The Sorption and Transformation of Tylosin and Progesterone by Soils

THESIS

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

Allison Jean Kreinberg

Graduate Program in Environmental Science

The Ohio State University

2012

Master's Examination Committee:

Professor Yu-Ping Chin, Advisor

Professor Warren Dick

Professor Roman Lanno

Copyrighted by

Allison Jean Kreinberg

2012

Abstract

Growth promoters (GPs) are antibiotic or hormone compounds given to livestock at sub-therapeutic levels in order to promote growth rates and feed efficiency. Many GPs have limited bioavailability, meaning that a fraction of the compound will pass through the target animal without being metabolized. The application of manure containing the unmetabolized fraction as fertilizer can result in the introduction of GPs to the environment. The subsequent transport and environmental fate of GPs is largely controlled by the underlying soil matrix. Tylosin (a macrolide antibiotic) and progesterone (a naturally produced hormone) are both GPs with limited bioavailability that have been detected in surface waters in the United States. Despite their widespread use, the interactions of both of these compounds with soils are poorly understood. The purpose of this work was to help elucidate the fate of these two compounds in soil systems.

Batch sorption experiments were conducted with five different sterilized soils and both tylosin and progesterone in order to determine their interactions with soils. Kinetics experiments were completed in order to determine the amount of time required to reach quasi-equilibrium between the sorbed and aqueous phase. Sorption isotherms were also completed, with initial aqueous concentrations ranging from 0.05-5.0 µM and 0.5-5.0 µM for progesterone and tylosin, respectively. Analysis of the aqueous concentration after

ii quasi-equilibrium was reached was conducted via reverse phase high performance liquid chromatography (RP-HPLC). Both tylosin and progesterone were found to undergo strong non-linear sorption based upon Freundlich transformations of the isotherm data.

Concentration dependent partition coefficients were calculated for both compounds using the Freundich isotherm fits and an initial aqueous concentration of Caq=0.5 µM. This resulted in average log Koc=2.94 ± 0.10 for progesterone and log Koc=2.95 ±0.18 for tylosin. While no work has been published regarding the sorption of progesterone to soils, these results are similar to values reported for tylosin.

Mass balances yielded acceptable recoveries for progesterone; however, almost

40% of the added tylosin was lost by the time quasi-equilibrium were reached (7 days).

Based upon the hypothesis that this loss could be due to abiotic metal-oxide mediated degradation, the fate of tylosin in the presence of goethite (ɑ-FeOOH) and birnessite (δ-

MnO2) (two minerals ubiquitous to soils) was investigated. Tylosin was added to a suspension of the metal oxide at pH=6.5 and its aqueous concentration measured over time by reverse phase high pressure liquid chromatography (RP-HPLC) in order to determine the rate of tylosin transformation. Tylosin appeared to undergo pseudo-first order transformation kinetics, with half lives of 27 h and 139 h with goethite and birnessite, respectively. Possible transformation pathways and products were hypothesized based upon UV-Vis absorbance and liquid chromatography-mass spectrometry results. The transformation of tylosin in the presence of metal oxides has not previously been reported, and suggests an additional degradation pathway for tylosin in soils. These results, in addition to a better understanding of the sorption of tylosin and

iii progesterone, help us to better comprehend the ultimate environmental fate of both compounds in soils.

Acknowledgments

The completion of this work would not have been possible without the guidance and knowledge of my advisor, Yu-Ping Chin. I would also like to thank Marcella Card for her excellent training and advice. I am grateful for the assistance of Katie Albanese, Brandon

McAdams, Tingting Liu, Victor Perez, and Julie Sheets in completing sample analysis.

The continual friendship and enthusiam of the Chin research group members (both past and present) has been invaluable. Countless thanks are owed to the many friends who have always helped me enjoy my work and my life. In addition, I am grateful for the constant encouragement and support of my family. Financial support was provided by

The Ohio State University and the National Science Foundation.

Vita

June 2006 ...... Orange High School, Pepper Pike, OH

2010...... B.S. Biochemistry, Miami University,

Oxford, OH

2010 - Present ...... Graduate Teaching and Research Associate,

School of Earth Sciences, The Ohio State

University, Columbus, OH

Fields of Study

Major Field: Environmental Science

Table of Contents

Abstract ...... ii

Acknowledgments...... v

Vita ...... vi

List of Tables ...... ix

List of Figures ...... x

Chapter 1: The environmental impact of growth promoters used in agriculture...... 1

1.1 Introduction ...... 1 1.2 Background ...... 3 1.3 Compounds of interest ...... 7 1.4 Statement of objectives ...... 14

Chapter 2: Tylosin and progesterone sorption to soils...... 15

2.1 Introduction ...... 15 2.2 Methods...... 25 2.3 Results and discussion ...... 30 Kinetic studies ...... 30 Sorption isotherms ...... 31 Progesterone ...... 31 Tylosin ...... 38 2.4 Conclusions ...... 44

Chapter 3: Tylosin transformation by soil mineral oxides...... 57

3.1 Introduction ...... 57 vii

3.2 Methods...... 64 Mineral synthesis ...... 64 Transformation kinetics ...... 66 3.3 Results and Discussion ...... 69 Transformation kinetics ...... 69 Transformation pathways...... 73 3.4 Conclusions ...... 77

Chapter 4: Conclusions and implications of this work...... 90

4.1 Summary of results ...... 90 4.2 Environmental significance ...... 94 4.3 Future work ...... 95

Complete list of references ...... 98

viii

List of Tables

Table 2.1. Physical and chemical soil properties...... 26

Table 2.2. Linear parameters and partition coefficients for progesterone with initial aqueous concentrations from 0.5-5 µM...... 31

Table 2.3 Freundlich parameters and partition coefficients for progesterone soil sorption isotherms...... 32

Table 2.4. Linear correlation coefficients for progesterone over varying initial aqueous concentration ranges...... 34

Table 2.5. Linear parameters and partition coefficients for tylosin with initial aqueous concentrations from 0.5-5.0 µM...... 38

Table 2.6. Freundlich parameters and calculated partition coefficients for tylosin...... 39

Table 2.7. Percent weight of iron and manganese oxides (measured as Fe2O3 and MnO2) in soils...... 43 Table 3.1. The ratio of abundance of fragment ions to the tylosin parent peak after reacting with goethite for 7 days...... 75

List of Figures

Figure 1.1. The chemical structure of the four different forms of tylosin (A-D)...... 15

Figure 1.2. The chemical structure of progesterone...... 16

Figure 2.1. Anticipated transport pathways of veterinary pharmaceuticals in the environment ...... 46

Figure 2.2. Progesterone soil sorption kinetics with Finley soil and an initial nominal aqueous concentration of 3 µM...... 47

Figure 2.3. Tylosin soil sorption kinetics with Drummer soil and an initial nominal aqueous concentration of 3 µM...... 48

Figure 2.4. Linear sorption isotherms for progesterone with initial nominal aqueous concentrations from 0.5-5 µM. Vertical error bars represent one standard deviation...... 49

Figure 2.5. Freundlich sorption isotherms for progesterone with initial nominal aqueous concentrations from 0.5 -5 µM. Vertical error bars represent one standard deviation. .... 50

Figure 2.6. Linear sorption isotherms for progesterone with initial nominal aqueous concentrations from 0.05-0.5 µM. No error bars are depicted because these are the results of a single experiment...... 51

Figure 2.7. Linear sorption isotherms for progesterone with initial nominal aqueous concentrations of 0.05-5.0 µM. No error bars are included because the lower concentrations were only determined once...... 52

Figure 2.8. Linear soil sorption isotherms for tylosin with initial nominal aqueous concentrations from 0.5-5.0 µM. Vertical error bars represent one standard deviation. .. 53

Figure 2.9. Freundlich sorption isotherms for tylosin and initial nominal aqueous concentrations from 0.5 – 5 µM. Vertical error bars represent one standard deviation. . 54

Figure 2.10. Representative mass balances for progesterone based upon triplicate extraction of the soil with methanol. No error bars are included because these results are from a single trial using Finley soil ...... 55

Figure 2.11. Representative mass balances for tylosin based upon triplicate extraction of the soil with methanol. No error bars are included because these results are from a single trial using Drummer soil...... 56

Figure 3.1. XRD powder scan of synthesized birnessite (green) compared to a birnessite reference pattern (orange vertical lines above background). The birnessite from which the reference pattern was calculated has chemical composition Mn2O4K0.46At1.4. (Post et al., 1990)...... 80

Figure 3.2. Kinetics of transformation of tylosin by goethite (ɑ-FeOOH) at pH=6.5 and an initial nominal tylosin concentration of ~1 µM. Error bars represent one standard deviation...... 81

Figure 3.3. Kinetics of transformation of tylosin by birnessite at pH=6.5 and an initial nominal tylosin concentration of ~1 µM. Error bars represent one standard deviation .. 82

Figure 3.4. Kinetics of transformation of tylosin by goethite (ɑ-FeOOH) at pH=6.5 and an initial nominal tylosin concentration of ~1 µM under sterile conditions. Error bars represent one standard deviation. Error bars for sterile conditions are not included because only one replicate was completed...... 83

Figure 3.5. Kinetics of transformation of tylosin by birnessite (δ-MnO2) at pH=6.5 and an initial nominal tylosin concentration of ~1 µM under sterile conditions. Error bars represent one standard deviation. Error bars for sterile conditions are not included because only one replicate was completed...... 84

Figure 3.6. Kinetics of transformation of tylosin by birnessite and goethite at pH=6.5 with an initial nominal tylosin concentration of ~ 1 µM and a DOM concentration of 20 mM. No error bars are present because these results represent a single replicate...... 85

Figure 3.7. RP-HPLC chromatogram for tylosin in contact with goethite for 1 h (blue) and 7 d (red), with absorbance at 287 nm...... 86

Figure 3.8. RP-HPLC chromatogram for tylosin in contact with birnessite for 1 h (blue) and 7d (red), with absorbance measured at 287 nm...... 87

Figure 3.9. The complexation of tylosin A with birnessite via the formation of a six- membered ring as a precursor to the formation of tylosin B...... 88

Figure 3.10. The complexation of tylosin B with birnessite via the formation of a six- membered ring as a precursor to the formation of product A...... 89

xii

Chapter 1: The environmental impact of growth promoters used in agriculture.

1.1 Introduction

The development of industrial agriculture has significantly increased the amount of livestock produced in the United States each year. Through the early 21st century, approximately 90 million cattle, 60 million swine, and 7 billion chickens were raised in the United States each year (MacDonald, 2008; USDA, 2009; USDA, 2012). Industrial agriculture has resulted in a higher percentage of these animals being raised at higher densities. While 690,000 producers raised 20 billion pounds of livestock in 1984, by

2000 only 95,000 producers raised 26 billion pounds of pork (Chee-Sanford et al., 2008).

Between 1982 and 2002, the number of large farms raising any kind of animal in the

United States increased by 234% (GAO, 2008). These large farms are also known as concentrated animal feeding operations (CAFOs). In order to be considered a CAFO, a farm must raise at least 1,000 cattle, 2500 large hogs, or 125,000 broiler chickens.

However, many large farms greatly exceed these minimum standards. Based on a survey of chicken producers, 8.44 billion chickens were raised on approximately 17,400 farms in

2006, correlating to an average of over 483,000 chickens per farm (MacDonald, 2008).

In addition to size, the farm must keep livestock in a confined environment for 45 days or

1 more in each year and livestock are not allowed to forage on the property in order to be considered a CAFO (GAO, 2008).

Because large numbers of animals are kept in confined spaces on CAFOs, veterinary pharmaceuticals (VPs) are commonly used to treat and prevent illnesses, which can spread easily in the crowded conditions, as well as accelerate growth.

Antibiotics are currently approved for four different purposes in the United States: disease treatment, disease control, disease prevention, and growth promotion. Both disease control and prevention involve giving medication to animals that do not exhibit clinical signs of disease in order to prevent possible infections. Growth promotion, which is also known as feed efficiency, involves treating healthy and growing animals with pharmaceuticals in order to rapidly increase weight gain (GAO, 2011). A significant portion of VPs are now used exclusively as growth promoters (GPs); approximately 11.2 million kg of antibiotics were administered as non-therapeutic feed additives in one year alone (Mellon et al., 2001)

Steroid hormone implants are widely used as GPs in the United States, and have been approved for this purpose since the 1950’s. As of 1989, it was estimated that 95% of cattle in the United States received steroids as GPs at some point in their life cycle

(Kuchler et al., 1989). Steroid hormones in cattle can improve weight gain by 5-20%, feed efficiency by 5-12% and lean meat growth by 15-25%. It is estimated that these improvements in growth rate results in the production of 2 billion additional pounds of beef each year (Kenney et al., 1989). Hormones are not currently approved as growth promoters for swine or poultry, meaning that a large portion of their usage is dedicated

2 towards cattle. Both naturally and synthetically produced hormones are currently approved for use as growth promoters. Natural hormones include estrogen, progesterone, and testosterone, whereas synthetic hormones include trenbolone acetate and zeranol

(FDA, 2011). Because of their ability to mimic the effect of natural hormones and impact the endocrine system in mammals, these steroid hormones are also classified as endocrine disrupting compounds (EDCs).

1.2 Background

The widespread usage of GPs has led to concerns about their environmental impact. Both hormones and antibiotics have limited bioavailability, meaning that a fraction of the compound will pass through the target animal without being metabolized.

This fraction will instead be excreted in either its original form or as an active metabolite.

The bioavailability of several different classes of antibiotics used as GPs varies widely

(9-75%) for swine and poultry (Anadón et al., 1999). This means that up to 90% of the antibiotic could be excreted from the animal unaltered.

Although many EDCs are already found in animal waste due to in vivo production, the use of additional growth promoters increases the concentration of these compounds found in their waste. It has been estimated that in 2000, farm animals raised in the United States excreted 330 tons of hormones (Lange et al., 2002). The bioavailability of hormone GPs ranges between 35-91%, resulting in the excretion of a significant fraction of parent compound (Lange et al., 2002). The sum of several different classes of antibiotics has been found to approach 1 mg l-1 in swine manure

(Campagnolo et al., 2002). Similarly, both naturally and synthetically produced EDCs have been detected in the manure of several different species of animals at elevated concentrations (Sarmah et al., 2006(1)).

Because of the high concentrations of GPs in animal waste, the disposal of animal manure thus presents the opportunity for the introduction of these compounds into the environment. For traditional farming systems, introduction of VPs to the environment can occur via direct deposition of animal wastes on pasture or grazing lands. However, because animals in CAFOs are kept in confined areas, their waste is collected and stored in holding tanks or lagoons. The large number of animals on these farms can result in the production of huge volumes of waste. For example, a dairy farm meeting the minimum number of cows to be considered a CAFO (700 heads) could produce almost 18,000 tons of manure annually. Far larger industrial farms, such as those used in beef cattle production (~ 140,000 heads of cattle,) can produce over 1.6 million tons of manure annually (GAO, 2008). Leakage from these manure storage tanks provides a route of entry for GPs into the environment. In Iowa, storage lagoons at least 7 acres in size are legally permitted to leak up to 16 million gallons of waste annually (Osterberg et al.,

2004). Both antibiotics and antibiotic-resistant microbes have been detected in surface waters and ground waters near storage lagoons (Osterberg et al., 2004). In addition, EDC concentrations in both the sediment profile and groundwater below a dairy farm storage lagoon were found in concentrations over 150 ng kg soil-1 (Arnon et al., 2008).

After storage in waste lagoons, the animal waste from CAFOs is frequently applied to crop fields as fertilizer. For example, roughly 39% of poultry litter was

4 applied by farmers to their own fields, with an average application rate of 1 acre fertilized by waste from 4000 birds (MacDonald, 2008). The application of poultry litter has been found to contribute to the concentration of an EDC in surface waters due to runoff from the field during storms (Nichols et al., 1996). Likewise, runoff from agricultural fields after the application of manure as fertilizer has been attributed to elevated levels of antibiotics in surface waters (Kay et al., 2005; Kuchta et al., 2009). The transport of GPs to surface waters is prevalent, with levels of both antibiotics and hormones found in surface waters across the United States in concentrations up to the µg l-1 range (Kolpin et al., 2002).

GPs in surface waters can impact aquatic ecosystems. For example, laboratory studies have found that EDCs contained in poultry litter have the ability to induce effects, which can alter the sex ratios of fish populations (Yonkos et al., 2010). Various EDCs have been measured in surface waters at concentrations which have been shown to affect the reproductive system of fish in laboratory experiments (Mills et al., 2005). Laboratory tests have shown that antibiotics could have toxic effects on aquatic macrophytes at environmentally relevant concentrations (Brain et al., 2005).

In addition to being transported to surface waters, some fraction of GPs can also be retained in the soil after the application of animal waste as fertilizer. Both antibiotics and hormones have been found to strongly adsorb to the organic or mineral fraction of soils. For many compounds, this is due both to the hydrophobicity of GPs and structure dependent interactions between the compound and various soil fractions e.g., organic matter, mineral phase, clay fractions (Tolls, 2001). This sorption can result in the

5 accumulation of GPs in soils, particularly with multiple applications of manure as fertilizer (Kemper, 2008). Transport of soil particulates (to which GPs are sorbed) to surface waters can result in an additional transport pathway (Pederson et al., 2003).

Antibiotics can still retain antimicrobial properties in the sorbed phase, thereby negatively impacting soil microbial communities (Chandler et al., 2005). Adsorption can shield GPs from microbial degradation, although biodegradation of such compounds does frequently occur (Ingerslev et al., 2001(1)). Both the antimicrobial effect on some species and the ability of other species to use GPs as substrates can result in changes in the microbial community structure of soils (Thiele-Bruhn et al., 2005; Westergaard et al.,

2001). In addition to biodegradation, VPs can also be transformed in soils via abiotic processes. These processes, which include hydrolysis, oxidation-reduction, and substitution reactions, are frequently mediated by the soil inorganic fraction (Thiele-

Bruhn, 2003; Colucci et al, 2001; LaKind et al., 1989; Chun et al., 2006(1)). EDCs are readily transformed in soil systems, both through biotic and abiotic processes. However, they do not appear to undergo complete mineralization, meaning that a fraction could be converted to products which retain endocrine-disrupting capabilities (Pan et al., 2009).

