributable to the relative abundance of protonated versus unprotonated TCS and that at very low pH

(orders of magnitude below its pKa) TCS would be relatively unreactive, similar to observations noted for the photolysis of Me-TCS.

Carbonate and bicarbonate, formed from the deprotonation of carbonic acid, appear to have the potential to interact with the hydroxyl group of TCS, as they are both

70 capable of binding an H+. In contrast, sulfate and chloride anions are the conjugate bases of very strong acids and will not interact with H+ ions in solution or with the H+ of the hydroxyl group of TCS. The potential effects of carbonate/bicarbonate on the kobs of TCS and its role in enhancing the photolysis of TCS at pH values well below its pKa are discussed below.

4.4.2 Mechanism of bicarbonate/carbonate enhancement

Experimental results indicated that carbonate/bicarbonate and phosphate are the constituents enhancing the photolysis of TCS, but does not involve photosensitization. As this enhancement is not due to the formation of carbonate radicals, we suggest that it is simply the carbonate/bicarbonate ions that are influencing TCS photolysis. Based on our results a mechanism was proposed in which negatively charged carbonate or bicarbonate ions form an ion complex with TCS via the interaction with the TCS hydroxyl group, leading to the deprotonation of TCS below its pKa followed by subsequent photodegradation (Figure 4.7). The bicarbonate ion has a partial negative charge which would be attracted to the positively charged hydrogen on the hydroxyl group of TCS, and this interaction would make it more likely that TCS would become deprotonated, even at this lower pH, explaining the bicarbonate-mediated photolysis enhancement seen in our experiments.

4.4.3 Environmental Implications

Most natural waters have a wide range of pH values ranging from slightly acidic

(~ 5) to more alkaline (9 or even higher), and the photolysis of the conjugate base of TCS is the dominant specie involved in its photofate. Based on this, photolysis would be significantly slower in waters with pHs well below the pKa of TCS, which is contrary to our experimental results. Under acidic conditions where bicarbonate can still exist at significant levels, e.g. near its pKa, and can form an ion-pair complex with TCS whereby the proton on the TCS hydroxyl functional group interacts with both the buffer and the parent compound. While less reactive than the conjugate base form of TCS, this protonated form of TCS does have an increased likelihood of undergoing photolytic degradation when exposed to buffers such as bicarbonate and phosphate compared to

TCS not exposed to buffers.

4.5 Figures

Figure 4.1 Photolysis of buffered and unbuffered waters. Comparisons of the kobs, represented by the slopes of the regression lines, of TCS in Milli-Q water, and US EPA moderately hard water and NaHCO3 water (both containing 96 mg/L NaHCO3) at different pHs (a) 6.55 (b) 6.84 (c) 9.12. All of the kobs values were significantly different from one another. 73

Figure 4.2 Effects of bicarbonate and sulfates. Comparison of the effects of the different constituents of EPA water on TCS photolysis at pH ~6. Degradation rate constants, or kobs, represented by the slopes of the regression lines, were statistically -1 different from one another, with the sulfate solution having a kobs of 0.0394 min and the -1 bicarbonate solution a kobs of 0.3288 min .

Figure 4.3 Effects of different buffers. Effects of two different buffers on the photolysis of TCS at pH 6.73. Degradation rate constants, or kobs, are represented by the slopes of - the regression lines. Triclosan degraded about three times faster in NaHCO3 (1.0383 min 1 -1 ) vs phosphate (0.3129 min ); these kobs values were statistically different from one another.

Figure 4.4 Photolysis with DABCO. The effects of 1 mM DABCO on the photolysis of TCS in NaHCO3 and phosphate buffers at pH 5.74. The kobs (represented by the slope of the regression line) of TCS was faster with DABCO in relation to the NaHCO3 solutions (0.0494 min-1 without, 0.0875 min-1 with), but slower with DABCO in relation to the -1 -1 phosphate solutions (0.0759 min without, 0.0599 min with). All of the kobs values were significantly different from one another.

