Antibiotic Sorption and Transport in Porous Media

MISSING IMAGE

Material Information

Title:
Antibiotic Sorption and Transport in Porous Media Effect of Solution Chemistry, Divalent Metals and Colloids
Physical Description:
1 online resource (114 p.)
Language:
english
Creator:
Chen, Hao
Publisher:
University of Florida
Place of Publication:
Gainesville, Fla.
Publication Date:

Thesis/Dissertation Information

Degree:
Doctorate ( Ph.D.)
Degree Grantor:
University of Florida
Degree Disciplines:
Soil and Water Science
Committee Chair:
Ma, Lena Q
Committee Members:
Jawitz, James W
Nkedi-Kizza, Peter
Gao, Bin

Subjects

Subjects / Keywords:
antibiotic -- cations -- colloid -- sorption -- transport
Soil and Water Science -- Dissertations, Academic -- UF
Genre:
Soil and Water Science thesis, Ph.D.
bibliography   ( marcgt )
theses   ( marcgt )
government publication (state, provincial, terriorial, dependent)   ( marcgt )
born-digital   ( sobekcm )
Electronic Thesis or Dissertation

Notes

Abstract:
Antibiotics have been frequently detected in soils and ground water; however, their transport behaviors in soils remain largely unknown.The fate and transport of antibiotics in soils are influenced by solution chemistry such as pH, ionic strength, and divalent medals. I examined the effect of solution chemistry on retention and transport of two antibiotics sulfamethoxazole(SMZ) and ciprofloxacin(CIP) under saturated sand media. Laboratory columns packed with quartz sand were used to test the effects of solution pH and ionic strength (IS). In general, SMZ manifested a much higher mobility than CIP. Almost all SMZ transported through the columns within one pore volume in deionized water, but no CIP was detected in the effluents under the same condition. Perturbations in solution pH and IS showed no effects on SMZ transport in the sand columns. When pH was increased to 9.5,however, ~93% of CIP eluted from the sand columns I then examined CIP retention and transport under the influence of Cu and Ca cations by either mixing CIP with Cu/Ca in solution or preloading CIP onto sand then using Cu/Ca solution to mobilize CIP. Though the amount of Fe/Al oxides on native sand surface was limited, it significantly impeded CIP transport, delaying CIP breakthrough curve.  In clean sand where Fe/Al oxides were removed, presence of Ca and Cu both significantly promoted CIP transport by reducing the retardation factor R. In native sand, due to CIP’s strong complexation ability with Fe/Al, only Cu promoted CIP transport. Compared to clean sand, native sand not only sorbed 10 times more CIP (50 mg kg-1)than clean sand.  Though native sand had limited Fe/Al oxides on surface, they were the major sites for CIP sorption via complexation with CIP’s carboxyl group. Strong sorption colloids like montomorillonite can facilitate CIP transport.  In short, the chemical property of antibiotics greatly influenced their fate and transport in porous media with SMZ being much mobile than CIP. Metal oxides on solid phase surface and solution chemistry including pH, IS and cations significantly influenced the fate of CIPin porous media.
General Note:
In the series University of Florida Digital Collections.
General Note:
Includes vita.
Bibliography:
Includes bibliographical references.
Source of Description:
Description based on online resource; title from PDF title page.
Source of Description:
This bibliographic record is available under the Creative Commons CC0 public domain dedication. The University of Florida Libraries, as creator of this bibliographic record, has waived all rights to it worldwide under copyright law, including all related and neighboring rights, to the extent allowed by law.
Statement of Responsibility:
by Hao Chen.
Thesis:
Thesis (Ph.D.)--University of Florida, 2012.
Local:
Adviser: Ma, Lena Q.
Electronic Access:
RESTRICTED TO UF STUDENTS, STAFF, FACULTY, AND ON-CAMPUS USE UNTIL 2013-12-31