In addition to affecting microbial communities, antibiotics can be taken up from soils into plant tissues (Kumar et al., 2005; Boxall et al., 2006). Although only small levels are taken up, there are concerns about the impact on human health because this process has been observed in crops grown for human consumption. There appears to be little uptake of hormone EDCs from soil into plants, although some EDCs can affect the symbiotic bacteria found in the nodules of alfalfa plants (Shore, 2009). Due to the various

6 transformation and transport processes described above, the half lives of GPs in soils vary greatly. The half lives of antibiotics in soils ranges from 1 to 180 days, and are largely dependent on adsorption and microbial processes (Thiele-Bruhn, 2003). Because they are more labile, hormone GPs appear to undergo faster degradation, with half lives estimated to range from less than 1 to 30 days (Ying et al., 2002).

1.3 Compounds of interest

While the environmental fate of many GP compounds is well known, others are less well studied. This work focuses on two compounds whose fate in soils is poorly understood. One compound I studied was tylosin, a member of the macrolide class of antibiotics. Macrolides are clinically important antibiotics which are used to treat infections caused by Gram-positive bacteria in both humans and animals. Although macrolides are considered a critically important class of antibiotics for human medicine, both tylosin and erythromycin are currently permitted for use as VPs. Tylosin is approved for use as a growth promoter for both swine and poultry, and is also approved to treat illnesses in the latter (GAO, 2011).

All macrolides have a central 14-, 15- or 16-member lactone ring with one or more sugar groups and additional functional groups bonded to it (Gaynor et al., 2003).

Although tylosin has four forms, all four contain a 16-member central lactone ring, with variations in functional groups (Figure 1.1). Tylosin is found as a mixture of all four structures, with approximately 80-90% of the total comprised of tylosin A (Sarmah et al.,

2006(2)). For the remainder of this paper, the mixture of all four compounds will solely

7 be referred to as tylosin unless otherwise specified. Tylosin A has an amino-containing sugar (mycaminose) and two neutral sugars (mycarose and mycinose). The charged form of tylosin is due to the protonation of the amine group in mycaminose, which occurs at pKa = 7.2 (Boxall et al., 2006). Macrolide molecules derive their antimicrobial properties from their ability to bind to the large ribosomal subunit and stop cell growth by inhibiting protein synthesis (Gaynor et al., 2003). This binding is largely mediated by specific interactions between the lactone and the ribosome, although hydrophobic partitioning also plays a role in the binding process. The hydrophobic partitioning of tylosin is limited because of its low octanol-water partition coefficient (log Kow=1.63; [Boxall et al., 2006]). This also correlates to a limited bioavailability; for swine; only ~22% was taken up after oral administration (EMEA, 1997).

The low bioavailability of tylosin in swine is important because swine are commonly treated using macrolide antibiotics. In a survey of U.S. pig producers in 2000,

63% of farms indicated that they used antibiotics in feed for growth promotion purposes

(NAHMS, 2000). Tylosin was the most common antibiotic given to adult pigs in feed and was used by 56.3% of producers. In addition, 30.7% of farms reported administering tylosin to swine via injection and 4.1% reported providing tylosin in the pigs’ water supply (NAHMS, 2000). This is in stark contrast to Europe, where the European Union

(EU) banned the use of tylosin as a growth promoter in 1999. This ban, which limited the use of five additional compounds, was put in place because of their similarity to medically important antibiotics used to treat humans (Stein, 2002).

Because of its widespread usage and limited bioavailability, tylosin is frequently found in animal manure. Tylosin will readily adsorb to manure because of its organic composition (Loke at al., 2002). This sorption could prevent the biodegradation of tylosin; however, microbial degradation under both aerobic and anaerobic conditions

(due to storage of manure in lagoons) has been reported (Kolz et al., 2005). Despite both degradation and sorption, residual tylosin was still detected in manure slurries after eight months. Once this manure is applied as fertilizer, it is likely that the residual tylosin could be transported to surface waters and soil communities.

In a survey of surface waters in the United States, tylosin was found in 13.5% of streams sampled with an average concentration of 0.04 µg l-1 and a maximum detected concentration of 0.28 µg l-1 (Kolpin et al., 2002). In surface water under laboratory conditions, tylosin A undergoes rapid isomerization followed by slower degradation, with a half life of approximately 200 days (Hu et al., 2007). However, it is feasible that exposure to natural sunlight would result in a shorter half-life because the experimental apparatus filtered out the predominant wavelengths at which tylosin absorbs (Hu et al.,

2007). The isomerization product, which was hypothesized to be γ/δ-cis-tylosin, has decreased antibiotic activity compared to tylosin, meaning that it is less likely to impact environmental microbial communities (Werner et al., 2007). In addition to abiotic photodegradation, tylosin can undergo biotic degradation in surface waters with an average lag time of 40 days and a half life ranging from 24-54 days (Ingerslev et al,

2001(1)). This lag time is likely necessary for the growth of tylosin-resistant microbial species. A cyanobacteria species (Microcystis aeruginosa) was found to be sensitive to

9 aqueous tylosin concentrations above 0.034 mg l-1, whereas a green algae species

(Selenastrum capricornutum) was less sensitive with toxicity occurring above 1.38 mg l-1

(Halling-Sørenson, 2000).

Tylosin has also been found to have an effect on soil microbial communities. The addition of tylosin to soils resulted in a change in the number and types of soil microbes present (Westergaard et al., 2001). Westergaard and co-workers (2001) found that the soil microbial populations initially decreased and then increased. This pattern is likely due to the death of some microbes after the tylosin application, with the subsequent degradation of tylosin by resistant species. The number of fungi did not change a week after the addition of tylosin, suggesting that the tylosin had no direct effect on fungi.

However, the fungal biomass increased after 7 d, likely due to a decrease in competition for nutrients as some bacteria are killed by the tylosin. The structure of the microbial community remained changed after 2 months, likely due to the increased presence of tylosin-resistant species. A greater abundance of tylosin-resistant bacteria have also been found in agricultural soils which have been treated with waste from animals who are treated with GPs (Onan et al., 2003). The proportion of tylosin-resistant bacteria were found to be significantly lower in soils collected from farms which only used antibiotics for therapeutic purposes (~2%) as compared to soils collected from farms where tylosin was used as a GP (~6%). Far higher proportions of resistant bacteria were found in the manure of animals which were treated sub-therapeutically, with 30% and 69% of species showing resistance in cattle and swine manure, respectively.

In addition to biotic processes, tylosin undergoes abiotic processes such as sorption in soil systems. Tylosin has been found to adsorb strongly to soils regardless of organic carbon content. While tylosin undergoes some hydrophobic partitioning, the observed sorption has been non-linear, suggesting that it also undergoes specific interactions with the soils (Sassman et al., 2007; Rabølle et al., 2002; Zhang et al., 2011).

These specific interactions have been hypothesized to include cation exchange processes between the charged species and functional groups in the organic matter or clay fraction of soils. In addition, tylosin has been found to adsorb to two clays within the mineral fraction of soils (Essington et al., 2010). This sorption was hypothesized to be due to specific interactions between tylosin and hydroxyl groups on the surface of the clays.

The environmental fate of progesterone, a naturally occurring female steroid hormone (NIEHS, 2005) is even less well understood than that of tylosin. Like all naturally produced steroid estrogens, progesterone is composed of four ring groups, one cyclopentane and three cyclohexane rings (Figure 1.2). Progesterone has two ketone functional groups and two methyl groups which affect its behavior. Because of a lack of proton-accepting functional groups, progesterone remains neutral at all pH values.

Progesterone is far more hydrophobic than tylosin, with log Kow=3.87 (Neale et al.,

2008). Despite its increased hydrophobicity, progesterone also has limited bioavailability, with less than 10% uptake in cattle after oral administration of the drug

(EMEA, 1999). Progesterone is naturally produced by all mammals, with far higher levels produced by females. Because it is a hormone, progesterone is commonly used in humans as an intrauterine contraceptive and to treat endocrine-related diseases (NIEHS,

2005). While EDCs are not currently approved for use as growth promoters in swine or poultry, progesterone can be used to control ovulation in livestock (NIEHS, 2005). In cattle, progesterone is commonly used in a 10:1 ratio in conjunction with 17ß-estradiol as a growth promoter (Song et al., 2000). This combination of hormones has been found to increase the rate of daily weight gain in cattle by 4.6-14.0%. In addition, 200 mg of progesterone with 20 mg of 17ß-estradiol improved the ratio of food consumed to weight gained by 1.8% (Song et al., 2000).

Like tylosin, the low bioavailability of progesterone means that it will be excreted by livestock and can enter the environment. However, because of natural production, it is difficult to determine what fraction of progesterone in the environment results from the use of GPs. For example, progesterone was found in the waste of poultry which had not been treated with GPs due to in vivo production of the hormone (Lorenzen et al., 2004).

In addition, progesterone was detected in the manure of both cattle which were not treated with any GPs and those which received other EDCs but not progesterone (Bartelt-

Hunt et al., 2012). Progesterone was also detected in runoff from feedlots for cattle which received EDCs, with an average concentration of 59.5 ng l-1 and a maximum concentration of 570 ng l-1 (Bartelt-Hunt et al., 2012). Under aerobic conditions in manure, testosterone was found to completely degrade within 12 h, after a lag phase of 5-

9 h (Borch et al., 2009). Despite this rapid degradation, progesterone was detected in 5% of streams within the rangeland of roaming beef cattle, with a maximum concentration of

27 ng l-1. The elevated level of progesterone (and other EDCs) in the streams was attributed to input from cattle manure (Kolodziej et al., 2007). In a general survey of

12 streams in the United States, progesterone was also found in 4.3% of streams, with an average concentration of 0.11 µg l-1 and a maximum concentration of 0.199 µg l-1 (Kolpin et al., 2002).

Limited work has been completed regarding the effect of progesterone on aquatic ecosystems. Progesterone can undergo photolysis at wavelengths higher than 340 nm, with a half-life of 79 h (Borch et al., 2009). When present in aquatic systems, progesterone is likely to have effects similar to those of other EDCs regarding toxicity and effects on reproductive organ development. It has been suggested that progesterone could have an effect on gonad development and function similar to that of other progestogins (i.e. 17ɑ-hydroxy progesterone) (Kolok et al., 2008). Progesterone has also been found to induce production of female reproductive proteins in immature female rainbow trout; however, its potency was nearly 1000 times less than that of other EDCs

(Kolok et al., 2008).

Similar to the lack of knowledge regarding the aquatic fate of progesterone, little is known regarding the fate and impact of progesterone in soils. To the best of my knowledge, no work has been completed to determine the sorption behavior of progesterone in soils. Progesterone has been found to adsorb strongly to isolated soil organic matter fractions and analogs (Neale et al., 2008). However, these analogs serve as a poor representation of natural soils, and thus provide limited information regarding the behavior of progesterone in natural systems. Although no estimates of half-life for progesterone in soils have been calculated, 70% of the progesterone spiked into soil samples remained after 25 days (Hu et al., 2011). Hu and coworkers (2011) added

13 progesterone to soils in order to investigate its effect on soil nematode communities, and they found that progesterone reduced the total number of soil nematodes and increased the male:female sex ratio.

1.4 Statement of objectives

Because of the varied impacts similar GPs can have on the environment, it is important to understand the environmental fate of progesterone and tylosin. Currently, little is known regarding their impact on soil ecosystems. The purpose of this work was to help elucidate the fate of these two compounds in sterilized soil systems. I first determined the kinetics and partition coefficients of sorption between both tylosin and progesterone and five soils (as described in Chapter 2). Based upon these results, I suggested likely mechanisms of interaction between both compounds and the soils which would affect their sorption behavior. Further, I investigated the interactions between tylosin and pure mineral oxides, as described in Chapter 3, based upon tylosin losses that could not be explained by sorption alone. The half lives and possible transformation products were determined for tylosin in contact with an iron and manganese oxide, two ubiquitous minerals present in nearly all soils. This work is essential for understanding the environmental fate and interactions of tylosin and progesterone in soils.

Figure 1.1. The chemical structure of the four different forms of tylosin (A-D).

Figure 1.2. The chemical structure of progesterone.

Chapter 2: Tylosin and progesterone sorption to soils.

2.1 Introduction

Over the past several decades, animal production in the United States has flourished. As of 2002, it was estimated that there were over 100 million cattle, 60 million swine, and 8.6 billion chickens in the United States (NASS, 2002). Many of these animals are now raised on large industrial farms, known as concentrated animal feeding operations (CAFOs), where the animals live in high-density conditions with little exposure to natural environments. For example, 690,000 producers raised 20 billion pounds of livestock in 1984. However, by 2000 only 95,000 producers raised 26 billion pounds of pork (Chee-Sanford et al., 2008). The number of CAFOs for all species of livestock in the United States increased by 234% between 1982 and 2002 (GAO, 2008).

The extensive use of veterinary pharmaceuticals (VPs) has helped facilitate the growth of CAFOs; it is estimated that 18-29 million pounds of antibiotics are used to treat animals in the United States in a single year (Anderson et al., 2003). Although many VPs are used to treat illnesses, growth promoters (GPs) are antibiotics or hormones used to enhance growth rates and/or prevent illnesses in animals living in CAFOs.

Because of their ability to mimic naturally produced compounds and interfere with

17 endocrine signaling, these hormones are commonly classified as endocrine disrupting compounds (EDCs) (Card et al., 2012). The use of GPs is extremely common; approximately 11.2 million kg of antibiotics were administered as non-therapeutic feed additives in one year alone (Mellon et al., 2001). Likewise, almost 90% of all cattle in the United States received at least one implant throughout their lifespan which contained growth promoting hormones (USDA, 2000). The widespread use of GPs is of environmental concern because many have limited bioavailability, meaning that a significant fraction of the compound will pass through the animal and be excreted either in its original structure or as a bioactive metabolite. For example, both tylosin and progesterone have limited bioavailability; tylosin is ~22% bioavailable when orally administered to swine (EMEA, 1997) and progesterone is ~10% bioavailable in cattle when orally administered (EMEA 1999). Elevated levels of many pharmaceuticals have been detected in animal waste. For example, the sum of several different classes of antibiotics found in swine manure slurry samples was found to approach 1 mg l-1

(Campagnolo et al., 2002). Elevated levels of EDCs, both naturally and synthetically produced, have also been found in the manure of several different species of animals

(Sarmah et al., 2006). The bioactive fraction which is excreted and present in manure can then enter the environment through direct or indirect pathways (Figure 2.1, [Kemper,

2008]).

For freely roaming animals, direct introduction of VPs to the environment can occur via deposition of animal wastes in pasture or grazing lands. However, for animals raised on CAFOs, manure is frequently collected and stored in large lagoons or surface

18 ponds instead of being distributed naturally. The stored manure is often applied to crop fields as a fertilizer, resulting in the widespread introduction of VPs to the environment

(Song et al., 2010). After manure is applied to fields, the environmental fate of these bioactive molecules is largely controlled by the underlying soil matrix. The compounds can run off the surface of the field and enter surface water systems, leach through the soil environment to enter groundwater systems, be taken up by plants (unless the manure was applied pre-emergence) or be retained in the soil compartment (Burnison et al., 2003;

Arikin et al., 2008; Kay et al., 2005; Boxall et al., 2006).

Because soils have such a profound effect on the transport of these compounds, it is important to understand the environmental fate and impact of these compounds in soil ecosystems. Several different classes of antibiotics which are used as GPs have been found in soils, including macrolides (of which tylosin is a member), tetracyclines, sulphonamides, and fluoroquinones (Kemper, 2008). These compounds have been found in agricultural soils in concentrations ranging from 450-900 µg kg-1 for tetracyclines, 13-

67 µg kg-1 for macrolides, and 6-52 µg kg-1 for fluoroquinones (Thiele-Bruhn, 2003). In addition, tetracyclines and their transformation products can be detected in soils for months after the application of manure containing these compounds, suggesting that soil could serve as an environmental reservoir for antibiotics (Aga et al., 2005). The effects of these compounds on soil microbial communities are unclear, with some studies reporting suppressed microbial growth and others finding increased rates of metabolism (Thiele-

Bruhn, 2005). However, there are fears that prolonged exposure of soil microbes to low concentrations of antibiotics could lead to the development of resistant microbial

19 populations (Khachatourians, 1998). An additional transport pathway has been observed via the uptake of veterinary medicines from soils into plants (Boxall et al., 2006).

Similarly, EDC compounds have also been detected at high levels in both soil and soil water runoff for up to three months after the application of agricultural manure (Kjaer et al., 2007). Under anaerobic conditions, several EDCs have been found to be resistant to microbial degradation, leading to persistence in the soil (Ying et al., 2005). However, under aerobic conditions these compounds are readily degraded and could possibly change the soil microbial community structure due to the input of metabolic products

(Ying et al., 2005; Chun et al., 2006(2)). The transport of organic contaminants in soils will affect their environmental impact on both soil and aqueous ecosystems; thus, it is important to understand processes which control transport.

The sorption of organic contaminants to soils plays a central role in controlling their transport to aqueous ecosystems. Although sorption can be due to several different mechanisms, the overall observed sorption to soils can be described using a partition coefficient. This partition coefficient, Kd, is a ratio of the sorbed and aqueous concentrations of a contaminant after it has reached equilibrium with a soil (equation 1)

(1)

One of the dominant mechanisms of sorption for many compounds is hydrophobic partitioning of the GP into the organic fraction of soils (elaborated on below). Because hydrophobic partitioning can play such a large role in sorption, normalizing for the fraction of organic carbon present in a soil (resulting in the normalized partition

20 coefficient Koc, equation 2) can help minimize variability in observed Kd values (Tolls,

2001).

(2)

The sorption of VPs to soils, as described by Kd or Koc, is often due to several different, and sometimes competing, interactions. As mentioned above, one of the most dominant contributors to soil sorption is the hydrophobic partitioning of organic contaminants onto soil organic matter (SOM). Such partitioning occurs because of an increase in the free energy of the system as the hydrophobic organic contaminant sheds the water molecules which surround it in an aqueous system and instead interacts with organic sorbents via non specific interactions such as van der Waals interactions

(Schwarzenbach et al., 2003). Previously, this partitioning has been related to similar processes such as the distribution of a compound between water and an immiscible organic solvent. One such “proxy” parameter is the octanol-water partition coefficient

(Kow). Correlations between the two partition coefficients have been calculated using a linear relationship of the form log Koc= a log Kow + d, where a and d are regression coefficients (see Baker et al., 1997 for a summary of several studies). Baker and coworkers developed a general equation based upon observed measurements of Kow and

2 Koc (log Koc= 0.903 log Kow + 0.094; R =0.91) in order to predict Koc values without experimental validation. However, this equation, which is commonly known as a single parameter linear free energy relationship (sp-LFER), is only valid for compounds with

Kow ~ 1.7-7.0. Below this range, hydrophobic partitioning is likely less important and

21 above this range almost all of the compound will be in the non-aqueous phase, leading to difficulties in an accurate measurement of the partition coefficient. In addition, this sp-

LFER assumes that the free energies of transfers for both octanol-water and SOM-water are linearly related and additive (with respect to functional groups and molecular fragments). However, specific molecular interactions between a compound and one of the sorbent phases could contribute different interaction energies, resulting in an overall non- linear relationship (Goss et al., 2001). The energies of these interactions, including more specific van der Waals interactions (beyond London Forces) and H-bond interactions, are accounted for in poly-parameter models i.e., pp-LFERs. These more complex models can accurately predict adsorption when more than one type of interaction (i.e., both van der Waals and H-bond interactions) occur, which is likely in complex soil systems. In addition, pp-LFERs are advantageous because they can accurately describe the behavior of several compound classes with different interactive groups (i.e., phenols, amines, etc) using a single equation.