Figure 4.5 Methyl triclosan photolysis. The effects of pH and NaHCO3 on methyl- triclosan (Me-TCS) photolysis compared to dark reactions and photolysis at pH 10. Methyl-triclosan did not degrade in any solution, nor did TCS left in the dark. Triclosan in Milli-Q water at pH 10.22 rapidly degraded, with a kobs, represented by the slope of the regression line, of 2.6361 min-1.

Figure 4.6 Effects of bicarbonate concentration on photolysis rates. The effects of NaHCO3 (9.6, 96, and 960 mg/L) concentration on TCS photolysis. At a pH of 6.55 the Kobs, represented by the slopes of the regression lines, increased with increasing NaHCO3 -1 -1 -1 concentration (0.0789 min , 0.1435 min , and 0.1687 min , respectively). The kobs of the 9.6 mg/L solution was significantly different from the 96 and 960 mg/L solutions, however the 96 and 960 mg/L solutions were not statistically different from one another.

Figure 4.7 Proposed TCS mechanism. Proposed mechanism for the ion-pair formation and subsequent enhancement of TCS photolysis at pHs below the pKa of TCS. “B” represents a bicarbonate ion.

4.6 References

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Wong-Wah-Chung, P., S. Rafqah, G. Voyard, & M. Sarakha. 2007. Photochemical behavior of triclosan in aqueous solutions: kinetic and analytical studies. Journal of Photochemistry and Photobiology A: Chemistry 191: 201-208. Zepp, R. J., J. Hoigne, & H. Bader. 2007. Nitrate-induced photooxidation of trace organic chemicals in water. Environmental Science and Technology 21:443-450. Zhao, W., W. Ma, C.Chen, J. Zhao, & Z. Shuai. 2004. Efficient degradation of toxic organic pollutants with Ni2O3/TiOs-x/Bx under visible irradiation. Journal of the American Chemical Society 126: 4782-4783.

Chapter 5: Conclusion

This study investigated the aquatic toxicity and photofate (in the presence of artificial sunlight) of two widely used antimicrobial compounds, triclosan (TCS) and triclocarban (TCC) in the environment. This study focused on understanding how photolytic processes are influenced by complex solution matrices such as dissolved organic matter (DOM) presence, type, pH, and the presence of buffers. A summary of findings for each of the parts of this study is provided below:

1. Investigation into the influence of dissolved organic matter on the photolysis of

TCC revealed that the products formed during TCC photolysis in the presence of

DOM were inherently more toxic to Daphnia magna than the parent compound

and those formed by direct photolysis of the analyte. The toxicity of TCC

photoproducts formed by direct photolysis significantly decreased relative to the

toxicity of the parent compound, while its toxicity of TCC in US EPA

moderately hard water after photolysis (containing DOM) remained similar

(slightly lower LC50) to the parent compound. This was attributed to the

formation of toxic chloroaniline and chlorophenyl isocyanate products resulting

from the interaction of TCC with reactive intermediates e.g., triplet and reactive

oxygen species formed from irradiated DOM. My toxicity tests with these

individual compounds revealed high acute toxicity to D. magna, and is similar to

that observed for the indirect photolysis of TCC.

2. My assessment of mixture interactions between TCS, TCC, and their photolysis

products formed with and without DOM further supported the premise that

mixtures are extremely complex and difficult to predict or understand. My

findings demonstrated that mixtures of these compounds and their products could

exert synergistic, antagonistic, or additive toxicities, relative to the binary mixture

used (Chapter 3). DOM did not appear to influence the toxicity of TCS or its

photolysis products in this study, as opposed to the effects on TCC as described

above.

3. The presence of buffers such as carbonate in aquatic systems can significantly

enhance the rate of TCS direct photolysis at pH values well below its pKa. I

attributed this observation to the formation of an ion-pair between bicarbonate

(and to a lesser extent carbonate) and the hydroxyl group of TCS, which

ultimately causes partial deprotonation of the analyte adjacent to the buffer

molecule. As a result the macro-scale observation of the “unprotonated”

compound is significantly faster in the presence of these buffers relative to

distilled water at the identical pH.

Based on the results of this study, I conclude that safer alternatives to these two compounds need to be implemented in soaps and household cleaning products.