Record Information

Source Institution:
UFRGP
Rights Management:
Applicable rights reserved.
Classification:
lcc - LD1780 2012
System ID:
UFE0044959:00001


This item is only available as the following downloads:


Full Text

PAGE 1

1 A NTIBIOTIC SORPTION AND TRANSPORT IN POROUS MEDIA: EFFECT OF SOLUTION CHEMISTRY DIVALENT METALS AND COLLOIDS By HAO CHEN A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2012

PAGE 2

2 2012 Hao Chen

PAGE 3

3 To my parents and grandparents

PAGE 4

4 ACKNOWLEDGMENTS I would like to sincerely thank my advisor, Dr. Len a Q. Ma, for her support, guidance, and encouragement during my graduate program at the University of Florida. I w as really blessed to be under her mentorship over the past four years. My research and lab oratory training was also influenced by the dedicate d efforts of my co chair Dr. Bin Gao. I am also very g rateful to him for his continued encouragement and support for all these years. I would also thank all my committee members, Dr. Peter Nkedi Kizza Dr. James W. Jawitz and Dr. Roy D. Rhue for their gen erous help and support during my journey as a graduate student at the University of Florida. Yin Dr. Wenchuan Ding, Xin Wang Anhui Huang Letuzia Oliveira Rujira Tisarum Jay Lessl Piyasa Ghosh Ky Gress Rui Liu Xiaol ing Dong and Yingjia Zhu for their ongoing professional and personal support The Biogeochemistry of Trace Metals group was like a huge family to me. I am also grateful to Huimin Sun Congrong Yu Yuan Tian Lei Wu Ying Yao Mandu Inyang and Yu Wang for providing a great environment to study and conduct research. I learned a lot from both groups. Finally, I extend my gratitude to my previous supervisors Dr. Yuanhua Dong and Ms. Qiong An, for their encourag ement and help during the application process to the University of Florida. Finally, I would like to express very special thanks to my parents for their tremendous and unconditional love and support. Only with their love and encouragement could I gain self confidence and ability to take on challenges and overcome difficulties in my life. With their constant encouragement and prayers I have been able to achieve my goal.

PAGE 5

5 My roommates and friends in Gainesville played a major role in creating a home away from home atmosphere. I would like to convey my gratitude to Yixing Zhu, Xiao ling Liao, Fang Wang, Xi Huang and all my friends.

PAGE 6

6 TABLE OF CONTENTS page ACKNOWLEDGMENTS ................................ ................................ ................................ .. 4 LIST OF TABLES ................................ ................................ ................................ ............ 9 LIST OF FIGURES ................................ ................................ ................................ ........ 10 LIST OF ABBREVIATIONS ................................ ................................ ........................... 12 ABSTRACT ................................ ................................ ................................ ................... 13 1 ANTI BIOTIC S IN THE ENVIRONMENT ................................ ................................ 15 Origin of Antibiotic in the Environment ................................ ................................ .... 15 Antibiotics in Soil ................................ ................................ .............................. 16 Antibiotics in Water ................................ ................................ ........................... 19 Characteristics and Sorption of Selected Antibiotics in Soils ............................ 21 Re search Objectives ................................ ................................ ............................... 22 2 EFFECTS OF PH AND IONIC STRENGTH ON SULFAMETHOXAZOLE AND CIPROFLOXACIN TRANSPORT IN SATURATED POROUS MEDIA ................... 26 Introduction ................................ ................................ ................................ ............. 26 Materials and Methods ................................ ................................ ............................ 28 Materials ................................ ................................ ................................ ........... 28 Analy sis of Antibiotics ................................ ................................ ....................... 29 Column Experiments ................................ ................................ ........................ 30 Modeling Antibiotic Transport in Saturated Porous Media ................................ 32 Results and Discussion ................................ ................................ ........................... 33 SMZ and CIP Transport in Water Saturated Porous Media .............................. 33 Effect s of pH on SMZ and CIP Transport ................................ ......................... 35 Effects of Ionic Strength on SMZ and CIP Transport ................................ ........ 36 Conclusions ................................ ................................ ................................ ............ 37 3 INTERACTIONS OF CU AND CA WITH CIPROFLOXACIN SORPTION AND DESORPTION ONTO SATURATED POROUS MEDIA ................................ ......... 45 Introduction ................................ ................................ ................................ ............. 45 Materials and Methods ................................ ................................ ............................ 48 Materials ................................ ................................ ................................ ........... 48 CIP Sorption onto Sand ................................ ................................ .................... 48 Flumequine Sorption onto Sand ................................ ................................ ....... 50 Analysis of Antibiotics and Cu and Ca ................................ .............................. 50 Results and Discussion ................................ ................................ ........................... 50 CIP Species in Solution ................................ ................................ .................... 50