The specific interactions accounted for in pp-LFERs can occur with different types of organic matter. For example, positively charged organic compounds have been shown to form complexes with the carboxyl and phenol groups of dissolved organic matter (DOM) (Arnold et al., 1998). These same complexes are likely to form with SOM because it contains the same types of functional groups as DOM. These interactions, which involve the exchange of a protonated contaminant with a metal cation species in complex with a negatively charged organic matter functional group, are related to the

22 observed correlation between positively charged organic compounds and the cation exchange capacities of soils (Sassman et al., 2005).

Molecules can also sorb to the mineral phase of soils, including clays and metal oxide minerals. Organic contaminant sorption to clay surfaces has been found to be related to surface area, and depending on size, some compounds will sorb to the interlayer spaces of expanding clays in addition to outer faces and edges (Tolls, 2001).

Because clay minerals are often negatively charged due to isomorphic substitution, they can undergo cation exchange processes with organic contaminants (Gu et al., 2005(1)).

Several different classes of organic compounds have been found to adsorb to metal oxides found in the mineral phase of soils (Gu et al., 2005(1); Figueroa et al., 2005; Van

Emmerik et al., 2003). Adsorption to metal oxides has frequently been found to be pH dependent, suggesting that specific interactions between charged functional groups play an important role.

Sorption of organic contaminants to soils can also be affected by variations in soil composition. Weber and coworkers (1992) described the distributed reactivity model

(DRM), in which sorption onto several different heterogeneous solid components contributed to the observed overall soil sorption. For example, they found that the extracted shale fraction of their soil exhibited a far different reactivity and sorptive behavior than the bulk soil. In addition to varying solid composition, sorption can also be affected by the composition of soil organic matter. SOM exists as a gradient between two phases- ‘rubbery’ or ‘amorphous’, in which the interactions between humic molecules are weak and transitory; or ‘glassy’, in which interactions between molecules

23 are more rigid and are relatively longer-lived (Pignatello, 1998), the latter being similar to the “shale” component of soils. Sorption of organic contaminants to the rubbery phase of SOM is commonly assumed to occur by a dissolution mechanism, which can be described using a linear model. However, organic compounds can also diffuse through and sorb to the longer-lived internal pores formed within the glassy phase of the SOM.

Sorption to pores is likely competitive due to limits in size and availability, thus contributing a non-linear portion to the overall sorption behavior of a contaminant.

Because of the limited bioavailability and wide usage of tylosin and progesterone,

I am interested in the sorption of these compounds to soils. Past studies have looked at the sorption of tylosin to soils (Rabolle et al., 2000; Sassman et al., 2007; Ter Laak et al.,

2009; Zhang et al., 2011). However, they have observed widely varying Koc values, ranging over approximately two orders of magnitude. In addition, several different mechanisms have been hypothesized that attempt to explain tylosin sorption, including hydrophobic partitioning and specific interactions between charged tylosin species and surface exchange sites. There appears to be little previous work completed regarding the sorption of progesterone to soils, with a single Koc value published based upon a limited set of field observations (Lopez de Alda et al., 2002). The aim of this work was to experimentally determine the sorption of tylosin and progesterone to five different soils.

Based upon the observed sorption behavior, I suggest possible mechanisms of interaction between both compounds and soils. These results can help provide important insights into the mobility of these compounds in soils, which plays a central role in their environmental fate.

2.2 Methods

Partition coefficients were measured for both target compounds with five different soils. Both the Drummer 36 and Coloma 32 soils are soils from Indiana and their collection has been described previously (Khan et al., 2009; Li et al., 1999). The Finley control soil is a soil from Ohio and details regarding its properties can be found in Card et al., 2012. Two additional soils were collected from the tundra in Alaska in July 2011.

One of the soils (Alaskan Burned), was collected from a site (GPS coordinates:

66o38’56” N, 150o09’52” W) which had been subjected to a wildfire two years previously; the other (Toolik control) was collected from a site (GPS coordinates

68o38’13” N, 149o37’52” W) which had not been affected by wildfires and is actually a peat (Table 2.1). At both sites, plant growth and surface soil were removed and the topsoil was collected to a depth of approximately 0.2 m. Topsoil samples were collected from six to eight randomly selected sites within a 5 m x 5 m grid and mixed to ensure homogeneity. After mixing, the samples were air dried and passed through a 2-mm sieve to remove large plant debris and rocks; the soils were then stored at 4 oC. The soils were analyzed by A&L Great Lakes Laboratories, Inc. (Fort Wayne, IN) to determine their physical and chemical properties (Table 2.1).

Table 2.1. Physical and chemical soil properties. Soil Origin ƒoc pH Sand:Silt: Clay CEC (%) (%:%:%) (meq 100 g-1) Coloma 32 Indiana 1.1 5.9 88:7:5 4.3 Drummer 36 Indiana 4.0 7.5 17:47:36 15.5 Finley Control Ohio 7.7 7.0 17:46:37 20.2 Alaskan Burned Alaska 7.1 6.5 39:31:30 18.5 Toolik Peat Alaska 75.6 6.4 47:34:19 28.1

The soils were also analyzed to determine the concentration of iron and manganese oxides present. Two soils (Alaskan Burned and Toolik Peat) were analyzed by A&L

Great Lakes Laboratories, Inc. and were normalized to the remaining soils, which were analyzed via X-ray fluorescence (XRF) using a PANalytical MagiX Pro PW2440 XRF spectrometer. Bulk soil samples were fused into beads with lithium tetraborate flux

(Chemplex, Palm City, FL) at a ratio of 1:4 (w:w) sample/flux with a Philips Perl'X bead maker using a platinum crucible and casting dish. The samples were analyzed under vacuum using a rhodium X-ray tube and K(ɑ) radiation. Instrument parameters were set to 57 and 60 kV and 70 and 66 mA for Mn and Fe, respectively.

For all experiments, soil (1.0 g and 0.1 g dry weight for tylosin and progesterone, respectively) was added to a 50-mL Corex centrifuge tube and brought to field capacity using Milli-Q water (Elix 10 reverse osmosis and Milli-Q UV Plus, Millipore, Billerica,

MA). Due to the much higher ƒoc value for Toolik peat, 0.5 g dry weight of soil was added to each Corex tube for tylosin experiments in order to ensure that the aqueous analyte concentrations were above the limit of detection after quasi-equilibrium was reached. The soils were then incubated for 24 hours in the dark at room temperature before being autoclaved at 121oC (Tuttnauer Brinkmann 2340M, Tuttnauer USA Co.

Ltd., Hauppauge, NY) for 1 h. This process was repeated twice more, for a total of three autoclave sessions, in order to minimize microbial degradation of the analytes.

Autoclaving has been shown to have minimal effect on the sorption behavior of organic contaminants (Lotrario et al., 1995). All solutions were made in 5 mM CaCl2 (EM

Science, Gibbstown, NJ), which was autoclaved at 121o C for 1 h. Calcium chloride was used to maintain a constant ionic strength and to replicate natural soil pore water conditions. Tylosin soil sorption is stable at ionic strengths below ~0.1 M, so the addition of CaCl2 should have negligible effects (Ter Laak et al., 2006). Although no studies have included progesterone, changing ionic strength had minimal effect on the soil sorption of similar hormone compounds (Lai et al., 2000). For all experiments, respective experimental blanks and controls, which contained either CaCl2 with soil or analyte solution without soil were prepared using the above procedure. Stock solutions were made by dissolving tylosin tartrate (used as received, MP Biomedicals, LLC.,

Solon, OH) or progesterone (>99% purity, Sigma Aldrich, St. Louis, MO) in HPLC grade methanol (J.T. Baker, Center Valley, PA) to a concentration of 2 mM. Stock solutions were kept in the dark at room temperature for no longer than one month.

For the kinetics experiments, an analyte solution was made by diluting the stock solution with autoclaved CaCl2 to a concentration of 3 µM. Ten ml of this solution was added to each 50 ml Corex tube to which the amount of soil listed above was added

(except for experimental controls which only received CaCl2). The soil-solution slurries were then incubated with occasional manual shaking in the dark for various time intervals

(with two samples sacrificed at each time point) from 1 h to 7 d until quasi-equilibrium

27 between the sorbed and aqueous phases was reached. I define quasi-equilibrium as an invariant change in the sorbed concentration relative to the time scale of the experiment.

It is necessary to define quasi-equilibrium because true equilibrium could take months to reach due to slow sorption into the glassy domain of the organic fraction of our soil

(Pignatello et al., 1996).

For sorption isotherm experiments, the analyte stock solution was diluted using autoclaved CaCl2 to six different nominal concentrations within the range of 0.5-5 µM.

Ten mL of each solution was added to two replicate samples (except for experimental controls). Where standard deviations are included, the depicted data represents the results of three trials of an experiment. The soil-solution slurries were incubated with occasional shaking in the dark for 24 h and 7 d for progesterone and tylosin samples, respectively. Low concentration soil sorption isotherms for progesterone were also completed using a similar procedure, except with the addition of 0.4 g dry weight of soil and 40 mL of solution to each 50-mL Corex tube and initial nominal aqueous concentrations ranging from 0.05-0.5 µM. After a 24-h incubation period with shaking, the samples were filtered as described below and the filtrate was passed through an

Empore C18 SPE disk (3M Co., St. Paul, MN), which had been preconditioned with 20 ml of methanol and 100 ml of Milli-Q water. Progesterone was then eluted from the disk using 5 ml of methanol, which was analyzed by reverse phase high pressure liquid chromatography (RP-HPLC). Details are provided below.

After incubation, each sample was centrifuged for 30 min at 1600 rpm and 23oC

(Beckman GS-6R, Beckman Coulter Inc., Brea, CA). The supernatant was filtered

28 through a 0.6 µm pore size glass fiber filter (Sterlitech, Kent, WA) before analysis via

HPLC-UV/vis. Unless explicitly stated that mass balances were completed, sorbed concentrations were calculated by difference based upon the initial and aqueous concentration after quasi-equilibrium was reached and were not directly measured. To complete mass balances, the supernatant was removed, the residual solution was determined gravimetrically and then the soil underwent triplicate extraction with 4 mL of methanol and a 2-h incubation time. The methanol supernatant was filtered and analyzed using RP-HPLC.

The supernatant was analyzed using reverse phase high-performance liquid chromatography (either a Waters 1515 or 1525 pump and 717plus autosampler, Waters

Corp., Milford, MA), with a UV-vis detector (Waters 2487 dual λ absorbance detector).

The mobile phase for progesterone was 75:25 (v/v) methanol/Milli-Q, and for tylosin

50:50 (v/v) 5 mM phosphoric acid (ACS reagent grade, Acros Organics, Morris Plains,

New Jersey) and acetonitrile (HPLC grade, J.T. Baker). The flow rate was 1.0 ml min-1 and injection volume was set at 100 µl. The column was a Restek Pinnacle DB packed with C18 (length, 150 mm; inside diameter, 4.6 mm; particle size 5.0 µm). Wavelengths of detection were 287 and 245 nm for tylosin and progesterone, respectively.

Wavelengths of detection were determined by observing absorbance maxima in the range of 200 – 800 nm (Cary 1 UV-Vis Spectrophotometer, Agilent, Santa Clara, CA).

2.3 Results and discussion

Kinetic studies

Quasi-equilibrium was reached between the aqueous and sorbed phase after 24 h and 7 d for progesterone and tylosin, respectively, as shown in representative plots

(Figure 2.2 and 2.3). I believe that this is the first data set reported for progesterone soil sorption kinetics. My data is within the time span (1-48 h) reported for similar hormone compounds to reach quasi-equilibrium with soils (Casey et al., 2005; Khan et al., 2009;

Lai et al., 2000). For tylosin, my kinetics show considerably longer times needed to reach quasi-equilibrium relative to those reported in other studies (24 – 60 h) (Rabølle et al., 2000; Sassman et al., 2007; Zhang et al., 2011). Rapid sorption within the first hour was observed both in this and other studies (Allaire et al., 2005; Ingerslev et al., 2001(2)), however, slower sorption continued for several days. It is possible that the previous studies could have underestimated the time required to reach equilibrium. Such underestimation could occur if small changes in aqueous concentration due to continued slow sorption are instead attributed to analytical error (Pignatello et al., 1996). In previous studies of other organic contaminants, although the aqueous concentration appeared to undergo a negligible change over a period of hours, longer contact increased the amount sorbed by up to an order of magnitude (Ball et al., 1991). It is likely that the large molecular size of tylosin would retard its diffusion through the glassy phase of the

SOM, requiring longer times for sorption within the micropores to reach equilibrium.

Sorption Isotherms

Progesterone

Soil sorption isotherms for progesterone were plotted using a linear relationship between the sorbed and aqueous concentrations after quasi-equilibrium was reached

(Figure 2.4). Partition coefficient, Kd, values were calculated based upon the slopes of the linear regressions for the isotherms (Table 2.2).

Table 2.2. Linear parameters and partition coefficients for progesterone with initial aqueous concentrations from 0.5-5 µM.

a b c Kd R Koc Log Koc Finley 42.7 0.960 555 2.74 Drummer 38.6 0.988 965 2.98 Coloma 11.2 0.986 1018 3.01 Burned 32.3 0.976 455 2.66 Toolik peak 161.3 0.988 213 2.33 a µmol sorbed µM aqueous-1 kg -1 b correlation coefficient for linear regression c L kg-1

Based upon the observed slightly non-linear shape of the sorption isotherms and the poor linear correlation coefficients for some soils, the data for the interactions of progesterone with all soils were also converted to the logarithmic transformation of the Freundlich equation (Figure 2.5).

Non-linearity of sorption isotherms suggests that more than one sorption mechanism contributes to the overall sorption of the analyte (Weber et al., 1992). This more complicated behavior is better represented by a Freundlich transformation of the

31 data, which is a mathematical model that accounts for processes with different energies of adsorption (i.e., different sorption mechanisms) (Chiou, 2002). Utilization of the

Freundlich transformation results in the introduction of two parameters, Kf and N, the

Freundlich adsorption coefficient and a measure of isotherm nonlinearity, respectively

(equation 3).

(3)

Typically, N values within the range of 1.0 ± 0.05 are indicative of linear sorption i.e., obey the simple partition coefficient expression (equation 1), while N values less than 1.0 suggest that at lower concentrations sorption sites show stronger affinity for sorbing molecules (Pignatello et al., 2006). For all soils, the sorption isotherms for progesterone were non-linear, with N ranging from 0.76 to 1.06 (Table 2.3).

Table 2.3 Freundlich parameters and partition coefficients for progesterone soil sorption isotherms.

a b c,d d Soil Kf N R Kd Koc Log Koc Coloma 10.52 1.06 0.995 10 920 2.96 Drummer 43.27 0.88 0.992 47 1176 3.07 Finley 52.80 0.82 0.991 59 777 2.89 Burned 41.98 0.76 0.998 49 696 2.84 Toolik Peat 175.35 0.85 0.995 195 258 2.41 a µM l kg-1 b correlation coefficient of linear regression c -1 determined at Caq= 0.5 µmol l d l kg-1

In an effort to compare sorption isotherms in a meaningful manner, concentration- dependent soil partition coefficients (Kd) were calculated for each soil using equation 4

-1 and Caq=0.5 µmol l , which was near the lowest aqueous concentration for all soils after equilibrium was reached.

(4)

It is important to use a small aqueous concentration because of the non-ideal behavior of organic contaminants at high concentrations (Chiou, 2002). The calculated Kd value was then used to calculate the organic carbon content normalized soil partition coefficient

(Koc) using equation 2, where ƒoc is the fraction organic carbon content of each soil. The

Koc values for progesterone were similar regardless of whether the linear or Freundlich transformation of the isotherms was used. Based upon the linear regressions, the average log Koc=2.84 ±0.17 whereas for the Freundlich transformation the average partition coefficient was slightly higher, with log Koc=2.94 ± 0.10 (where the error represents one standard deviation). The Koc values for Toolik peat were not included in calculations of the average because of the extremely high fraction of organic carbon as compared to the other four soils tested.

Because of the observed non-linearity at higher concentrations, sorption isotherms were completed at lower concentrations (initial Caq=0.05-0.5 µM) with three soils. The sorption isotherms at lower concentrations were more linear for the Finley and Burned soils than the isotherms for the same soils at higher concentrations, based upon a comparison of their correlation coefficients (Figure 2.6, Table 2.4). It is possible that the

Drummer soil does not follow this trend because the lowest data point (nominal initial

Caq=0.05 µM) was not included due to leakage of the filter assembly during extraction.

Additionally, sorption isotherms calculated for all concentrations tested were equivalently

33 or more linear than the isotherms for just the higher concentrations, based upon their respective linear correlation coefficients (Figure 2.7, Table 2.4).

Table 2.4. Linear correlation coefficients (R values) for progesterone over varying initial aqueous concentration ranges. Initial aqueous concentrations Low High All (0.05-0.5 µM) (0.5-5.0 µM) (0.05-5.0 µM) Finley 0.994 0.960 0.982 Drummer 0.968 0.988 0.994 Burned 0.996 0.976 0.999

Adsorption at low concentrations has previously been attributed to sorption to either high-surface-area carbonaceous materials (HSACM) or micropores within the glassy phase of SOM (Xia et al., 1999). Sorption to HSACM, which consists of charcoal or other hard carbon geosorbents e.g., shale like components described by Weber et al.

(1992) and is found in almost all soils, is possible because of its high surface area and the presence of specific high-affinity adsorption sites on its surface (Chiou, 2002). Linear behavior at lower concentrations could occur if there is an abundance of micropore sorption sites, allowing all progesterone molecules to adsorb without competition or through additional sorption mechanisms.

There is limited data available regarding the interaction of progesterone with soils. To date, no one has conducted sorption isotherms with progesterone and soils.