TCS has been found in concentrations as high as 2.3 µg/L (0.008 µM) in surface waters and 10 µg/L (0.035 µM) in wastewater treatment plant effluent (WWTPE)

(Lopez-Avila et al., 1980; Kolpin et al., 2002), and is known to form potentially hazardous chlorinated dioxins when photolyzed (Buth et al., 2009). TCC has been

found at even higher concentrations (as high as 6.75 µg/L or 0.021 µM) in surface waters (Halden and Paull, 2005), and my research demonstrates that it forms chloroanilines when photolyzed in the presence of DOM, which I demonstrated to have acute toxicity to Daphnia magna. While the toxic concentrations for TCS (1

µM) and its photolysis products (8 µM) are significantly higher than levels found in the environment, the potential for bioaccumulation in organisms should be considered when assessing potential (especially chronic) toxicity. Further, my research on TCC demonstrates that its photoproducts exhibit much higher and known acute toxicity than TCS. While many countries such as those in the European

Union (European Chemicals Agency) have banned the use of TCS, and many companies have voluntarily removed it from their products (personal correspondence with The Procter and Gamble Company, Cincinnati, OH), less initiative has been taken for TCC.

My research demonstrates the importance of taking into consideration ALL natural water constituents when assessing both the photodegradation of the analyte

(including the ubiquitous carbonate system present in all natural waters)and its potential risk (as well as the toxicity of its derivatives) after discharge to the environment. My research has shown for example that carbonate/bicarbonate can influence photolysis of TCS in a totally unexpected manner. Given that nearly all waters are buffered by the carbonate system it is conceivable that TCS would be more photo-recalcitrant in very poorly buffered low pH (e.g., streams affected by acid mine drainage) sunlit surface waters. Thus, I suggest that other chemical

87 constituents present in natural waters should also be considered when assessing the environmental fate of aquatic contaminants. Finally, my research demonstrated that

DOM changes the photoproduct composition of TCC photolysis resulting in the formation of highly toxic derivatives. While my research demonstrate the acute toxicity of these compounds at higher than concentrations present in natural waters the chronic toxicity of these substances remain unknown.

To date there have been few studies that have integrated basic reaction kinetic studies to toxicity. Future studies should include assessment of the interactions of more complex mixtures, including tertiary mixtures to mixtures containing dozens of chemicals commonly discharged to aquatic systems as mixtures to more accurately and precisely describe the potential exposure and effects on aquatic life. The photolysis and toxic effects of other chemicals that are prominent in aquatic systems besides TCS and TCC should also be considered e.g., surfactants, chelating agents, etc. as my research only investigated a simplistic binary combination of TCS, TCC, and their photolysis products in the presence of DOM.

5.1 References

Beyond Pesticides. June 2015. EU to ban triclosan, while EPA and FDA reject calls for U.S. ban. Accessed 2 May 2016. http://beyondpesticides.org/dailynewsblog. European Chemicals Agency. 2015. The Biocidal Products Committee adopts 11 opinions. http://echa.europa.eu. Halden, R.U., & D.H. Paul. 2005. Co-occurrence of triclocarban and triclosan in U.S. water resources. Environmental Science and Technology 39: 1420-1426. Kolpin, D.W., E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber, & H.T. Buxton. 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams 1999-2000: a national reconnaissance. Environmental Science and Technology 36: 1202-1211. Lopez-Avila, V., & R.A. Hites. 1980. Organic compounds in industrial wastewater. Their transport into sediments. Environmental Science and Technology 14: 1382-1390.

Appendix A: Chemical Structures

Figure A.1 Structure of triclosan.

Figure A.2 Structure of triclocarban.

Figure A.3 Structure of 4-chloroaniline.

Figure A.4 Structure of 3,4-dichloroaniline.

Figure A.5 Structure of 4-chlorophenyl isocyante.

Figure A.6 Structure of 3,4-dichlorophenyl isocyante.

Appendix B: Mass Spectra of TCC and Photolysis Products

Figure B.1 Mass spectra of photolyzed TCC in EPA-DOM water. Parent TCC was not detected in this solution.

Figure B.2 Mass spectra of TCC photolyzed in EPA with DOM. Detection was set for CA and CPI products.