PAGE 7

7 Characteristics of Sand ................................ ................................ .................... 51 CIP Sor ption Isotherm onto Clean and Native Sand ................................ ........ 52 CIP Sorption onto Sand Preloaded with Cu and Ca ................................ ......... 54 Premixing Cu and Ca with CIP Red uced CIP Sorption onto Clean Sand ......... 56 Premixing CIP with Cu and Ca changed CIP sorption onto native sand .......... 57 Comparison of CIP Sor ption with Probe Compound Flumequine ..................... 58 Environmental Implication ................................ ................................ ....................... 59 Conclusions ................................ ................................ ................................ ............ 60 4 INFLUENCE OF CU AND CA ON CIPROFLOXACIN TRANSPORT IN SATURATED POROUS MEDIA ................................ ................................ ............. 67 Introduction ................................ ................................ ................................ ............. 67 Materials a nd Methods ................................ ................................ ............................ 69 Materials ................................ ................................ ................................ ........... 69 Column Experiments ................................ ................................ ........................ 69 Impacts of Aque ous Ca and Cu on CIP Transport in Sand Media .................... 69 Impacts of Aqueous Ca and Cu on Mobilization of CIP Presorbed onto Sand Media ................................ ................................ ................................ ... 70 Modeling CIP Transport in Saturated Porous Media ................................ ........ 71 Results and Discussion ................................ ................................ ........................... 71 CIP Species in Solution and Characteristics of Sa nd ................................ ....... 71 CIP Transport in Saturated Sand Porous Media ................................ .............. 72 Presence of Cu and Ca promoted CIP transport in Clean Sand ....................... 74 Presence of Cu Promoted CIP Transport in Native Sand ................................ 76 Aqueous Cu and Ca Increased Transport of Presorbed CIP in Clean Sand .... 77 Sand ................................ ................................ ................................ .............. 78 Conclusions ................................ ................................ ................................ ............ 79 5 COLLOID FACILITATED CIP TRANSPORT IN SATURATED POROUS MEDIA .. 86 Introduction ................................ ................................ ................................ ............. 86 Materials ................................ ................................ ................................ ........... 88 Batch Adsorption Experiments ................................ ................................ ......... 88 Column Experiments ................................ ................................ ........................ 89 CIP Colloid Co transport in Sand Media ................................ ........................... 89 Colloids Mobilization of CIP Presorbed onto Sand Media ................................ 89 Results and Discussion ................................ ................................ ........................... 90 CIP Sorption Isotherm onto Sand, Kaolinite and Montmorillonite ..................... 90 CIP, Kaolinite and Montmorillonite Transport in Saturated Sand Column ........ 91 CIP Colloid Co transport in Sand Media ................................ ........................... 91 Transport of Presorbed CIP in Native Sand under the Influence of Colloids .... 92 Environmental Implication ................................ ................................ ....................... 93 Conclusions ................................ ................................ ................................ ............ 93 6 CONCLUSIONS AND FUTURE DIRECTIONS ................................ ...................... 99