-1 Nonetheless, a Kd value of 204 l kg has been estimated based upon the ratio of the aqueous and sediment concentrations of progesterone along the Anoia and Cardener 34

River (López de Alda et al., 2002). This value is significantly higher than the Kd values calculated here (except for the Toolik soil, which is a peat and has an organic matter content more than an order of magnitude higher than the other soils). Moreover, their value is not a true “measured” partition coefficient as determined by a sorption isotherm, but is indirectly determined based upon a small set of field data (López de Alda et al.,

2002).

In contrast to the soil literature, the partition coefficients (Kom) between progesterone and several organic matter fractions and analogs (Aldrich humic acid, alginic acid, and tannic acid) were determined (Neale et al., 2008). At pH=7, log Kom ranged from 3.57-4.59, with the highest sorption observed for Aldrich humic acid. It is impossible to make a direct comparison of these Kom values with our partition coefficients because of their selection of organic matter fractions as compared to bulk soils. Commercially available humic acids (such as Aldrich humic acid) have been shown to differ significantly from natural organic matter in several important chemical properties, such as a higher composition of ash and much higher aromaticity and carboxylic acid content (Malcolm et al., 1986). Thus, commercially available humic acid is a poor substitute for natural organic matter (Chiou et al., 1987). It is also difficult to compare our results to these values because there are many other factors (i.e., clay content, soil structure) which influence sorption behavior in natural soils in addition to

SOM.

Although few attempts have been made to calculate Koc for progesterone, a larger body of work exists that has reported the sorption of similar EDC compounds to soils.

Concentration-specific Koc values were calculated for four similar EDC compounds (17β- estradiol [E2], estrone, melengestrol acetate, and ɑ-zearalanol) using the identical

Drummer soil used in this study (Card et al., 2012). My concentration specific Koc value for progesterone with Drummer soil (log Koc= 3.07) was within the range determined for these compounds (log Koc=2.96-3.41), indicating that progesterone and the other EDC substances interact with SOM components in a similar fashion. In addition to the E2 studied by Card et al. (2012), 17ɑ-ethynyl estradiol (EE2) and testosterone also exhibited non-linear behavior with soil from the Drummer series (N=0.62-0.77; Lee et al., 2003).

Partition coefficients were calculated using a linear model despite the observed non- linearity, making it difficult to compare their values to our progesterone partition coefficients calculated using the Freundlich model. Despite this discrepancy, both the Kd and Koc values for progesterone with Drummer 36 fall within the range calculated for

-1 these EDCs (Kd=23.4-83.2 l kg , log Koc=2.91-3.46). In addition to Drummer soils, non- linear behavior of EDCs was observed for several soils from New Zealand by Sarmah et al. (2008). Based upon the reported N values (N=0.73-1.1), they used the Freundlich

-1 model to calculate concentration specific Kd values of 14.08-235.94 l kg for E2 and

-1 EE2. Correcting for ƒoc and using a Caq=0.5 mg l , they calculated log Koc values of 3.12

± 0.24 and 2.90 ± 0.32 for E2 and EE2, respectively. My calculated log Koc values for progesterone for all soils (except for Toolik peat) fall within these ranges. As stated previously, sorption to Toolik peat is likely affected by its extremely high organic carbon content.

Based upon the structure and estimated octanol-water partition coefficient of progesterone (log Kow=3.87 [Neale et al., 2008]), we would expect hydrophobic partitioning of progesterone to be a dominant sorption mechanism even though the molecule possesses H donor/acceptor functional groups. However, it is likely that there is some other contributing sorption mechanism, as suggested by the non-linearity of our

Freundlich parameter at higher concentrations. Indeed, this is supported by the fact that sorption to Toolik peat has the lowest Koc value due to its very high organic matter content. SOM will readily adsorb to the surfaces of soil minerals, with preferential sorption to the specific active sites and micropores that are present (Kaiser et al., 2003).

The high organic fraction of the Toolik peat could out compete progesterone for these sites, thereby preventing progesterone from undergoing specific sorption.

Cation exchange reactions can be an important sorption mechanism, but it is unlikely that progesterone undergoes specific interactions with carboxyl or phenol groups because it is neutral at all pH values. As discussed above, adsorption at low concentrations has previously been attributed to sorption to HSACM, such as charcoal. If this were a dominant mechanism in our soil systems at low concentrations, we would expect to see significantly greater non-linear behavior with the Alaskan burned soil because of the higher fraction of HSACM that would be present as a result of the wildfires which recently occurred. However, the Alaskan Burned soil does not exhibit noticeably different behavior than the other soils, suggesting that another mechanism between progesterone and our soils controls sorption at the lower concentrations. One possible explanation for the observed non-linearity could be interactions between

37 progesterone and the interlayer spacing of clay minerals in the soils. Enrofloxacin, an antibiotic of a similar size as progesterone (MW=359.4 g/mol [Tolls, 2001]), has been shown to cause a concentration-dependent increase of the interlayer spacing of montmorillonite, with greater increases in spacing occurring at higher enrofloxacin concentrations (Nowara et al., 1997). In addition, the sorption of enrofloxacin to clays appears to be mediated by a carboxyl group, which can also be found in progesterone, further suggesting the likelihood of a similar interaction between progesterone and clays.

Tylosin

Soil sorption isotherms for tylosin were also plotted using a linear relationship between the sorbed and aqueous concentrations after quasi-equilibrium was reached

(Figure 2.8). Partition coefficient, Kd, values were calculated based upon the slopes of the linear regressions for the isotherms (Table 2.5).

Table 2.5. Linear parameters and partition coefficients for tylosin with initial aqueous concentrations from 0.5-5.0 µM.

a b c c Kd R Koc Log Koc Finley 80.6 0.965 391 2.59 Drummer 45.2 0.994 1130 3.03 Coloma 9.1 0.991 827 2.92 Burned 69.1 0.985 973 2.99 Toolik 80.6 0.966 107 2.03 a µmol sorbed µM aqueous-1 kg -1 b correlation coefficient for linear regression c l kg-1

The sorption of tylosin to all soils was also described by the Freundlich equation (Figure

2.9). Like progesterone, tylosin exhibited non-linear behavior with all soils, with N ranging from 0.65 - 0.81 (Table 2.6).

Table 2.6. Freundlich parameters and calculated partition coefficients for tylosin.

a b c,d d Soil Kf N R Kd Koc Log Koc Coloma 10.38 0.81 0.999 11.83 1075 3.03 Drummer 43.68 0.81 0.997 49.67 1242 3.09 Finley 30.85 0.73 0.996 37.16 482 2.68 Burned 54.70 0.70 0.996 67.46 950 2.98 Toolik Peat 75.58 0.65 0.998 96.01 127 2.10 a µM l kg-1 b correlation coefficient of linear regression c -1 determined at Caq= 0.5 µmol l d l kg-1

The Koc values for tylosin appeared to be similar regardless of whether the linear or

Freundlich transformation of the soil sorption isotherms was used. Based upon the linear regressions, the average log Koc=2.88 ±0.20 whereas for the Freundlich transformation the average partition coefficient was slightly higher, with log Koc=2.95 ± 0.18 (where the error represents one standard deviation). The Koc values for Toolik peat were not included in calculations of the average because of the extremely high fraction of organic carbon as compared to the other four soils tested.

While linear sorption of tylosin to soils has been reported (Ter Laak et al., 2009), non-linear sorption of tylosin to soils has more commonly been observed, with N values ranging from 0.83-1.03 (Rabølle et al., 2000, Zhang et al., 2011) to significantly non-

39 linear (0.42-0.80 [Sassman et al., 2007]). While our calculated Kd values are within a

-1 similar range as those from Rabølle and coworkers (2000) (Kd=8.3-128 l kg ), our Koc

-1 values were only comparable with respect to two soils (Koc=553-771 l kg ) with the two

-1 additional soils having Koc values an order of magnitude larger (Koc=5664-7988 l kg ).

Because tylosin desorption was only on the order of 13-14% from these two high Koc soils, I suspect that tylosin undergoes irreversible specific interactions with the soils or transformation (biotic or abiotic). With respect to the latter process, degradation can be mistaken for sorption, resulting in an overestimation of Koc. Zhang et al. (2010)

-1 measured Kd that ranged from 1.89-8.44 l kg . While two of their calculated Koc values

-1 were within the same order of magnitude as our values (Koc=127 and 365 l kg at Caq=5

-1 -1 mg l ), their third soil had a far higher calculated Koc=10825 l kg . Further, the surface area of this soil was 2-113 times greater than that of the other two soils. If tylosin undergoes surface-mediated specific interactions, this soil could have a greater number of available sites, resulting in significantly more sorption. The work by Sassman (2007) which looked specifically at the sorption of tylosin A, used Caq=0.2 µM to calculated Kd

-1 values an order of magnitude larger (Kd=721-5520 L kg ) than the values calculated here

-1 except for one soil (Eustis-25, Kd=2.23 L kg ). Because many of these soils have far lower ƒoc than the ones used here, it is possible that there are more sites available for specific interactions without inhibition due to sorption of DOM to the mineral surface.

The observed non-linearity of tylosin sorption was attributed to ion exchange processes with the mineral phase, based upon an observed correlation between the CEC of the soils and the sorption behavior (Sassman et al., 2007).

Previous researchers have considered the contribution of several different mechanisms to the overall observed non-linearity of tylosin sorption and very large Koc values relative to its octanol water partition coefficient. Sassman (2007) correlated the cation exchange capacity of tylosin to soils, suggesting that electrostatic interactions could contribute to tylosin sorption. Tylosin is a weak base with a pKa of 7.2 (Boxall et al., 2006), meaning that its protonated species will be dominant at the pH of all the soils investigated, except for Drummer. This could allow it to form ionic complexes with

SOM hydroxyl or carboxyl functional groups or negatively charged clay minerals. In addition, both Sassman (2007) and Zhang (2011) suggested that the higher molecular weight of tylosin (916.12 g mol-1 for tylosin A; [Boxall et al., 2006]) could prevent it from accessing all the sorption domains of the soils, particularly at lower concentrations, thus contributing to non-linearity. Although antibiotic sorption to clay interlayers has been shown to increase the spacing between these layers, this increase is not observed with tylosin, suggesting that it is too large to enter these gaps (Kumar et al., 2005).

In addition to the observed non-linearity of tylosin sorption to these soils, the calculated Koc values were far higher than would be predicted based upon the octanol- water partition coefficient of tylosin (log Kow=1.63; [Boxall et al., 2006]).

Underestimation of sorption behavior based upon Kow values has previously been attributed to several possible phenomena, including either concurrence of several different sorption mechanisms (discussed above) or sorption to reactive mineral phases

(Weber et al., 1992). The sorption of tylosin to clay minerals has been hypothesized to occur predominantly via cation exchange processes (Essington et al., 2010). Further, in

41 addition to hydrophobic partitioning, analysis of sorbed macrolide-oxide complexes with

Fourier Transform Infrared Spectroscopy (FT-IR spectroscopy) suggest that surface complexation at the carbonyl group of roxithromycin contributes to sorption (Feitosa-

Felizzola et al., 2009). This carbonyl group, located on the central lactone ring, is conserved in tylosin, suggesting a mechanism for specific interactions between tylosin and metal oxides in soils.

Evidence for this type of mechanism has been observed for roxithromycin and clarithromycin, two macrolide antibiotics with similar structures to tylosin. They have been found to adsorb to the surface of iron and manganese oxides (Feitosa-Felizzola et al., 2009). Although adsorption of macrolides to Fe and Mn oxides in soils was observed over 24 h in the Feitosa-Felizzola (2009) study, these interactions resulted in degradation of the macrolide compounds over a longer time period (up to 16 days). If tylosin were forming similar complexes and undergoing transformation in our soils, we would expect to see a change in the mass balance over time.

In order to investigate this, mass balances were completed for both tylosin and progesterone. Whereas progesterone underwent minimal degradation in the time required to reach quasi-equilibrium (Figure 2.8), the amount of recovered tylosin decreased over time (Figure 2.9). Microbial degradation of organic contaminants can also occur over this time span, however, microbial degradation of tylosin required a 3-week “lag period” after addition of the compound to soils (Sassman et al., 2007). In addition, I autoclaved our soil which should have prevented microbial activity in our samples.

Decreasing tylosin mass balance over time could also be explained by irreversible sorption. However, previous studies have found limited desorption hysteresis, with negligible differences between sorption and desorption isotherms for tylosin in contact with soils over 24 h (Zhang et al., 2011). Zhang and coworkers (2011) suggest hysteresis may occur as tylosin reaches equilibrium with the micropores in the glassy phase; however, this would likely occur over a much longer time frame (i.e., months). I hypothesize that metal-oxide mediated degradation of tylosin in our soils occurs and is similar to that observed by Feitosa-Felizzola (2007) and colleagues for other macrolide compounds. This hypothesis is supported by the presence of both Fe and Mn oxides in our soils, as determined by XRF. Iron and manganese concentrations ranged from 2.09-

7.99% and 0.059-0.175%, which were measured as percent weight of Fe2O3 and MnO2 in the bulk soil, respectively (Table 2.7).

Table 2.7. Percent weight of iron and manganese oxides (measured as Fe2O3 and MnO2) in soils. Fe Mn Finley 5.22 0.071 Drummer 3.71 0.087 Coloma 2.09 0.058 Burned 7.99 0.175 Toolik peat 4.37 0.06

It is possible that this proposed degradation mechanism could contribute to the disparities in reported tylosin Koc values due to variations in metal oxide concentrations, which have not been reported for previous sorption isotherm studies. This is the basis of the study described in Chapter 3. 43

2.4 Conclusions

Based upon the completed soil sorption isotherms, both tylosin and progesterone exhibited strong, non-linear sorption with the six soils used in this study. These sorption isotherms are the first completed for progesterone, and suggest that it exhibits sorption behavior similar to other EDCs. I hypothesize that observed non-linearity could be due to the diffusion of progesterone into the interlayer spacing of clay molecules contained within the mineral fraction of soils based upon the observed behavior of structurally similar compounds. Future studies to determine the sorption of progesterone to isolated clay fractions (with a comparison of sorption to both swelling and single-layer clays) could help provide evidence for this hypothesis. In addition, relating the abundance of

HSACM in our soils to the observed behavior of progesterone at lower concentrations could help provide evidence for or against the specific interaction of progesterone with these soil fractions and should be investigated in the future.

Based upon its lower Kow, tylosin exhibited stronger sorption than anticipated, although our observed Koc values were similar to those previously published. Tylosin exhibited non-linear sorption behavior, which has previously been attributed to specific interactions between charged tylosin species and mineral surfaces or retarded diffusion into micropores due to tylosin’s large size. However, we suggest that tylosin could also be undergoing sorption and degradation in contact with metal oxide species. Further investigations were conducted in order to study these interactions and are described in the following chapter. The high Koc values of both compounds suggest that they will

44 undergo limited transport to surface water systems and instead could accumulate in soil environments.

Figure 2.1. Anticipated transport pathways of veterinary pharmaceuticals in the environment.

0.12

0.1

1 - 0.08

0.06

0.04

µmolsorbed soil g

0.02

0 0 0.5 1 1.5 2 2.5 Time (days)

Figure 2.2. Progesterone soil sorption kinetics with Finley soil and an initial nominal aqueous concentration of 3 µM.

0.03

0.025

0.02

0.015

µmolsorbed soil g 0.01

0.005

0 2 4 6 8

Time (days) Figure 2.3. Tylosin soil sorption kinetics with Drummer soil and an initial nominal aqueous concentration of 3 µM.

0.35 Finley Drummer 0.3

Coloma

1 - 0.25 Burned 0.2 Toolik

0.15

sorbedµmolsoil g 0.1

0.05 0 0 1 2 3 4 5 6 µM aqueous

Figure 2.4. Linear sorption isotherms for progesterone with initial nominal aqueous concentrations from 0.5-5 µM. Vertical error bars represent one standard deviation.

log (µM aqueous) -1 -0.5 0 0.5 1 0

-0.5

) 1 -

-1

-1.5 Finley -2 Drummer

sorbed(µmol log g soil Coloma -2.5 Burned

-3 Toolik

Figure 2.5. Freundlich sorption isotherms for progesterone with nominal initial aqueous concentrations from 0.5 -5 µM. Vertical error bars represent one standard deviation.

0.02

0.016

1 -

0.012

Finley sorbed g soil 0.008

ol Drummer µm Burned 0.004

0 0 0.1 0.2 0.3 0.4 0.5 µM aqueous Figure 2.6. Linear sorption isotherms for progesterone with nominal initial aqueous concentrations from 0.05-0.5 µM. No error bars are depicted because these are the results of a single experiment.

0.16

0.14 0.12

1 - 0.1

0.08 Finley

0.06 Drummer Burned 0.04

µmolsorbed g soil 0.02

0 0 1 2 3 4 5 µM aqueous

Figure 2.7. Linear sorption isotherms for progesterone with nominal initial aqueous concentrations of 0.05-5.0 µM. No error bars are included because the lower concentrations were only determined once.

0.35 Finley Drummer 0.3

Coloma

1 - 0.25 Burned

Toolik 0.2

0.15

sorbedµmolsoil g 0.1

0.05 0 0 1 2 3 4 5 6 µM aqueous

Figure 2.8. Linear soil sorption isotherms for tylosin with nominal initial aqueous concentrations from 0.5-5.0 µM. Vertical error bars represent one standard deviation.

log (µM aqueous) -2 -1.5 -1 -0.5 0 0.5 1 0 Finley

-0.5 Drummer

) 1 - Coloma -1 Burned Toolik -1.5

-2

(µM sorbedsoil log g -2.5

-3 Figure 2.9. Freundlich sorption isotherms for tylosin with nominal initial aqueous concentrations from 0.5 – 5 µM. Vertical error bars represent one standard deviation.

110 100 90 80 70 60 50

% remaining% 40 30

10 0 0 0.5 1 1.5 2 2.5 Time (days) Figure 2.10. Representative mass balances for progesterone based upon triplicate extraction of the soil with methanol. No error bars are included because these results are from a single trial using Finley soil.

100 90 80

60 50

40 % remaining% 30

20 10 0 0 1 2 3 4 5 6 7 8 Time (days) Figure 2.11. Representative mass balances for tylosin based upon triplicate extraction of the soil with methanol. No error bars are included because these results are from a single trial using Drummer soil.

Chapter 3: Tylosin transformation by soil mineral oxides.