Figure B.3 Mass spectra of 4-chloroaniline standard.

100

Figure B.4 Mass spectra of 3,4-dichloroaniline standard.

101

Figure B.5 Mass spectra of 4-chlorophenyl isocyanate standard.

102

Figure B.6 Mass spectra of 3,4-dichlorophenyl isocyanate standard.

103

Figure B.7 Comparison of photolyzed TCC and product standards. Combination of above mass spectra showing the corresponding peaks between photolyzed TCC in EPA- DOM and the proposed degradation product standards.

104

Figure B.8 Mass spectra of TCC photolyzed in EPA water without DOM. None of the degradation product peaks are seen.

105

Appendix C: LC50 Curves

106

Figure C.1 LC50 curve of parent TCC in EPA water.

107

Figure C.2 LC50 curve of parent TCC in EPA-DOM water. Water contained 12 mg/L DOM

108

Figure C.3 LC50 curve of photolyzed TCC in EPA water

109

Figure C.4 LC50 curve of photolyzed TCC in EPA water

110

Figure C.5 LC50 curve of photolyzed TCC in EPA water.

111

Figure C.6 LC50 curve of photolyzed TCC in EPA-DOM water. Water contained 12 mg/L DOM

112

Figure C.7 LC50 curve of photolyzed TCC in EPA-DOM water. Water contained 12 mg/L DOM

113

Figure C.8 LC50 curve of photolyzed TCC in EPA-DOM water. Water contained 12 mg/L DOM

114

Figure C.9 LC50 curve of 4-CA in EPA water.

115

Figure C.10 LC50 curve of 4-CA in EPA water.

116

Figure C.11 LC50 curve of 3,4-DCA in EPA water.

117

Figure C.12 LC50 curve of 3,4-DCA in EPA water.

118

Figure C.13 LC50 curve of 4-CPI in EPA water.

119

Figure C.14 LC50 curve of 4-CPI in EPA water.

120

Figure C.15 LC50 curve of 3,4-DCPI in EPA water.

121

Figure C.16 LC50 curve of 3,4-DCPI in EPA water.

122

Figure C.17 LC50 curve of parent TCS in EPA water.

123

Figure C.18 LC50 curve of parent TCS in EPA-DOM water. Water contained 12 mg/L DOM in EPA water.

124

Figure C.19 LC50 curve of photolyzed TCS in EPA water.

125

Figure C.20 LC50 curve of photolyzed TCS in EPA-DOM water. Water contained 12 mg/L DOM in EPA water

126

Figure C.21 TU50 curve of parent TCC and parent TCS mixture in EPA.

127

Figure C.22 TU50 curve of parent TCC and parent TCS mixture in EPA.

128

Figure C.23 TU50 curve of parent TCC and parent TCS mixture in EPA.

129

Figure C.24 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

130

Figure C.25 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM in EPA.

131

Figure C.26 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

132

Figure C.27 TU50 curve of mixture of parent TCC in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

133

Figure C.28 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

134

Figure C.29 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

135

Figure C.30 TU50 curve of mixture of parent TCS in EPA water and TCC photolyzed in EPA-DOM water. Water contained 12 mg/L DOM.

136

Figure C.31 TU50 curve of mixture of TCC photolyzed in EPA-DOM water and TCS photolyzed in EPA water. EPA-DOM water contained 12 mg/L DOM.

137

Figure C.32 TU50 curve of mixture of TCC photolyzed in EPA-DOM water and TCS photolyzed in EPA-DOM water. EPA-DOM water contained 12 mg/L DOM.

138

Figure C.33 TU50 curve of mixture of TCC and TCS photolyzed in same solution of EPA-DOM water. Water contained 12 mg/L DOM.

139

Appendix D: NMR of Synthesized Methyl-Triclosan

140

Figure D.1 1H NMR of synthesized Me-TCS.

141

Figure D.2 13C NMR of synthesized Me-TCS.

142

Figure D.3 HMQC NMR of synthesized Me-TCS. Heteronuclear Multiple Quantum Coherence (HMQC) NMR, a form of 2-dimensional NMR, of synthesized Me-TCS.