PAGE 8

8 APPENDIX A ADSORPTION ISOTHERMS OF CIP TO SAND : A 1) NATIVE SAND, A 2) CLEAN SAND ................................ ................................ ................................ ....... 102 B SCANNING ELECTRON MICROSCOPE IMAGE OF NATIVE SA ND .................. 104 LIST OF REFERENCES ................................ ................................ ............................. 105 BIOGRAPHICAL SKETCH ................................ ................................ .......................... 114

PAGE 9

9 LIST OF TABLES Table page 2 1 Basic properties of sulfamethoxazole (SMZ) and ciprofloxacin (CIP), the pKa values are from Lucida et al. (2000) and Vazquez et al. (2001), respectively. ... 38 2 2 Summary of experimental conditions and model parameters (SMZ = sulfamethoxazole and CIP = ciprofloxacin) ................................ ......................... 39 3 1 Basic properties of ciprofloxacin (CIP) ................................ ................................ 61 3 2 Stability constant (log k) of CIP with Cu, Ca, Fe and Al ( 1996 ) ........................... 62 3 3 Basic property of sand ................................ ................................ ........................ 62 3 4 The impact of Ca or Cu on CIP sorption onto clean and native sand (3 g sand with 1 mg L CIP for 24 h) ................................ ................................ ................. 63 4 1 Transport model parameters and experiment con ditions for CIP transport in sand columns. ................................ ................................ ................................ 80 4 2 Partition coefficient of CIP onto clean and native sand after shaking 3 g of sand with 0.7 M CIP plus water, 1 mM Ca or 1 mM Cu for 24 h. ...................... 80

PAGE 10

10 LIST OF FIGURES Figure page 2 1 Transport of SMZ, CIP, and bromide in saturated sand columns under DI water conditions. ................................ ................................ ................................ 40 2 2 Distribution of retained CIP in the sand columns after extended DI water flushing. ................................ ................................ ................................ .............. 41 2 3 Effect of pH on the transport of SMZ and CIP in satu rated sand columns .......... 42 2 4 Effect of IS on the transport of SMZ in saturated sand columns at pH 5.6 and pH 9.5 ................................ ................................ ................................ ................. 43 3 1 CIP sorptio n coefficient Kd* by clean and native sand at 1 mg L CIP different cation concentrations. ................................ ................................ ........... 64 3 2 Sorption ability of sand of 0.01 mmol CIP or FQ in DI water, 1 mmol CaCl 2 or CuCl 2 solutio n ................................ ................................ ................................ ..... 65 3 3 Possible interactions in the system ................................ ................................ ..... 66 4 1 Possible interactions of CIP with clean and native sand. ................................ .... 81 4 2 Transport of bromide in native and clean sand and CIP in clean sand media under DI water condition. ................................ ................................ .................... 82 4 3 CIP transport in saturated sand m edia where it was mixed with:1 mM Cu in clean sand, 1 mM Ca in clean sand and Cu in native sand. ............................... 83 4 4 Mobilization of presorbed CIP onto sand after column was flushed with DI water, 1 mM Ca, and 1 mM Cu for 5 pore volumes ................................ ............ 84 4 5 Distribution of sorbed CIP in different layers of the saturated native sand column after flushed with DI water, 1 mM Ca or 1 mM Cu. ................................ 85 5 1 Transport of kaolinite and montmorillonite in sand media under DI water. ......... 94 5 2 Premixed CIP and kaolinite co transport in saturated sand medi a. .................... 95 5 3 Premixed CIP and montmorillonite co transport in saturated sand media. ......... 96 5 4 Mobilization of presorbed CIP onto sand after column was flushed with montmorillonite. ................................ ................................ ................................ .. 97 5 5 Mobilization of presorbed CIP onto sand after column was flushed with Kaolinite. ................................ ................................ ................................ ............. 98

PAGE 11

11 A 1 Native sand isotherm CIP sorption isotherm ................................ .................... 102 A 2 Clean sand isotherm CIP sorption isotherm ................................ ..................... 103 B 1 Scanning elec tron microscope Image of native sand ................................ ....... 104