3.1 Introduction

As discussed in Chapter 2, the sorption of tylosin to soils has been previously studied. However, little research has examined the interaction of tylosin with the soil mineral phase, despite the fact that it often accounts for over 90% of the total solids in a bulk soil (Sparks, 2003). The inorganic portion of soils consists of both primary minerals, which are unaltered since their formation, and their weathering products, which are known as secondary minerals. Because weathering is ubiquitous, secondary minerals are prevalent in almost every soil system. Phyllosilicates (also known as clay minerals) are secondary minerals which are found mainly in the clay (< 2 mm) fraction of soils and consist of Si-O tetrahedral structures assembled in sheets of either tetrahedral or octahedral thickness. Due to the substitution of Al3+ for Si4+, phyllosilicates often have a net negative charge, with pHpzc ranging from less than pH~4.0 to up to pH~6.5 (Sakurai et al., 1988). This negative charge is often balanced by the interaction of cations with hydroxyl and oxygen groups on the outer edge of the clay (Sparks, 2003). Cations can also interact with surface hydroxyl groups in the interlayer spaces of clays where two tetrahedral sheets coordinate with a single octrahedral sheet in the middle. Antibiotics

57 have previously been found to adsorb within the interlayer spacing of clays, increasing the distance between the sheets (Nowara et al., 1997; Chang et al., 2009).

Although they are less abundant than phyllosilicates, metal oxides also play an important role in soil chemistry. Metal oxides refer to either amorphous or crystalline Al,

Fe, and Mn hydroxides, oxyhydroxides, and hydrous oxides. Goethite (ɑ- FeOOH) is one of the most thermodynamically stable iron oxides, and so is common in soils of all climates. Goethite has a central Fe ion surrounded by three O2- and three OH- groups to give FeO3(OH)3 octahedra, which link to form chains (Cornell et al., 2003). Both synthetic and naturally occurring goethite exhibit acicular (or needle-like) crystal shapes, with specific surface areas usually ranging from 8-200 m2 g-1 (Cornell et al., 2003).

Manganese oxides are often found in soils in their amorphous form, coating surface particles and fissures, or as large nodules (Sparks, 2003). Birnessite (δ-MnO2), the most common Mn oxide, has a structure similar to phyllosilicate clays, with a combination of Mn3+ and Mn4+ at the center of linked octahedra, with an interlayer between which contains exchangeable cations and water molecules (Dixon et al., 2002).

Because of the presence of Mn3+ and frequent substitution, birnessite has a negative surface charge at environmentally relevant pH values. For birnessite, the pHpzc~1.5 whereas for goethite it is much higher, at pHpzc~8-9 (McKenzie, 1989; Cornell et al.,

2003).

Because both Mn and Fe can act as Lewis acids, exposed atoms at mineral surface sites in aqueous conditions will coordinate with hydroxyl or water molecules (which then dissociate). In addition to structural hydroxyl groups, this will result in mineral surfaces

58 covered in hydroxyls that act as ligands (Cornell et al., 2003). Water molecules can react with the surface hydroxyl groups via hydrogen bonds to form a “hydration shell” around the metal oxide surface. The water molecules in this layer are more ordered than the surrounding aqueous solution and thus have different properties, such as a lower dielectric constant (Cornell et al., 2003). Cations are frequently drawn to the surface of mineral oxides due to the presence of deprotonated hydroxyl groups. However, due to both their organized shell of water molecules and the hydration shell on the mineral surface, they are frequently limited in their approach to the mineral surface by a distance of 3-6 nm (the width of 1-2 water molecules, [Johnston et al., 2002]). These non-specific interactions are known as outer sphere complexes.

Ions which have a high affinity for specific sites and can form coordination bonds or have electrostatic interactions with surface functional groups can form inner sphere complexes. Inner sphere complexes occur when the aqueous species binds directly to surface functional groups without steric inhibition from water molecules. Because these complexes form stronger bonds than outer sphere complexes, they are more likely to be irreversible or undergo slower kinetics (Johnston et al., 2002). Inner sphere complexes can be monodentate, where a single bond forms between the mineral and the adsorbate, or bidentate, where two bonds form between a metal atom at the mineral surface and the adsorbate. In a variation of monodentate complexation, differing numbers of bonds can form between functional groups of the adsorbate and the mineral surface, resulting in bi- or tri-nuclear complexes (Deng et al., 2002).

The formation of inner sphere complexes can sometimes result in the dissolution or reduction of mineral surfaces. Dissolution can occur after adsorption of a species to a metal atom, polarizing and weakening the metal-O bond in the mineral structure and making it easier to cleave. Dissolution most commonly occurs after the formation of mononuclear (often bidentate) inner sphere complexes, which cause the greatest depolarization of the metal-O bond (Cornell et al., 2003). The reduction of metal species within the mineral structure also results in a weakening of the structural metal-O bonds.

Metal reduction is also facilitated by formation of inner sphere complexes, which allows electron transfer between the metal atom and the adsorbate, provided that the adsorbate is capable of donating electrons and the reduction potential is favorable (Cornell et al.,

2003). Reduction is frequently followed by dissolution, resulting in an increase in the reduced metal species in the aqueous soil environment. This reduction and subsequent dissolution is often facilitated by adsorption of organic contaminants, with a correlation between the number of phenolic or carboxylic functional groups and the rate of metal reduction (Chen et al., 2003). An additional pathway for dissolution of metal oxides involves biotic reduction, in which the oxidation of organic compounds is coupled with metal oxide reduction in order to provide energy for soil microbial species (Chen et al.,

2003).

Many different classes of organic contaminants, including VPs, will adsorb to the surface of soil minerals through inner sphere complexes. Tetracycline (and similar compounds of the same antibiotic class which are commonly used as GPs) will readily sorb to both clay and mineral oxide surfaces. Sorption is likely due to specific

60 interactions between tetracycline functional groups and surface charge sites on the mineral fraction (Gu et al., 2005(1); Pils et al., 2007). The adsorption of tetracyclines to mineral oxide surfaces is the initial step in surface-mediated transformation reactions.

For example, tetracyclines have been found to undergo dehydration and epimerization

3+ reactions after sorption to Al oxides (Chen et al., 2010). In addition, MnO2 underwent reductive dissolution due to the transfer of electrons during surface-mediated oxidation and transformation of tetracyclines (Rubert et al., 2006).

The interactions of fluoroquinolones (FQs), another common class of antibiotics have also been studied with both clays and mineral oxides. In addition to sorption within the interlayer of swelling clays, specific adsorption of several FQs appears to be dependent on a keto acid structure; without the keto acid structure, sorption to clays drops by two orders of magnitude (Nawara et al, 2007). This ketone functional group is also important for the sorption of ciprofloxacin (a FQ) with hydrous iron oxide, due to its ability along with a carboxylic functional group to form a bidentate inner sphere complex with Fe (Gu et al., 2005(2)). FQs have also been found to undergo transformation in the presence of MnO2. This transformation is mediated by the formation of a surface complex between the FQ and the MnO2 surface (Zhang et al., 2005). After complexation, an aromatic nitrogen within a piperazine ring donates an electron to the MnO2, resulting in the formation of a radical FQ compounds which can undergo further oxidation reactions to form transformation products (Zhang et al., 2005). Several other classes of antibiotic GPs have also been found to undergo oxidative transformation after sorption to

MnO2 (Zhang et al., 2008). These transformations typically follow complicated kinetics,

61 with a pseudo-first order initial reaction rate followed by more complicated rates which did not fit models at later stages (Zhang et al., 2008).

The addition of dissolved organic matter (DOM) has been found to have an effect on the formation of inner sphere complexes between GPs and soil mineral oxides. DOM has been found to have varying effects, ranging from a slight (5-20%) increase in transformation rates to a 10-15 fold decrease in reaction rates (Kang et al., 2004; Chen et al., 2010(2); Feitosa-Felizzola et al., 2009). DOM will readily adsorb to the surfaces of soil minerals, with preferential sorption to the specific active sites that are present (Kaiser et al., 2003). The hydrophobicity of DOM has been found to have an effect on its sorption to mineral surfaces (Kaiser et al., 1997). If DOM is more hydrophobic than the organic contaminant it could outcompete the compound for preferential sorption to the mineral surface. This would block active sites from the contaminant and prevent its sorption, resulting in a decrease in the rate of transformation of organic contaminants. In addition, DOM could also have an indirect effect on reaction rates by reductively dissolving the metal oxide. The reduced aqueous metal species can then compete for reactive sites and sorb to the surface of the oxide, limiting the number of available active sites (Klausen et al., 1997).

Conversely, the addition of DOM has been found to increase the rate of transformation, largely by serving as a mediator for transformation reactions. Smolen et al. (2003) found that organic acids could reduce Fe3- in goethite, whereby the reduced ferrous iron would then adsorb to the surface of the goethite and facilitate the transformation of a nitrobenzene contaminant. The addition of humic acid substituents

(phenol containing compounds) was found to increase the transformation rate of organic contaminants in the presence of birnessite (Kang et al., 2004; Park et al., 1999). This is likely due to the fact that birnessite can oxidize either the contaminant or the phenolic compound, which can then couple to itself or the other fraction. The addition of bulk humic acid (HA) at low concentrations was also found to increase transformation rates of a pesticide, likely due to birnessite-mediated complex formation between the pesticide and the HA. However, at higher HA concentrations, inhibition of the transformation was observed (Kang et al., 2004), possibly due to adsorption of the humic acid to the active sites of the birnessite, as was previously described (Kaiser et al., 2003).

Little work has been completed regarding the interaction of macrolide antibiotics with inorganic soil components. The adsorption of tylosin to two common clays has previously been described (Essington et al., 2010). Tylosin was found to adsorb readily to montmorillonite at pH values below that of its pKa (pH=7.2); however, only 5% of added tylosin sorbed to the surface of kaolinite. Essington and coworkers (2010) proposed that tylosin sorption to clay surfaces is due to specific interactions, such as ion exchange processes. The adsorption of erythromycin A, a macrolide antibiotic with a similar structure to tylosin, was measured in contact with soils which had been saturated with different ions (i.e., Fe3+, H+, and others) (Kim et al., 2004). This adsorption was followed by degradation, with different degradation pathways for the Fe3+ and H+ saturated soils. The Fe3+ saturated soil promoted loss of a neutral group from the central lactone ring, whereas the H+ saturated soil promoted both the loss of the neutral sugar group and the formation of an internal oxygen-containing ring structure (Kim et al.,

2004). Other macrolides, clarithromycin and roxithromycin, were found to adsorb to ferrihydrite and amorphous MnO2 via surface complexation at the carbonyl group located on the central lactone ring at the carbon 1 position (Feitosa-Felizzola et al., 2009). In addition to adsorption, both compounds underwent transformation in contact with both mineral oxides, with the major transformation pathway likely being hydrolysis of an ether linkage to a sugar functional group. Feitosa-Felizzola and coworkers (2009) suggested that this transformation pathway was facilitated by surface interactions with the metal oxides.

I hypothesize that tylosin would likely undergo metal oxide-surface mediated transformations, similar to those of clarithromycin and roxithromycin. If true, this would provide an explanation for the decreasing mass balances found for tylosin in contact with soils described in Chapter 2. In order to investigate if such transformations occur, tylosin was added to suspensions of goethite (ɑ-FeOOH) and birnessite (δ-MnO2) and its aqueous concentration was measured over time. An additional experiment was conducted in order to determine the effect of dissolved organic matter (DOM) on the rate of transformation of tylosin in contact with mineral oxides. This study is important for understanding the fate, including residence time, of tylosin in soil systems.

3.2 Methods

Mineral Synthesis

Goethite was previously synthesized based upon the method described by Wang et al. (1997). The synthesis and identification of goethite was previously described

(Meier, 1999). After synthesis, the goethite was stored at room temperature in the dark until use. Birnessite was synthesized based upon the reduction method described by Luo et al. (2000). Briefly, a solution of 15 g of KMnO4 (certified ACS grade, Fisher

Scientific, Fair Lawn, NJ) and 30 g of KOH (ACS grade, BDH, TOWN) in 300 mL of

MQ was stirred for five minutes and then added, with stirring, to a solution of 30 g of

KOH in 100 mL of neat ethanol (>99.5+%, Acros Organics, Morris Plains, NJ) and 200 mL of MQ. This mixture was stirred for 3 minutes before being poured into 125-mL plastic bottles. The bottles were then heated at 85o C for 24 h. The manganese oxide product plus any remaining supernatant was washed with 500 mL of MQ using a gravity vacuum filtration set up and an 8-µM qualitative cellulose filter (Whatman, United

Kingdom). The product was then transferred from the filter paper to a glass Petri dish and dried at 60o C for 24 h. The mineral which formed was ground using a ceramic mortar and pestle before identification and use in experiments.

Identification was completed via powder X-ray diffraction using a PANalytical

X’Pert Pro X-ray diffractometer in the Subsurface Energy Materials Characterization

Laboratory (SEMCAL), School of Earth Sciences, Ohio State University. Data were collected from 5 to 70 degrees 2-theta using Cu Kɑ radiation, a tube current of 40 mA, and a tube voltage of 45 kV. The resulting powder scan was analyzed using the search match algorithm in X’Pert High Score Plus and the crystallography open database

(COD). The experimental scan (Figure 3.1) best matches that of a calculated birnessite pattern (Downs et. al, 1993). Original data were obtained for a K-bearing birnessite (Post et al., 1990). Diffraction lines matching the reference material include the following d-

65 spacings: 7.06, 3.53, 2.55, 2.48, 2.39, 1.77, 1.5, 1.47, and 1.42 Å. Some expected peaks

(2.25 and 2.12 Å) in the reference pattern were missing in the experimental pattern. In addition, the presence of additional peaks not attributed to birnessite suggests possible contamination and it is likely that some amorphous manganese oxide or other trace phases are present in the sample.

Transformation Kinetics

Bulk degradation experiments were completed in order to investigate the interaction of tylosin with mineral oxides. Two mM Na2CO3 (certified ACS grade, Fisher

Scientific) was used as a buffer and it was adjusted to pH=6.5±0.1 before use with solutions of either HCl (ACS grade, BDH) or NaOH (certified ACS, Fisher Scientific).

Forty mg of metal oxide was added to a 125-mL glass Erlenmeyer flask, while 40 mL of buffer was added to both this flask and an additional empty flask, which served as an experimental control. The flask containing the metal oxide was then sonicated for 10 minutes to ensure suspension of the oxide. After sonication, 2 mM tylosin stock solution in methanol (described in Chapter 2) was added to both flasks to a concentration of 1 µM.

Both flasks were stirred using Teflon coated stir bars throughout the duration of the experiment and were kept on cork rings to prevent incidental heating from the stir plates.

Rubber corks wrapped in Parafilm were used to stopper both flasks, which were wrapped in aluminum foil to prevent light-mediated degradation.

At time points ranging from 1 h – 7 d, duplicate 500-µL aliquots were removed from each flask and added to 500 µL of methanol (HPLC grade, EMD, TOWN) in a 1.5-

66 ml micro-centrifuge tube. Methanol was added because it has been shown to preferentially adsorb to oxide surfaces, thereby quenching the reaction (Spitz et al, 1986).

Each sample was then centrifuged for 30 minutes at 30,000 rcf (Beckman Microfuge E,

Beckman Coulter Inc., Brea, CA). Between 5- 12% of the analyte was lost during processing due to adsorption of tylosin to the walls of the plastic microfuge tubes.

Immediately after centrifuging, the samples were transferred to 1-ml amber glass vials and stored at 4o C until analysis via reverse phase high-performance liquid chromatography, RP-HPLC.

A similar experiment was completed in order to investigate the effect of DOM on tylosin degradation rates. For this experiment, Suwannee River natural organic matter

(SRNOM) was used because it is considered representative of organic material originating from terrestrial sources (Chin et al., 2004). The collection and isolation of the

SRNOM used in this experiment was previously completed and is described elsewhere

(Mash, 2001). In order to determine the effect of DOM on tylosin transformation rates,

-1 SRNOM was added to 2 mM Na2CO3 buffer to a concentration of 20 mg l . After addition of the SRNOM, the buffer was adjusted to pH=6.5 ±0.1 and the experiment was conducted as described above.

Minimal microbial degradation of tylosin in soils has been reported over time periods ranging from 48 h to 20 d (Allaire et al, 2006; Sassman et al, 2001) and we would expect few microbes to be present in this experimental design to a lack of organic substrates. However, a sterile experiment using methods similar to those described above was completed in order to ensure that biological activity did not contribute to observed

67 tylosin degradation. The experimental design was the same except that the automatic pipette tips used to transfer aliquots and the microfuge tubes were autoclaved for 15 min at 121oC (Tuttnauer Brinkmann 2340M, Tuttnauer USA Co. Ltd., Hauppauge, NY). The buffer pH was adjusted to pH=6.5 ± 0.1 before being autoclaved for 1 h under the same conditions. The metal oxide was added to the Erlenmeyer flask and both flasks were autoclaved for 30 min under the same conditions before addition of the buffer and tylosin and completion of the experiments.

Analysis of the aqueous tylosin concentration at all time points was completed using RP-HPLC (either a Waters 1515 or 1525 pump and 717plus Autosampler, Waters

Corp., Milford, MA), with UV-Vis detection (Waters 2487 dual λ absorbance detector) at

287 nm. The mobile phase was 50:50 (v/v) 5 mM phosphoric acid (ACS reagent grade,

Acros Organics, Morris Plains, New Jersey) and acetonitrile (HPLC grade, J.T. Baker).

The flow rate was 1.0 ml min-1 and injection volume was set at 100 µl. A Whatman reverse phase guard column was used to protect the main column, a Restek Pinnacle DB reverse phase analytical column (C18 stationary phase, length, 150 mm; inside diameter,

4.6 mm; particle size 5.0 µm).

Analysis of transformation products was conducted at the end of the experiment

(7 days). An aliquot of each sample was diluted with methanol to a final 1:1 ratio in a

25-mL Corex centrifuge tube. The samples were then centrifuged for 1 h at 1600 rpm and 23oC (Beckman GS-6R, Beckman Coulter Inc., Brea, CA). The supernatant was filtered through a 0.6-µm pore size glass fiber filter (Sterlitech, Kent, WA) before analysis via liquid chromatography-triple quadrupole mass spectrometry (LC-MS). For

LC-MS, a Thermo Scientific system was used, which included an Accela binary pump and autosampler. The same Pinnacle DB C18 column described above was used with a flow rate of 0.4 ml min-1; however, the injection size was changed to 20 µl. Gradient elution was used with two mobile phase: phase A, 0.1 % formic acid (~98% purity,

Fluka, St. Louis, MO) in acetonitrile (v/v) and phase B, 0.1% formic acid in Milli-Q water. The mobile phase composition was held at 100% A for 2 min, before decreasing to 80%/20% A/B from 2-10 minutes. After HPLC separation, the analytes were analyzed using a Finnegan TSQ Quantum Discovery Max (Thermo Fisher Scientific, Waltham,

MA) triple quadrupole mass spectrometer run in MS mode. Analysis was completed using positive ion mode with electrospray ionization (ESI). The mass spectrometer was operated under the following conditions: capillary temperature, 300 oC; sheath gas pressure, 2 psi; auxiliary gas pressure, 2 psi; collision gas, argon; spray voltage, 5 kV.