143

Appendix E: Toxicity of Dissolved Organic Matter

144

Table E.1. Toxicity of DOM. Daphnia magna mortality in EPA water, EPA water with dissolved organic matter (DOM), and EPA water photolyzed with DOM. There was no mortality seen in any treatment after 96 hours, indicating that all toxic effects noted in other tests were solely from TCS, TCC, or their photoproducts

24 48 72 96 Replicate hours hours hours hours 1 0 0 0 0 EPA Water 2 0 0 0 0 3 0 0 0 0 4 0 0 0 0 1 0 0 0 0 EPA water with 2 0 0 0 0 12mg/L DOM 3 0 0 0 0 4 0 0 0 0 EPA water with 1 0 0 0 0 12mg/L photolyzed 2 0 0 0 0 DOM 3 0 0 0 0 4 0 0 0 0

145

Appendix F: LC50s and Confidence Intervals

146

Table F.1 LC50s and 95% confidence intervals of single compounds.

DOM Compound Form LC50 (µM) 95%LCL 95%UCL (mg/L) Triclocarban Parent 0 0.086968983 0.039504716 0.191460787 Triclocarban Parent 12 0.146862468 0.049715568 0.433839648 Triclocarban Photolyzed 0 2.668077648 2.067427806 3.443234301 Triclocarban Photolyzed 12 0.032146466 0.017204448 0.06006559 Triclosan Parent 0 1.6235 0.920213 2.864214 Triclosan Parent 12 1.7647 1.174115 2.652282 Triclosan Photolyzed 0 8.4819 6.680821 10.76847 Triclosan Photolyzed 12 8.5131 2.700224 26.83979 4- Parent 0 0.082465302 0.034444324 0.235704614 chloroaniline 3,4- Parent 0 1.028895429 0.412608374 3.504869898 dichloroaniline 4- chlorophenyl Parent 0 >10* NA NA isocyanate 3,4- dichlorophenyl Parent 0 2.029909019 1.336539854 2.978640476 isocyanate * calculated LC50 was higher than the highest concentration tested

147

Table F.2 TU50s and 95% confidence intervals of mixtures.

Compound 1 Compound 2 TU50 95% LCL 95% UCL Parent Triclosan in Parent Triclocarban EPA in EPA 0.0246 0.006428 0.094317 Parent Triclosan in ph Triclocarban in EPA EPA-DOM 0.5298 0.40626 0.690889 Parent ph Triclocarban in Triclocarban in EPA EPA-DOM 0.475 0.340144 0.66408 ph Triclosan in ph Triclocarban in EPA-DOM EPA-DOM 1.45 0.534476 3.92427 Triclosan and Triclocarban ph 0.987 0.383241 2.54056 simultaneously in same soln ph Triclocarban in ph Triclosan in EPA EPA-DOM 0.679 0.233003 1.98119

148

Table F.3 Slopes and 95% confidence intervals of photolysis experiments.

Figure Solution pH Slope 95% LCL 95% UCL 1a Milli-Q 6.55 -0.038 -0.047 -0.029 Bicarbonate 6.55 -0.711 -0.807 -0.615 1b Milli-Q 6.84 -0.004 -0.164 0.157 Bicarbonate 6.84 -0.478 -0.526 -0.431 EPA 6.84 -0.292 -0.338 -0.246 1c MQ 9.12 -1.929 -2.307 -1.55 EPA 9.12 -2.86 -3.091 -2.63 2 Sulfate 6.00 -0.039 -0.049 -0.03 Bicarbonate 6.00 -0.329 -0.357 -0.3 3 Phosphate 6.73 -0.313 -0.34 -0.286 Bicarbonate 6.73 -1.038 -1.249 -0.827 4a Phosphate 4.74 -0.076 -0.08 -0.071 Phosphate+DABCO 4.74 -0.06 -0.062 -0.058 Bicarbonate 4.74 -0.049 -0.05 -0.048 Bicarbonate+DABCO 4.74 -0.088 -0.092 -0.083 4b Bicarbonate 8.09 -2.225 -2.454 -1.995 Bicarbonate 8.09 -2.05 -2.31 -1.79 Bicarbonate+IPP 8.09 -2.065 -2.308 -1.823 5 MQ 10.22 -2.636 -3.111 -2.161 6a 9.6mg/L Bicarbonate 6.75 -0.059 -0.127 0.01 96mg/L Bicarbonate 6.75 -1.038 -1.249 -0.827 960mg/L Bicarbonate 6.75 -2.158 -2.475 -1.841 6b 9.6mg/L Bicarbonate 6.55 -0.079 -0.104 -0.054 96mg/L Bicarbonate 6.55 -0.144 -0.167 -0.12 960mg/L Bicarbonate 6.55 -0.169 -0.223 -0.114