PAGE 12

12 LIST OF ABBREVIATIONS CIP Ciprofloxacin FQ Flumequine FQs Fluoroquinolones HPLC High performance liquid chromatography ICP Inductively Coupled Plasma PHzpc Zero point of charge of metal oxides PKa Acid dissociation constant PPCPs Pharmaceuticals and personal care products SMZ Sulfamethoxazole

PAGE 13

13 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy ANTIBIOTIC SORPTION AND TRANSPORT IN POROUS MEDIA: EFFECT OF SOLUTION CHEMISTRY DIVALENT METALS AND COLLOIDS By Hao Chen December 2012 Chair: Lena Q. Ma Major: Soil and Water Science A ntibiotics ha ve been frequently detected in soils and groundwater; however, their transport behaviors in soils remain largely unknown. The fate and transport of antibiotics in soils are influenced by solution chemistry such as pH, ionic strength, and divalent medals I examined the effect of solution chemistry on retention and transport of two antibiotics sulfamethoxazole (SMZ) and ciprofloxacin(CIP) under saturated sand media Laboratory col umns packed with quartz sand were used to test the effects of solution pH and ionic strength (IS). In general, SMZ manifested a much higher mobility than CIP. Almost all SMZ transported through the columns within one pore volume in deionized water, but no CIP was detected in the effluents under the same condition. Perturbations in s olution pH and IS showed no effects on SMZ transport in the sand columns. When pH was increased to 9.5, however, ~93% of CIP elute d from the sand columns I then examined CIP retention and transport under the influence of Cu and Ca cations by either mixing CIP with Cu/Ca in solution or preloading CIP onto sand then

PAGE 14

14 using Cu/Ca solution to mobilize CIP. Though the amount of Fe/Al oxides on native sand surface was limited, it significantly impeded CIP transport, delaying CIP breakthrough curve. In clean sand where Fe/Al oxides were removed, presence of Ca and Cu both significantly promoted CIP transport by reducing the retardation factor R. In transport. Compared to clean s and, native sand not only sorbed 10 times more CIP (50 mg kg 1) than clean sand Though native sand had limited Fe/Al oxides on surface, Strong sorption colloids like m ontomorillonite can facilitate CIP transport In short the chemical property of antibiotics greatly influence d their fate and transport in porous media with SMZ being much mobile than CIP Metal oxides on so l i d phase surface and solution chemistry incl uding pH IS and cations significantly influence d the fate of CIP in porous media

PAGE 15

15 CHAPTER 1 ANTI BIOTIC S IN THE ENVIRONMENT In the past decade, there has been growing concern regarding the release of antibiotics in the soil and water system ( Vulliet et al. 2011 ; Lapworth et al. 2012 ) Antibiotics are widely used in health care and agricultural industries for the treatm ent and prevention of diseases. High volume of manufacturing and consumption of antibiotics results in their release into the soil and water system inevitably which may impose a potential risk on the eco systems ( Nygaard et al. 1992 ; Brown et al. 2006 ) Significant increases of a ntibiotic resistance genes have been observed in many circumstances globally, which is plausibly related to the widesprea d antibiotics in the environment Origin of Antibiotic in the Environment I n the USA, more than 25 million pounds of were manufactured in 2000; more than half of these antibiotics were used for living stock ( Threats 2009 ) European Federation of Anima l health has reported in 1999, more than ten thousand tons of antibiotics has been used in European area, 65%of these antibiotics was in human medicine, 29% was for veterinary usage, and 6% was used for animal for growth promoters ( (FEDESA) 199 7 ) In Europe animal growth promoter usage of antibiotics has already been banned recently; in USA this usage of antibiotics for animal growth still remain. Small amount antibiotics also ha ve been used for plant to control certain bacterial diseases of high value fruit, vegetable, and ornamental plants. In the USA, typically less than 0.5% of antibiotics applied for p l ant disease (McManus et al., 2002). Most of the unused antibiotics have been disposed into the sewage system. Since the sewage system is not designed specifically to remove antibiotics during