The MS was run in both total ion and scanning mode, with scanning completed for 916,

772, 406, 192, 174, and 144 m/z fragments.

3.3 Results and discussion

Transformation Kinetics

In the absence of metal oxides, tylosin appeared to be relatively stable over 7 d, with a loss of approximately 20% of the initial concentration (Figure 3.2). This is similar to the reported aqueous stability of tylosin, with estimated losses of 0-20% over a similar time frame at pH=6.0-7.0 (Paesen et al., 1995; Loftin et al., 2008). Losses are likely due to sorption to the walls of the plastic centrifuge tubes and the glass sample vessel walls.

In the presence of goethite, tylosin underwent degradation and obeyed pseudo-first order kinetics (Figure 3.2). An exponential regression of the transformation curve resulted in a pseudo-first order rate constant, k, of 0.026 h-1 and a calculated half-life of 27 h

(R2=0.991). The change in aqueous concentration of tylosin in contact with birnessite was also measured. Although complex kinetics have previously been observed for organic contaminants in contact with birnessite (Zhang et al., 2008), the transformation of tylosin in contact with birnessite fits a pseudo-first order model (Figure 3.3), with an observed pseudo-first order rate constant k=0.005 h-1 and an estimated half life of 139 h

(R2=0.991).

Experiments conducted under sterile conditions had little effect on observed tylosin behavior with goethite (Figure 3.4). Tylosin still underwent pseudo-first order degradation, with k=0.028 h-1 and an estimated half life of 25 h (R2=0.989). This suggests that microbial degradation of tylosin in our system is minimal, and instead we are observing abiotic reactions. Biotic degradation is likely minimal due to a lack of substrates for microbes within the experimental design and the inhibition of microbial growth due to the antibiotic properties of tylosin. Under sterile conditions, the rate of tylosin degradation in the presence of birnessite increased (Figure 3.5). Tylosin degradation still obeyed pseudo-first order kinetics of transformation, with a pseudo-first order rate constant k=0.012 h-1 and a calculated half-life of 58 h (R2=0.996). It is unclear why autoclaving would result in an increase in transformation rates, but since only one replicate was tested, it is possible that these results are due to experimental error.

Feitosa-Felizzola and coworkers (2009) also reported first-order kinetics for the transformation of similar macrolide compounds with metal oxides. They observed similar rate constants (k=0.014 h-1 and k=0.010 h-1) for the interaction of ferrihydrite with clarithromycin and roxithromycin, respectively. While within the same order of magnitude, these k values result in the calculation of slightly longer half-lives (49 and 71 h, respectively) than was calculated for the interaction of tylosin with goethite. The

-1 reported half-lives for both compounds in contact with amorphous MnO2 (43 h and 42 h-1, respectively) were almost three times faster than was found for tylosin in contact with birnessite. The observed differences in degradation rates are likely due to differing transformation pathways (as discussed below).

A preliminary experiment was also conducted to investigate the effect of DOM on the rate of degradation of tylosin in contact with both mineral oxides (Figure 3.6).

Although tylosin transformation still exhibited first order kinetics with both minerals, the rates of transformation changed. For goethite with SRNOM, tylosin degradation had a rate constant of k=0.018 h-1 and an estimated half life of 39 h (R2=0.977), which is ~1.5 times slower than degradation in contact with goethite in the absence of DOM. Feitosa-

Felizzola and coworkers (2009) also reported a decrease in reaction rates for the transformation of clarithromycin and roxithromycin with oxides after the addition of

DOM, although they found that the addition of DOM caused the rate to slow down by a factor of 10-15. The greater effect of DOM on the kinetic data of Feitosa-Felizzola et al.

(2009) could be due to their use of commercially purchased humic acid as opposed to the isolated NOM used here. As discussed in Chapter 2, commercially purchased organic

71 matter is a poor substitute for NOM due to its differing chemical and physical properties

(Malcolm et al., 1986; Chiou et al., 1987). The observed inhibition is likely due to preferential sorption of the DOM to the clay surfaces, masking specific interaction sites from the antibiotics (Pils et al., 2007). In addition, DOM can serve as a Lewis base and contribute to the reductive dissolution of metal oxides. The reduced metal species can then adsorb to the mineral surface and reduce the number of available reactive sites, thus repressing reaction rates (Klausen et al., 1997). However, the aqueous metal concentration was not measured, making it difficult to determine the contribution of this mechanism to the observed results.

The addition of DOM was found to double the rate of tylosin degradation in contact with birnessite; an exponential regression resulted in a pseudo-first order rate constant of k=0.01 h-1 and a half life of 69 h (R2=0.976). These results suggest that the addition of DOM facilitates the degradation of tylosin in contact with birnessite.

Although many studies have observed an inhibition of transformation due to DOM, an increase in transformation rates due to the presence of DOM for phenol-containing compounds has also been reported (Park et al., 1999; Kang et al., 2004). It is possible that the DOM could be serving as an intermediate, thus facilitating transformation. In the presence of birnessite, phenols can be oxidized to quinone or radical groups (McBride,

1987). These products could then react with tylosin, resulting in increased transformation rates.

The sorption of DOM to mineral surfaces can be correlated to surface charge of the adsorbent with positively charged minerals will likely undergoing more sorption due

72 to electrostatic interactions (Meier et al., 1999). Due to the negative surface charge of birnessite under these experimental conditions (pHpzc~1.5, [McKenzie, 1989]), the sorption of SRNOM will likely be reduced due to repulsive forces. Greater sorption will likely be observed with goethite (pHpzc~8-9, [Cornell et al., 2003]), which could explain why repressed transformation kinetics were seen with goethite as opposed to birnessite.

However, these results are based upon a single trial and are only preliminary; thus, further replication is required in order to ensure the validity of these transformation rates.

Transformation Pathways

Feitosa-Felizzola and coworkers (2009) reported similar pathways for clarithromycin and roxithromycin in contact with both Fe and Mn oxides. They found that approximately 80% of the observed degradation occurred via hydrolysis of the ether linkage between the central lactone ring and the cladinose sugar group. Minor transformation pathways also included dealkylation of the amine group attached to the desosamine sugar group and hydrolytic cleavage of the central lactone between the ether and the carbon in the 13 position.

Although they reported similar pathways for both oxides, there appear to be differing transformation pathways for tylosin with goethite and birnessite. Within the

RP-HPLC chromatograms, a decrease in the tylosin A peak (~3.9 min elution time) is observed after contact with goethite (Figure 3.7). This suggests that tylosin undergoes a transformation which alters the conjugated bonds between the carbon at the 13 position and the ketone group at C9, thus reducing absorbance at 287 nm. This reduction in

73 absorbance could be due to cleavage of this chromophoric region, preventing absorbance at any wavelength. Another possibility is that addition to the ketone group or one of the double bonds could result in a hypsochromic shift, or a shift in absorbance to shorter wavelengths, due to a reduction in the number of electrons available for delocalization.

Absorbance was also measured at 221 nm and the absence of peaks at this wavelength suggests that tylosin likely undergoes cleavage in contact with goethite instead of addition reactions. Because transformation occurs within the chromophoric region, it is likely that tylosin binds to Fe3+ at the goethite surface via the ketone group at C9. The formation of complexes between metal oxide surfaces and oxygen-containing functional groups is common (Gu et al., 2005; Feitosa-Felizzola et al., 2009).

It is possible that the sugar groups are cleaved after the transformation of tylosin at the chromophoric region. The Fe-mediated cleavage of ether linkages has previously been observed in the transformation of erythromycin with Fe3+ saturated clays (Kim et al., 2004). We would not be able to detect these transformations by UV-Vis spectrophotometry; however, LC-MS analysis of the 7-d samples provided evidence for tylosin degradation via these pathways. Because the MS fragmentation pattern of tylosin results in the same molecules that we anticipate would form from mineral oxide-mediated transformation, we cannot use the presence of a specific fragment ion as evidence for degradation processes. However, if cleavage were occurring solely due to fragmentation in the MS, we would expect a constant ratio of the fragment ions to the parent peak at the same elution time for all samples even though the actual concentration of tylosin changes.

This also assumes that the LC-MS parameters are identical between the control and the

74 sample i.e., they are run on the same day under identical conditions. For tylosin in contact with goethite for 7 days, we found that the ratio of fragment ion to parent peak changed relative to the control sample which did not have goethite present (Table 3.1).

Table 3.1. The ratio of the abundance of fragment ions to the tylosin parent peak after reacting with goethite for 7 days.

m/z (fragment): 916.16 Control Goethite 916.16 1:1 1:1 (tylosin) 772 1.14:1 1.37:1 (tylosin-mycarose) 406 4.17:1 7.33:1 (lactone – sugars) 144, 174, 192 87.14:1 299.6:1 (sugar groups)

In addition, we found a proportional ratio between the fragments which were formed and the parent ion (i.e., an approximate triplicate increase in the signal for all three sugars in relation to the parent peak). This suggests the formation of the sugars (which are also the main fragments in the MS) in solution before the sample came in contact with the ion source of the MS.

For tylosin in contact with birnessite, a decrease in the tylosin A peak was also observed. However, there is a concomitant appearance of two additional peaks at earlier retention times (labeled ‘Product Peaks’ in Figure 3.8). Because the product peaks still absorb at 287 nm, it suggests that the chromophoric region in the degradates remain intact and tylosin is instead undergoing a different transformation pathway than that mediated by goethite. I hypothesize that product peak 2 in Figure 3.8 is likely tylosin B 75

(see Figure 1.1 for its structure). As discussed in Chapter 1, tylosin is usually present as a mixture of four different structures, with tylosin A and tylosin B being the first and second most abundant, respectively. The peak at ~3.3 min is present in standard solutions of tylosin, which we hypothesize is tylosin B based upon its lower hydrophobicity (and thus, earlier elution) as compared to tylosin A. Tylosin B has been observed to elute earlier than tylosin A in other studies using HPLC (Hu et al., 2007).

Because this product peak elutes at the same time as in the standard solutions, I suspect it is tylosin B; however, the use of a standard or nuclear magnetic resonance (NMR) of the isolated product would be necessary to conclusively identify its structure.

Tylosin B is formed from tylosin A via cleavage of an ether bond, releasing the mycarose group. Birnessite-mediated cleavage of ether bonds has been previously reported (Ahn et al., 2006; Chen et al., 2010). Feitosa-Felizzola et al. (2009) proposed that hydrolysis of the ether linkage between the sugar and the central lactone was likely the main transformation pathway for clarithromycin and roxithromycin in contact with

MnO2. Both Feitosa-Felizzola (2009) and Chen et al. (2010) proposed the formation of a six-membered chelate ring structure between the Mn oxide surface and the compound as a precursor to transformation. Thus, it is likely tylosin forms a similar precursor prior to transformation. I hypothesize that chelation occurs between the amine group on the mycaminose ring and the oxygen linking the two sugar functional groups (Figure 3.9).

Subsequent protonation of the mycarose ring results in cleavage of the ether linkage and formation of desmycosin. I anticipate that I should see accelerated kinetics at lower pH if this is the viable pathway.

Product peak 1 is likely tylosin B after cleavage of the ether linkage between the mycaminose ring and the central lactone structure (Figure 3.10). It is also possible that a six member ring could form due to chelation of the manganese oxide surface at the ether linkage (at C5) and the hydroxyl group on the central lactone ring at position 3 could form (Figure 3.10). This could catalyze cleavage of the ether bond, resulting in the central lactone without the mycaminose group. This is similar to the mechanism proposed by Feitosa-Felizzola and coworkers (2009), who suggested the formation of a ring structure involving the metal oxide surface, the oxygen, which forms the ether bond, and a carbonyl group on the lactone. The cleavage of the ether bond could then occur via protonation, as suggested for the formation of tylosin B (above). The formation of this product is supported by the fact that the mycaminose fragment ion (m/z=173) was detected via LC-MS before elution of the tylosin parent compound, suggesting that it was formed in solution before entering the ion source. The rate of transformation of tylosin in contact with birnessite could be slower than with goethite because of increased steric hindrance during the formation of the six membered ring as opposed to specific interactions at the ketone group with goethite.

3.4 Conclusions

Tylosin was transformed in the presence of goethite and birnessite. The transformation in contact with both minerals followed pseudo-first order kinetics, with faster degradation in the presence of goethite as opposed to birnessite (half lives of 27 h and 139 h, respectively). The difference in transformation rates is likely due to different

77 mechanisms. I hypothesize that the birnessite-mediated transformation of tylosin results in the formation of tylosin B (desmycosin) and tylosin B without the mycaminose group.

The formation of both of these products could be mediated by the development of six- member chelate complexes between the MnO2 surface and tylosin, with cleavage of the

C-O bonds occurring via protonation. If protonation is essential to this mechanism, the reaction could be suppressed under basic conditions and accelerated under acidic conditions. Future experiments should be conducted to investigate the effect of pH on the rate of transformation of tylosin in contact with birnessite.

For tylosin in contact with goethite, it is possible that cleavage occurs within the chromophoric region, based upon decreased absorbance at 287 nm without the development of additional peaks at 221 nm. It is possible that this cleavage is mediated via the complexation of the ketone group (at C9) with the goethite surface or possibly oxidation followed by reduction of goethite. However, additional analysis via LC-MS is necessary in order to identify possible transformation products and determine further details regarding their formation pathways.

Similar experiments were completed under sterile conditions in order to ensure that transformation was due to abiotic processes. The rate of transformation for tylosin in contact with goethite decreased slightly, while it increased by a factor of three for tylosin with birnessite. Repetition of this experiment is necessary to ensure that the increased transformation rate is not due to experimental error. Additional experiments were conducted in order to observe the effect of DOM on tylosin transformation rates. With the addition of DOM, tylosin underwent slower transformation in the presence of goethite

78 and faster transformation in the presence of birnessite. Again, further work is required in order to determine if this increase in transformation rate with birnessite is due to experimental error or because the phenolic groups within the DOM can serve as mediators for the transformation of tylosin. In addition, the negative surface charge of birnessite could prevent the sorption of DOM, preventing it from blocking active sites and slowing the transformation of tylosin. Experiments should be conducted in order to determine the rate of transformation of tylosin in the presence of birnessite and phenol- containing model compounds, which could be representative of DOM and its functional groups. Faster degradation rates with the addition of phenolic compounds could help support the hypothesis that they serve as a mediator.

The transformation of tylosin in contact with metal oxides has not previously been reported. Because metal oxides are an important constituent of the soil inorganic fraction, this research provides new insight into the behavior of tylosin in soils. For example, the observed degradation of tylosin provides an explanation for the decreasing mass balances observed for tylosin in contact with bulk soils (as described in Chapter 2).

Furthermore, differences in reported partition coefficients for tylosin could be due to the previously unaccounted for transformation of tylosin with metal oxides in the soils, which could be mistaken as sorption. In addition, the rate of transformation of tylosin in soils will affect its availability for transport, and thus its effect on both soil and surface water ecosystems. Thus, it is important to develop an understanding of the transformation processes of tylosin in contact with the soil minerals.

1.2

µ 0.8

0.6 Control 0.4 Goethite

( Concentration 0.2

0 0 50 100 150 200 Time (h)

Figure 3.2. Kinetics of transformation of tylosin by goethite (ɑ-FeOOH) at pH=6.5 and an initial nominal tylosin concentration of ~1 µM. Error bars represent one standard deviation.

1.2 Control

Birnessite 1

0.8

0.6

0.4

Concnetration (µM) Concnetration 0.2

0 0 50 100 150 200 Time (h)

Figure 3.3. Kinetics of transformation of tylosin by birnessite at pH=6.5 and an initial nominal tylosin concentration of ~1 µM. Error bars represent one standard deviation

1.2

µ 0.8 Control 0.6 Goethite 0.4 Control Sterile

Concentration ( Concentration Goethite Sterile 0.2

0 0 50 100 150 200 Time (h)

1.2

0.8

Control 0.6 Birnessite 0.4 Control (sterile)

(µM)Concentration Birnessite (sterile) 0.2

0 0 50 100 150 200 Time (h)

1.2 Control

1 Goethite

0.8 Birnessite

0.6

0.4

(µM)Concentration 0.2

0 0 50 100 150 200 Time (h)

0.0035 TYL A 1 h 0.0030 7 d

0.0025

0.0020

0.0015

(AU) Absorbance 0.0010

0.0005

0.0000 0 1 2 3 4 5 6 Time (min) Figure 3.7. RP-HPLC chromatogram for tylosin in contact with goethite for 1 h (blue) and 7 d (red), with absorbance at 287 nm.

0.0035 TYL A 1 h 0.003 7 d

0.0025 Product Peaks 0.002

0.0015 2

(AU) Absorbance 0.001

1 0.0005

0 0 1 2 3 4 5 6 Time (min) Figure 3.8. RP-HPLC chromatogram for tylosin in contact with birnessite for 1 h (blue) and 7d (red), with absorbance measured at 287 nm.

Figure 3.9. The complexation of tylosin A with birnessite via the formation of a six-membered ring as a precursor to the formation of tylosin B.

Figure 3.10. The complexation of tylosin B with birnessite via the formation of a six-membered ring as a precursor to the formation of product 1.

Chapter 4: Conclusions and implications of this work.

4.1 Summary of results

The overall purpose of this work was to look at the behavior of tylosin and progesterone in bulk soils and the inorganic soil fraction. Chapter 2 described my work regarding the interaction of tylosin and progesterone with sterilized soils. Based upon an initial aqueous concentration of 3.0 µM, progesterone was found to reach quasi- equilibrium between the sorbed and aqueous phase after 24 h. While it appears that no estimates for the time required to reach quasi-equilibrium have been published before, this is within the time range estimated for similar compounds (Casey et al.; 2005; Khan et al., 2009; Lai et al., 2000). Tylosin reached quasi-equilibrium after 7 d with the same initial aqueous concentration. This is considerably longer than the times previously reported in the literature (Rabølle et al., 2000; Sassman et al., 2007; Zhang et al., 2011).