149

Appendix G: Compiled Toxicity Data

150

Table G.1 Compiled Toxicity Data. Data from other studies.

Reported Time Molarity Compound Endpoint Conc. Organism Reference Period (µM) (µg/L ) TCS Lethality 48 hr 19.9 0.069 C. dubia Orvos 2002 TCS Lethality 48 hr 390 1.347 D. magna Orvos 2002 fathead TCS Lethality 96 hr 260 0.898 Orvos 2002 minnow Cell Density DeLorenzo TCS 30 min 3.55 0.012 algae Decrease 2008 DeLorenzo TCS Luminescence 15 min 53 0.183 V. fischeri 2008 Atrazine Luminescence 30 min 69400 321.758 V. fischeri Palma 2008 Chlorpyrifos Luminescence 30 min 2890 8.243 V. fischeri Palma 2008 Atrazine Lethality 48 hr 35500 164.588 D. magna Palma 2008 Chlorpyrifos Lethality 48 hr 740 2.111 D. magna Palma 2008 Increased Jap. medaka Ishibashi TCS 21 day 602 2.079 Vitellogenin larvae 2004 Increased Jap. medaka Raut and TCS 35 day 101.3 0.350 Vitellogenin larvae Angus 2010 TCC Growth Inhib. 24 hr 295 0.935 algae spp Gao 2015 TCS Growth Inhib. 24 hr 1063 3.671 algae spp Gao 2015 P. TCC Growth Inhib. 72 hr 29 0.092 Tamura 2013 subcapitata P. TCS Growth Inhib. 72 hr 5.1 0.018 Tamura 2013 subcapitata TCC Immobilization 48 hr 10 0.032 D. magna Tamura 2013 TCS Immobilization 48 hr 180 0.622 D. magna Tamura 2013 Juvenile TCC 96 hr 85 0.269 O. latipes Tamura 2013 Mortality Juvenile TCS 96 hr 210 0.725 O. latipes Tamura 2013 Mortality P. TCC Growth NOEC 72 hr 5.7 0.018 Tamura 2013 subcapitata P. TCS Growth NOEC 72 hr 0.53 0.002 Tamura 2013 subcapitata TCC Repro. NOEC 8 day 1.7 0.005 C. dubia Tamura 2013 TCS Repro.NOEC 8 day 30 0.104 C. dubia Tamura 2013 Hatching/Hatc TCC 9 day 24 0.076 D. rerio Tamura 2013 hling Survival Hatching/Hatc TCS 9 day 26 0.090 D. rerio Tamura 2013 hling Survival

151

Appendix H: Toxicity Data in SI Units

152

Table H.1 Toxicity data in SI units

DOM 96-h LC50 96-h LC50 Compound State (mg/L) (µM) (µg/L) TCS Parent 0 1.6235 470.06819 TCS Photolyzed 0 8.4819 2455.849326 TCS Parent 12 1.7647 510.951238 TCS Photolyzed 12 8.5131 2464.882974 TCC Parent 0 0.087 27.45546 TCC Photolyzed 0 2.6681 841.998998 TCC Parent 12 0.1469 46.358702 TCC Photolyzed 12 0.0321 10.130118 4-CA Parent 0 0.0824 10.511768 3,4-DCA Parent 0 1.0289 166.6992913 4-CPI Parent 0 >10* >1536.6* 3,4-DCPI Parent 0 2.0299 381.6435289

* calculated LC50 was higher than the highest concentration tested

153

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