This longer time could be due to slow diffusion and sorption of tylosin within the glassy phase of the SOM, which would be retarded due to tylosin’s large size and degradation processes.

Soil sorption isotherms were completed for both tylosin and progesterone with five different soils. With concentrations ranging from 0.05-5.0 µM, progesterone exhibited strong, non-linear sorption with greater non-linearity observed at higher

90 concentrations. Partition coefficients were calculated using Freundlich transformations of the data; using an initial aqueous concentration of 0.5 µM, log Koc=2.94 ±0.10

(excluding Toolik peat). While I believe no previous sorption isotherms have been published for progesterone, these Koc values are comparable to those reported for similar

EDC compounds (Card et al., 2012; Lee et al., 2003; Sarmah et al., 2008). In addition to hydrophobic partitioning (predicted based upon progesterone’s high log Kow value), the non-linearity of the sorption isotherms suggest that progesterone undergoes specific interactions with the soil. The more linear behavior of progesterone at lower concentrations (0.05-0.5 µM) plus the fact that progesterone does not exhibit more non- linear behavior with the Alaska “burned” soil suggests there is an abundance of micropore sites and competition for sites within high-surface-area carbonaceous materials

(HSACM) does not occur. Enrofloxacin, an antibiotic of similar size as progesterone, has been found to sorb within the interlayer spacing of clays (Nowara, 1997). Because this sorption is mediated by a ketone functional group that is conserved in progesterone, I hypothesize that the observed non-linearity is due to sorption of progesterone within the interlayer spacing of clays within the soils.

Tylosin also underwent strong, non-linear sorption in contact with the soils.

Partition coefficients were calculated using Freundlich transformations of the data; using an initial aqueous concentration of 0.5 µM, the concentration dependent log Koc=2.95 ±

0.18 (excluding Toolik peat). These Koc values are higher than anticipated based upon the Kow value for tylosin (Kow=1.63[Boxall et al., 2006]). Non-linear behavior for tylosin in contact with soils has been previously reported; likewise, our Koc values are within the

91 range observed by others (Rabølle et al., 2000; Sassman et al., 2007; Zhang et al., 2011).

While non-linearity has previously been attributed to specific interactions of tylosin with soils or exclusion from micropores due to its large size, I hypothesize that the observed non-linearity is likely due to the degradation of tylosin in contact with the mineral fraction of the soils. This is based upon the fact that similar macrolide compounds undergo transformation in contact with metal oxides (Feitosa-Felizzola et al., 2009), and after mass balances were completed, only 60% of the tylosin added to the soils was accounted for by the time quasi-equilibrium was reached (as opposed to over 95% for progesterone).

In order to determine if tylosin undergoes transformation with the mineral fraction in soil, I investigated its interactions with a manganese and iron oxide in suspension (as described in Chapter 3). Tylosin underwent pseudo-first order degradation in contact with both goethite (ɑ-FeOOH) and birnessite (δ-MnO2), with faster degradation in contact with goethite (t1/2=27 h) than in contact with birnessite (t1/2=139 h). The longer half-life for tylosin with birnessite is likely due to a different mechanism of transformation. Based upon the change in absorbance at 221 and 287 nm over one week, tylosin in contact with goethite likely undergoes cleavage within its chromophoric region, as degradates lack absorbance at both detection wavelengths. This cleavage is likely mediated by interactions between the goethite surface and the ketone group at C9.

Because of the appearance of new product peaks over time at 287 nm, it is unlikely that tylosin is changed at the chromophoric region in contact with birnessite. Instead, I hypothesize that a six-membered ring structure is formed between the amine group, the

92 ether linkage between the mycaminose and mycarose ring, and the manganese oxide surface. The formation of this ring structure likely mediates cleavage of the ether bond, resulting in the formation of tylosin B (desmycosin); the formation of a similar ring structure could result in the formation of an additional product which could be desmycosin without a mycaminose group.

Additional experiments were carried out in order to investigate the effect of dissolved organic matter (DOM) on the rate of transformation of tylosin in contact with metal oxides. The addition of DOM was found to slow the transformation of tylosin in contact with goethite by a factor of 1.5. DOM has previously been found to have an inhibitory effect on the transformation of contaminants with metal oxides (Feitosa-

Felizzola et al., 2009; Pils et al., 2007; Klausen et al., 1997), likely due to either competitive sorption of DOM to the oxide surfaces or the reductive dissolution and subsequent sorption of reduced metal species. Tylosin in contact with birnessite was degraded three times faster in the presence of DOM. Although less frequently reported in the literature (Kang et al., 2004; Park et al., 1999), this could be due to the ability of Mn to oxidize phenol groups within the DOM to quinone or radical structures at a faster rate than tylosin. These phenol groups could then react with tylosin, increasing its apparent rate of transformation in contact with birnessite. In addition, less DOM could adsorb to birnessite as compared to goethite due to its negative surface charge, blocking fewer surface sites and allowing the tylosin transformation reaction to continue uninhibited.

4.2 Environmental significance

The interactions of organic contaminants with soils can have a profound impact on their environmental fate. In particular, their sorption to soils can affect their availability for microbial degradation, runoff to surface waters, abiotic degradation, or uptake into plants (Boxall et al., 2006; Thiele-Bruhn, 2005; Ingerslev et al., 2001). Each of these pathways can have varying environmental impacts; for example, higher concentrations of antibiotics in soils could result in antibiotic resistant soil microbial populations (Khachatourians, 1998) or hormones in runoff that enters streams could contribute to the feminization of aquatic wildlife (Brain et al., 2005).

Because of its importance in controlling environmental fate, it is necessary to understand the interaction of organic contaminants with soil. While previous studies have investigated the sorption of tylosin to soils, they have reported varying partition coefficients ((Rabølle et al., 2000; Sassman et al., 2007; Zhang et al., 2011). My calculated log Koc values are towards the lower end of the previously published values; hopefully, my data will help provide a better estimate for the actual amount of tylosin sorbed in agricultural soils. Further, the transformation of tylosin in contact with metal oxides is a previously unreported degradation pathway for tylosin in soils. This could have an impact on its environmental fate; for example, there is the possibility that soils with higher concentrations of mineral oxides will degrade tylosin at a faster rate, thereby reducing its half-life in soils. In addition, this transformation could prevent tylosin’s accumulation in soils, and thus its continual impact on soil ecosystems.

It is important to understand the environmental fate of progesterone because it has been implicated as a possible carcinogen (NIEHS, 2005). Despite this, limited work has been completed regarding the interactions of progesterone with soils. Determining the time required for progesterone to reach quasi-equilibrium with soils will help us understand this process relative to the time required to transport it to surface waters, where it has been detected (Kolpin et al., 2002; Kolodziej et al., 2007). Because progesterone undergoes strong sorption to soils, a greater fraction of will likely be retained in the soil compartment instead of being transported. Progesterone’s strong sorption could protect it from microbial decay or other transformation processes, allowing soil to become a legacy source of progesterone.

4.3 Future work

As identified in Chapters 2 and 3, several future avenues of work remain regarding this research. For example, the interactions of progesterone with clays should be studied in order to test the hypothesis that it undergoes sorption within the interlayer spacing of clays. In particular, sorption isotherms with swelling and single-layer clays should be completed in order to determine if greater sorption is observed with the swelling clays. In addition, relating the abundance of HSACM in our soils to the observed behavior of progesterone at lower concentrations could help provide evidence for or against the specific interaction of progesterone with these soil fractions and should be investigated in the future.

Because there is no previous work regarding the interaction of tylosin with metal oxides, there remains a great deal of work to be done regarding this phenomenon. For example, further replicates are required in order to determine if the addition of DOM does enhance the rate of tylosin degradation in contact with birnessite or whether an experimental error was observed. If additional trials do suggest that DOM enhances the rate of tylosin transformation, further experiments could be conducted regarding the interaction of tylosin and birnessite in the presence of phenol-containing compounds as a controlled representation of DOM and its functional groups. Enhanced degradation rates in the presence of the known quinone or phenolic compounds could help support the hypothesis that they serve as a mediator. Additional work is needed to determine the rate of transformation of tylosin in contact with birnessite under sterilized conditions.

While some work has been completed using LC-MS in order to identify transformation products for tylosin in contact with metal oxides, further work is necessary. While I hypothesize that tylosin B is formed in contact with birnessite based upon the elution patterns in the HPLC chromatograms, this is not a definitive identification. In addition, further analysis via LC-MS is needed in order to support or disprove my hypothesis that the second product peak formed from tylosin in contact with birnessite is tylosin B with the mycaminose sugar cleaved. Similarly, while evidence of degradation of tylosin in contact with goethite has been observed via LC-MS, no products have been definitively identified, but rather indirectly determined. Further analysis should be completed in order to identify transformation products, with the particular goal of determining if cleavage of the lactone ring within the chromophoric region does occur.

While outside the scope of this work, it would be beneficial to determine the environmental fate of tylosin’s transformation products. In particular, determining whether these products retain antimicrobial properties could help us understand their environmental impact and fate.

Complete List of References

Aga, D. S., O’Connor, S., Ensley, S., Payero, J. O., Snow, D., Tarkalson, D. 2005. Determination of the persistence of tetracycline antibiotics and their degradates in manure-amended soil using enzyme-linked immunosorbent assay and liquid chromatography-mass spectrometry. J. Agric. Food Chem., 53, 7165-7171.

Allaire, S. E., Del Castillo, J., Juneau, V. 2005. Sorption kinetics of chlortetracycline and tylosin on sandy loam and heavy clay soils. J. Environ. Qual. 35, 969-972

Anadón, A., Martinez-Larranaga, M. R. 1999. Residues of antimicrobial drugs and feed additives in animal products: regulatory aspects. Livest. Prod. Sci., 59, 183-198.

Anderson, A. D., Nelson, J. M., Rossiter, S., Angulo, F. J. 2003. Public health consequences of use of antimicrobial agents in food animals in the United States. Microb. Drug Resist., 9, 373-380.

Arikin, O. A., Rice, C., Codling, E. 2008. Occurrence of antibiotics and hormones in a major agricultural watershed. Desalination, 226, 121-133.

Arnold, C. G., Ciani, A., Muller, S. R., Amirbahman, A., Schwarzenbach, R. P. 1998. Association of triorganotin compounds with dissolved humic acids. Environ. Sci. Technol., 32, 2976-2983.

Arnon, S., Dahan, O., Elhanany, S., Cohen, K., Pankratov, I., Gross, A., Ronen, Z., Baram, S., Shore, L. S. 2008. Transport of testosterone and estrogen from dairy-farm waste lagoons to groundwater. Environ. Sci. Technol., 42, 5521-5526.

Baker, J. R., Mihelcic, J. R., Luehrs, D. C., Hickey, J. P. 1997. Evaluation of estimation methods for organic carbon normalized sorption coefficients. Water Environment Research, 69, 1360145.

Ball, W. P., Roberts, P. V. 1991. Long-term sorption of halogenated organic chemicals by aquifer material. 2. Intraparticle diffusion. Environ. Sci. Technol., 25, 1237-1249.

Bartelt-Hunt, S. L., Snow, D. D., Kranz, W. L., Mader, T. L., Shapiro, C. A., van Donk, S. J., Shelton, D. P., Tarkalson, D. D., Zhang, T. C. 2012. Effect of growth promotants on

98 the occurrence of endogenous and synthetic steroid hormones on feedlot soils and in runoff from beef cattle feeding operations. Environ. Sci. Technol., 46, 1352-1360.

Borch, T., Davis, J. G., Yang, Y.-Y., Young, R. B. 2009. Occurrence of steroid sex hormones in the Cache la Poudre river, and pathways for their removal in the environment: completion report no. 216. http://www.cwi.colostate.edu/publications/ cr/216.pdf

Boxall, A. B. A., Johnson, P., Smith, E.J., Sinclair, C.J., Stutt, E., Levy, L. S. 2006. Uptake of veterinary medicines from soils into plants. J. Agric. Food Chem., 54, 2288- 2297.

Brain, R. A., Wilson, C. J., Johnson, D. J., Sanderson, H., Bestari, K., Hanson, M. L., Sibley, P. K., Solomon, K. R. 2005. Effects of a mixture of tetracyclines to Lemna gibba and Myriophyllum sibricium evaluated in aquatic microcosms. Environ. Pollut., 138, 425- 442.

Burnison, B. K., Hartmann, A., Lister, A., Servos, M. R., Ternes, T., Van Der Kraak, G. 2003. A toxicity identification evaluation approach to studying estrogenic substances in hog manure and agricultural runoff. Environ. Toxicol. Chem., 22, 2243-2250.

Campagnolo, E. R., Johnson, K. R., Karpati, A., Rubin, C. S., Kolpin, D. W., Meyer, M. T., Esteban, J.E., Currier, R. W., Smith, K., Thug, K. M., McGeehin, M. 2002. Antimicrobial residues in animal waste and water resources proximal to large-scale swine and poultry feeding operations. Sci. Total Environ., 299, 89-95.

Card, M. L., Chin, Y. P., Lee, L. S., Khan, B. 2012. Prediction and experimental evaluation of soil sorption by natural hormones and hormone mimics. J. Agric. Food Chem., 60, 1480-1487.

Casey, F. X. M., Hakk, H., Simunek, J., Larsen, G. L. 2004. Fate and transport of testosterone in agricultural soils. Environ. Sci. Technol., 38, 790-798.

Casey, F. X. M., Simunek, J., Lee, J. Larsen, G. L., Hakk, H. 2005. Sorption, mobility, and transformation of estrogenic hormones in natural soil. J. Environ. Qual., 34, 1372- 1380.

Chandler, Y., Kumar, K., Goyal, S. M., Gupta, S. C. 2005. Antibacterial activity of soil- bound antibiotics. J. Environ. Qual., 34, 1952-1957.

Chang, P.-H., Li, Z., Jiang, W.-T., Jean, J.-S. 2009. Adsorption and intercalation of tetracycline by swelling clay minerals. Appl. Clay. Sci., 46, 27-36.

Chee-Sanford, J. C., Mackie, R. I., Koike, S., Krapac, I. G., Lin, Y.-F., Yannarell, A. C., Maxwell, S., Aminov, R. I. 2008. Fate and transport of antibiotic residues and antibiotic resistance genes following land application of manure waste. J. Environ. Qual.,38, 1086- 1108.

Chen, J., Gu, B., Royer, R. A., Burgos, W. D. 2003. The roles of natural organic matter in chemical and microbial reduction of ferric iron. Sci. Tot. Environ., 307, 167-178.

Chen, W.-R., Huang, C.-H. 2010(1). Adsorption and transformation of tetracycline antibiotics with aluminum oxide. Chemosphere, 79, 779-785.

Chen, W.-R., Ding, Y., Johnston, C. T., Teppen, B. J., Boyd, S. A., Li, H. 2010. Reaction of lincosamide antibiotics with manganese oxide in aqueous solution. Environ. Sci. Technol., 44, 4486-4492.

Chin, Y.-P., Miller, P. L., Zeng, L., Cawley, K., Weavers, L. K. 2004. Photosensitized degradation of bisphenol A by dissolved organic matter. Environ. Sci. Technol., 39, 5888-5894.

Chiou, C. T. Partition and adsorption of organic contaminants in environmental systems, 1st edition.; Wiley: Hoboken, NJ, 2002.

Chiou, C. T., Klie, D. E., Brinton, T. I., Malcolm, R. L., Leenheer, J. A., MacCarthy, P. 1987. A comparison of water solubility enhancements of organic solutes by aquatic humic materials and commercial humic acids. Environ. Sci. Technol., 21, 1231-1234.

Chun, C. L., Penn, R. L., Arnold, W. A. 2006 (1). Kinetic and microscopic studies of reductive transformations of organic contaminants on goethite. Environ. Sci. Technol., 40, 3299-3304.

Chun, S., Lee, J., Radosevich, M., White, D. C., Geyer, R. 2006 (2). Influence of agricultural antibiotics and 17ß-estradiol on the microbial community of soil. J. Environ. Sci. Health, Part B. Food Contam., Agric. Wastes., 41, 923-935.

Colucci, M., Bork, H., Topp, E., 2001. Persistence of estrogenic hormones in agricultural soils: I. 17ß-estradiol and estrone. J. Environ. Qual., 30, 2070-2076.

Cornell, R. M., Schwertmann, U. The Iron Oxides: Structure, Properties, Reactions, Occurrences, and Uses, 2nd ed.; Wiley-VCH; Weinheim, Germany, 2003.

Deng, Y., Dixon, J. B. Soil organic matter and organic-mineral interactions. In Soil mineralogy with environmental applications; Dixon, D. B., Schulze, D. G., Eds.; Soil Science Society of America, Inc., Madison, WI, 2002, 69-109.

100

Dixon, J. B., White, N. G. Manganese Oxides. In Soil mineralogy with environmental applications; Dixon, D. B., Schulze, D. G., Eds.; Soil Science Society of America, Inc., Madison, WI, 2002, 367-389.

Downs, R. T., Bartelmehs, K. L., Gibbs, G. V., Boisen, Jr., M. B. 1993. Interactive software for calculating and displaying X-ray or neutron powder diffractometer patterns of crystalline materials. Am. Mineral., 78, 1104-1107.

Essington, M. E., Lee, J., Seo, Y. 2010. Adsorption of antibiotics by montmorillonite and kaolinite. Soil Sci. Soc. Am. J., 74, 1577-1588.

European Agency for the Evaluation of Medicinal Products (EMEA). 1997. Committee for veterinary medicinal products: tylosin summary report. http://www.emea.europa.eu/ docs/en _GB/document_library/Maximum_Residue_Limits__Report/2009/11/WC50001 5760.pdf

European Agency for the Evaluation of Medicinal Products (EMEA). 1999. Committee for veterinary medicinal products: progesterone summary report. http://www.ema. europa.eu/docs/ en_GB/document_library/Maximum_Residue_Limits__Report/2011/07/ WC500108427.pdf

Feitosa-Felizzola, J., Hanna, K., Chiron, S. 2009. Adsorption and transformation of selected human-used macrolide antibacterial agents with iron (III) and manganese (IV) oxides. Environ. Pollut., 157, 1317-1322.

Figueroa, R. A., Mackay, A. A. 2005. Sorption of oxytetracycline to iron oxides and iron oxide-rich soils. Environ. Sci. Technol., 39, 6664-6671.

Food and Drug Administration (FDA). 2011. Steroid hormone implants used for growth in food producing animals; http://www.fda.gov/AnimalVeterinary/SafetyHealth/ ProductSafetyInformation/ucm055436.htm

Gaynor, M., Mankin, M. S. 2003. Macrolide antibiotics: binding sites, mechanism of action, resistance. Curr. Top. Med. Chem., 3, 949-961.

Government Accountability Office. 2008. GAO-08-944: Concentrated animal feeding operations: EPA needs more information and a clearly defined strategy to protect air and water quality from pollutants of concern. http://www.gao.gov/assets/290/280229.pdf

Government Accountability Office (GAO). 2011. GAO-11-801: Agencies have made limited progress in addressing antibiotic use in animals. http://www.gao.gov/assets/330/ 323090.pdf

101

Gu, C. Karthikeyan, K. G. 2005 (1). Sorption of the antimicrobial ciprofloxacin to aluminum and iron hydrous oxides. Environ. Sci. Technol., 39, 9166-9173.

Gu, C., Karthikeyan, K. G. 2005 (2). Interaction of tetracycline with aluminum and iron hydrous oxides. Environ. Sci. Technol., 38, 476-483.

Halling-Sørenson, B. 2000. Algal toxicity of antibacterial agents used in intensive farming. Chemosphere, 40, 731-739.

Hu, D., Coats, J. R. 2007. Aerobic degradation and photolysis of tylosin in water and soil. Environ. Toxicol. Chem., 26, 884-889.

Hu, C., Hermann, G., Pen-Mouratov, S., Shore, L., Steinberger, Y. 2011. Mammalian steroid hormones can reduce abundance and affect the sex ratio in a soil nematode community. Agric., Ecosyst., Environ., 142, 275-279.

Ingerslev, F., Toräng, L., Loke, M. L., Halling-Sørenson, B., Nyholm, N. 2001(1). Primary biodegradation of veterinary antibiotics in aerobic and anaerobic surface water simulation systems. Chemosphere, 44, 865-872.

Ingerselv, F., Halling-Sørenson, B. 2001(2). Biodegradability of metronidazole, olaquindox, and tylosin and formation of tylosin degradation products in aerobic soil- manure slurries. Ecotoxicol. Environ. Saf., 48, 311-320.

Johnston, C. T., Tombacz, E. Surface chemistry of soil minerals. In Soil mineralogy with environmental applications; Dixon, D. B., Schulze, D. G., Eds.; Soil Science Society of America, Inc., Madison, WI, 2002, 37-69.

Kaiser, K., Zech, W. 1997. Competitive sorption of dissolved organic matter fractions to soils and related mineral phases. Soil Sci. Soc. Am. J., 61, 64-9.

Kaiser, K., Guggenberger, G. 2003. Mineral surfaces and soil organic matter. Eur. J. Soil Sci., 54, 219-236.

Kang, K.-H., Dec, J., Park, H., Bollag, J.-M. 2004. Effect of phenolic mediators and humic acid on cyprodinil transformation in the presence of birnessite. Water Res., 38, 2737-2745.

Kay, P. Blackwell, P. A., Boxall, A. B. A. 2005. Transport of veterinary antibiotics in overland flow following the application of slurry to arable land. Chemosphere, 59, 951- 959.

Kenney, J., Fallert, D. 1989. Livestock hormones in the United States. National Food Review, 12, 21-24. 102

Khachatourians, G. G. 1998. Agricultural use of antibiotics and the evolution and transfer of antibiotic-resistant bacteria. Can. Med. Assoc. J., 159, 1129-1136.

Khan, B., Qiao, X., Lee, L. S. 2009. Stereoselective sorption by agricultural soils and liquid-liquid partitioning of trenbolone (17ɑ and 17ß) and trendione. Environ. Sci, Technol., 43, 8827-8833.

Kim, Y.-H., Heinze, T. M., Kim, S.-J., Cerniglia, C. E. 2004. Adsorption and clay- catalyzed degradation of erythromycin A on homoionic clays. J. Environ. Qual., 33, 257- 264.

Klausen, J., Haderlein, S. B., Schwarzenbach, R. P. 1997. Oxidation of substituted anilines by aqueous MnO2: effect of co-solutes on initial and quasi-steady-state kinetics. Environ. Sci. Technol., 31, 2642-2649.

Kjaer, J., Olsen, P., Bach K., Barlebo, H. C., Ingerslev, F., Hansen, M., Halling- Sorensen, B. 2007. Leaching of estrogenic hormones from manure-treated soils. Environ. Sci. Technol., 41, 3911-3917.

Kolok, A. S., Sellin, M. K. 2008. The environmental impact of growth-promoting compounds employed by the United States beef cattle industry: history, current knowledge, and future directions. Rev. Environ. Contam. Toxicol., 195, 1-30.

Kolodziej, E. P., Sedlak, D. L. 2007. Rangeland grazing as a source of steroid hormones to surface waters. Environ. Sci. Technol., 41, 3514-3520.

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

Kolz, A. C., Moorman, T. B., Ong, S. K., Scoggin, K. D., Douglass, E. A. 2005. Degradation and metabolite production of tylosin in anaerobic and aerobic swine-manure lagoons. Water Env. Res., 77, 49-56.

Kuchler, F., McClelland, J., Offutt, S. E. 1989. Regulating food safety: the cost of animal growth hormones. National Food Review, 12, 25-33.

Kuchta, S. L., Cessna, A. J., Elliott, J. A., Peru, K. M., Headley, J. V. 2009. Transport of lincomycin to surface and ground water from manure-amended cropland. J. Environ. Qual., 38, 1719-1727.

103

Kumar, K., Gupta, S. C., Baidoo, S. K., Chander, Y., Rosen, C. J. 2005. Antibiotic uptake by plants from soil fertilized with animal manure. J. Environ. Qual., 34, 2082-2085.

Lai, K. M., Johnson, K. L., Scrimshaw, M. D., Lester, J. N. 2000. Binding of waterborne steroid estrogens to solid phases in river and estuarine systems. Environ. Sci. Technol., 34, 3890-3894.

LaKind, J. S., Stone, A. T. 1989. Reductive dissolution of goethite by phenolic reductants. Geochim. Cosmochim. Acta, 53, 961-971.

Lange, I. G., Daxenberger, A., Schiffer, B., Witters, H., Ibarreta, D., Meyer, H. H. D. 2002. Sex hormones originating from different livestock production systems: fate and potential disrupting activity in the environment. Anal. Chim. Acta, 473, 27-37.

Li, H., Lee, L. 1999. Sorption and abiotic transformation of aniline and ɑ-napthylamine by surface soils. Environ. Sci. Technol., 33, 1864-1870.

Loftin, K. A., Adams, C. D., Meyer, M. T., Surampalli, R. 2008. Effects of ionic strength, temperature, and pH on degradation of selected antibiotics. J. Environ. Qual., 37, 378- 386.

López de Alda, M. J., Asunción, Gil, A., Paz, E., Barceló, D. 2002. Occurrence and analysis of estrogens and progestogens in river sediments by liquid chromatography- electrospray-mass spectrometry. Analyst, 127, 1299-1304.

Lorenzen, A., Hendel, J. G., Conn, K. L., Bittman, S., Kwabiah, A. B., Lazarovitz, G., Masse, D., McAllister, T. A., Topp, E. 2004. Survey of hormone activities in municipal biosolids and animal manures. Environ. Tox., 19, 216-225.

Lotrario, J. B., Stuart, B. J., Lam, T., Arands, R. R., O’Connor, O. A., Kosson, D. S. 1995. Effects of sterilization methods on the physical characteristics of soil: implications for sorption isotherm analyses. Bull. Environ. Contam. Toxicol., 54, 668-675.

Luo, J. Zhang, Q., Sulb, S. L. 2000. Mechanistic and kinetic studies of crystallization of birnessite. Inorg. Chem., 39, 741-747.

MacDonald, J. M. 2008. The economic organization of U.S. broiler production. United States Department of Agriculture Economic Information Bulletin Number 38. http://www.ers.usda.gov/ publications/eib38/eib38.pdf

Malcolm, R. L., MacCarthy, P. 1986. Limitations in the use of commercial humic acids in water and soil research. Environ. Sci. Technol., 20, 904-911.

104

McBride, M. B. 1987. Adsorption and oxidation of phenolic compounds by iron and manganese oxides. Soil Sci. Soc. Am. J., 51, 1466-1472.

McKenzie, R. M. Manganese oxides and hydroxides. In Minerals in soil environments; Dixon, J. B., Weed, S. B., Eds.; Soil Science Society of America, Inc.: Madison, WI, 1989, 439-467.

Meier, M. 1999(1). Sorption and fractionation of unaltered natural organic matter in surface water and groundwater samples obtained from a small watershed. M.S. thesis, School of Earth Sciences, Ohio State University, Columbus, OH.

Meier, M., Namjesnik-Dejanovic, K., Maurice, P. A., Chin, Y.-P., Aiken, G. R. 1999(2). Fractionation of aquatic natural organic matter upon sorption to goethite and kaolinite. Chem. Geol., 157, 275-284.

Mellon, M., Benbrook, C., Benbrook, K. L. 2001. Hogging it: estimates of antimicrobial abuse in livestock. Union of Concerned Scientists, Cambridge, MA. http://www.ucsusa. org/food_and_agriculture/science_and_impacts/impacts_industrial_agriculture/hogging- it-estimates-of.html

Mills, L. J., Chichester, C. 2005. Review of evidence: are endocrine-disrupting chemicals in the aquatic environment impacting fish populations?. Sci. Tot. Environ., 343, 1-34.

National Institute for Environmental Health Sciences (NIEHS). 2005. Report on carcinogens, Eleventh Edition. http://ntp.niehs.nih.gov/?objectid=72016262-BDB7- CEBA-FA60E922B18 C2540

Neale, P. A., Escher, B. I., Schäfer, A. I. 2008. pH dependence on steroid hormone – organic matter interactions at environmental concentrations. Sci. Tot. Env. 407, 1164- 1173.

Nichols, D. J., Daniel, T. C., Moore, P. A., Edwards, D. R., Pote, D. H. 1996. Runoff of estrogen hormone 17ß-estradiol from poultry litter applied to pasture. J. Environ. Qual., 26, 1002-1006.

Nowara, A., Burhenne, J., Spiteller, M. 1997. Binding of fluoroquinolone carboxylic acid derivatices to clay minerals. J. Ag. Food Chem., 45, 1459-1463.

Onan, L. J., LaPara, T. M. 2003. Tylosin-resistant bacteria cultivated from agricultural soil. FEMS Microbiol. Lett., 220, 15-20.

Osterberg, D., Wallinga, D. 2004. Addressing externalities from swine production to reduce public health and environmental impacts. J. Am. Public Health, 94, 1703-1708.

105

Paesen, J., Cypers, W., Pauwels, K., Roets, E., Hoogmartens, J. 1995. Study of the stability of tylosin A in aqueous solutions. J. Pharm. Biomed. Anal., 13, 1153-1159.

Pan, B., Ning, P., Xing, B. 2009. Part V – sorption of pharmaceuticals and personal care products. Environ. Sci. Pollut. Res., 16, 106-116.

Park, J.-W., Dec, J., Kim, J.-E., Bollag, J.-M. 1999. Effect of humic constituents on the transformation of chlorinated phenols and anilines in the presence of oxidoreductive enzymes or birnessite. Environ. Sci. Technol., 33, 2028-2034.

Pederson, Y., Yeager, M., Suffet, I. 2003. Xenobiotic organic compounds in runoff from fields irrigated with treated wastewater. J. Agric. Food Chem., 51, 1360-1372.

Pignatello, J. J., Xing, B. 1996. Mechanisms of slow sorption of organic chemicals to natural particles. Environ. Sci. Tech., 30, 1-11.

Pignatello, J.J. 1998. Soil organic matter as a nanoporous sorbent of organic pollutants. Adv. Colloid Interface Sci., 76-77, 445-467.

Pignatello, J.J., Lu, Y., LeBoeuf, E. J., Huang, W., Song, J., Xing, B. 2006. Nonlinear and competitive sorption of apolar compounds in black carbon-free natural organic materials. J. Environ. Qual., 35, 1049-1059.

Pils, J. R. V., Laird, D. A. 2007. Sorption of tetracycline and chlortetracycline on K- and Ca-saturated soil clays, humic substances, and clay-humic complexes. Environ. Sci. Technol, 41, 1928-1933.

Post, J. E., Veblen, D. R. 1990. Crystal structure determinations of synthetic sodium, magnesium, and potassium birnessite using TEM and the Rietveld method. Am. Mineral., 75, 477-489.

Rabølle, M., Spliid, N. H. 2000. Sorption and mobility of metronidazole, olaquindox, oxytetracycline and tylosin in soil. Chemosphere 40, 715-722.

Rubert, IV., K. F., Pedersen, J. A. 2006. Kinetics of oxytetracycline reaction with a hydrous manganese oxide. Environ. Sci. Technol., 40, 7216-7221.

Sakurai, K., Ohdate, Y., Kyuma, K. 1988. Comparison of salt titration and potentiometric titration methods for the determination of zero point charge (ZPC). Soil Sci. Plant. Nutr., 34, 171-182.

Sarmah, A. K., Northcott, G. L., Leusch, F. D. L., Tremblay, L. A. 2006. (1). A survey of endocrine disrupting chemicals (EDCs) in municipal sewage and animal waste effluents in the Waikato region of New Zealand. 106

Sarmah, A. K., Meyer, M. T., Boxall, A. B. A. 2006.(2). A global perspective on the use, sale, exposure pathways, occurrence, fate and effects of veterinary antibiotics (Vas) in the environment. Chemosphere, 65, 725-759.

Sarmah, A. K., Northcott, G. L., Scherr, F. F. 2008. Retention of estrogenic steroid hormones by selected New Zealand soils. Environ. Int., 34, 749-755.

Sassman, S. A., Lee., L. S. 2005. Sorption of three tetracyclines by several soils: assessing the role of pH and cation exchange. Environ. Sci. Technol., 39, 7452-7459.

Sassman, S. A., Sarmah, A. K., Lee, L. S. 2007. Sorption of tylosin A, D, and A-aldol and degradation of tylosin A in soils. Environ. Toxicol. Chem. 26, 1629-1635.

Schwarzenbach, R. P., Westall, J. 1981. Transport of nonpolar organic compounds from surface water to groundwater. Laboratory sorption studies. Environ. Sci. Technol, 15, 1360-1367.

Shore, L. 2009. Effects of steroid hormones on aquatic and soil organisms. In Hormones and pharmaceuticals generated by concentrated animal feeding operations: transport in water and soil; Pruden, A., Shore, L. S., Eds.; Springer, New York, pp. 85-93.

Smolen, J. M., McLaughlin, M. A., McNevin, M. J., Haberle, A., Swantek, S. 2003. Reductive dissolution of goethite and the subsequent transformation of 4- cyanonitrobenzene: role of ascorbic acid and pH. Aquat. Sci., 65, 308-315.

Song, M. K., Choi, S. H. 2000. Growth promoters and their effects on beef production. Asian Austral. J. Anim., 14, 1230-135.

Sparks, D. L. Environmental Soil Chemistry, 2nd ed., Elsevier, San Diego, California, 2003.

Spitz, R. N., Barton, J. E., Barteau, M. A., Staley, R. H., Sleight, A. W. 1986. Characterization of the surface acid-base properties of metal oxides by titration/displacement reactions. J. Phys. Chem., 90, 4067-4075.

Stein, H. H. 2002. Experience of feeding pigs without antibiotics: a European perspective. Anim. Biotechnol., 13, 85-95.

Stephany, R. W. 2010. Hormonal growth promoting agents in food producing animals. Handb. Exp. Pharmacol., 195, 355-368.

107

Ter Laak, T. L., Gebbink, W. A., Tolls, J. 2006. The effect of pH and ionic strength on the sorption of sulfachloropyridazine, tylosin, and oxytetracycline to soil. Environ. Toxicol. Chem. 25, 904-911.

Thiele-Bruhn, S. 2003. Pharmaceutical antibiotic compounds in soils – a review. J. Plant. Nutr. Soil Sci., 166, 145-167.

Thiele-Bruhn, S., Beck, I. C. 2005. Effects of sulfonamide and tetracycline antibiotics on soil microbial activity and microbial biomass. Chemosphere, 59, 457-465.

Tolls, J. 2001. Sorption of veterinary pharmaceuticals in soils: a review. Environ. Sci. Technol., 35, 3397-3406.

United States Department of Agriculture (USDA). 2000. Part I: Baseline reference of feedlot management practices, 1999. http://www.aphis.usda.gov/animal_health/nahms/ feedlot/downloads/feedlot99/Feedlot99_dr_PartI.pdf

Van Emmerik, T., Angove, M. J., Johnson, B. B., Wells, J. D., Fernandes, M. B. 2003. Sorption of 17 ß-estradiol onto selected soil minerals. J. Colloid Interface Sci., 266, 33- 39.

Weber Jr., W. J., McGinley, P. M., Katz, L. E. 1992. A distributed reactivity model for sorption to soils and sediments. 1. Conceptual basis and equilibrium assessments. Environ. Sci. Technol. 26, 1955-1962.

Werner, J. J., Chintapalli, M., Lundeen, R. A., Wammer, K. H., Arnold, W. A., McNeill, K. 2007. Environmental photochemistry of tylosin: efficient, reversible photoisomerization to a less active isomer, followed by photolysis. J. Agric. Food. Chem., 55, 7062-7068.

Westergaard, K., Muller, A. K., Christensen, S., Bloem, J., Sørenson, S. J. 2001. Effects of tylosin as a disturbance on the soil microbial community. Soil. Biol. Biochem., 33, 2061-2071

Ying, G-G., Kookana, R. S. 2005. Sorption and degradation of estrogen-like endocrine disrupting chemicals in soils. Environ. Toxicol. Chem., 24, 2640-2645.

Yonkos, L. T., Fisher, D. J., Van Veld, P. A., Kane, A. S., McGee, B. L., Staver, K. W. 2010. Poultry litter-induced endocrine disruption in fathead minnow, sheepshead minnow, and mummichog laboratory exposures. Environ. Toxicol. Chem., 29, 2328- 2340.

108

Zhang, H., Huang, C.-H. 2005. Oxidative transformation of fluoroquinolone antibacterial agents and structurally related amines by manganese oxide. Environ. Sci. Technol., 39, 4474-4483.

Zhang, H., Chen, W.-R., Huang, C.-H. 2008. Kinetic modeling of oxidation of antibacterial agents by manganese oxide. Environ. Sci. Technol., 42, 5548-5554.

Zhang, Q., Yang, C., Dang, Z., Huang, W. 2011. Sorption of tylosin on agricultural soils. Soil Science, 176, 407-412.

109