PAGE 16

16 sewage treatment so majority of the antibiotics were not treated ( Le Corre et al. 2012 ) These antibiotics will enter the environmental and reach surface water and subsurface water ( Lapworth et al. 2012 ) and even drinking water ( Vulliet et al. 2011 ) Antibiotics used for human or animal treatment or as growth promoters are excreted and end up in manure. Unmetabolized antibiotic substances are often entered the aquatic envir onment. For example, fluoroquinolone antibiotics, was found in in hospital effluent with concentration as high as 200 g L 1 Ampicillin was found in German hospital effluence with concentration between 20 and 80 g L 1 ( Brown et al. 2006 ; Martins et al. 2008 ) Antibiotics in S oil There are several environmental factors influencing the toxicity of antibiotics in soi ls including soil property and chemical characteristics of antibiotics ( Kemper 2008 ) One of the most important issues for antibiotics in soil is their fate and t ransport Antibiotics used for v eterinary use are excreted by the animals and often end up in subsurface via manure used as agricultural fertilizer The loads of antibiotics shed by maturing have been estimated up to kilograms per he ctare ( Kemper 2008 ) Often, antibiotics are released into the subsurface with slight transform ation for a relatively long time period ( Vulliet et al. 2011 ) Chemical and physical behavior of these antibiotics in the soil depends on the ir molecular structure. A ntibiotics are classified into ionized amphiphilic or amphoteric groups, and for this reason, antibiotics can interact with different soil component s Due to the ir properties, such as their molecular structure, polarity and solubility, the sorption and desorption of these antibiotics in soils differ significant ly ( Kmmerer 2001 ; Carrasquillo et al. 2008 ) Some antibiotics like ciprofloxacin (CIP) are charged in

PAGE 17

17 natural pH and slightly water soluble and as a result highly immobile in soils. However, these highly retarded antibiotics can still be mobilized in certain circumstance. For example, the transport of tetracyclines seems to be restricted to fast preferential and macropore flow or to b e facilitated by co transport with mobile colloids such as dissolved organic matter. Most antibiotics are sorbed to solid phase rapidly Their antibiotic activity is decreased by sorption h owever, that does not mean a complete elimination of the ir antimic robial activity ( Chander et al. 2005 ) E xperimental studies on the anti bacteri al activity of soil bound tetracycline and tylosin showed that even th ough these compounds are tightly adsorbed by clay particles, they remain active ( Chander et al. 2005 ) showing antimicrobial effects that may influen ce the selection of antibiotic resistant bacteria in the terrestrial environment. If the application of contaminated manure in soil exceeds the ir degradation rate, it will result in their accumulation. In general, t hese sorbed compounds serve as a reservoi r of pollutants and further contaminate groundwater by leaching There are several potential pathways for antibiotics to enter the natural water bodies including groundwater aquifers ( Jones et al. 2001 ; Lapworth et al. 2012 ) Leaching from soils (e.g., agricultural fields and landfills) is one of the most important pathways for antibiotic transmission to water bod ies. For example, large volumes of animal wastes (manure and slurry) containing high doses of antibiotics have been commonly applied to agricultural crop fields as a source of organic fertilizer to increase crop yields. Although some antibiotic may be reta ined by the soils, the rest can be transported to surface and groundwater through surface runoff and inf iltration

PAGE 18

18 Several studies have been conducted to investigate the mechanisms controlling the retention and release of antibiotics in soils ( Heberer et al. 2004 ; Lorphensri et al. 2007 ) Comprehensive investigations ranging from large scale o bservations in the field to bench scale studies with batch experiments have advanced the knowledge of environmental fate of antibiotics. Most of the previous studies, however, mainly focused on the ir sorption, desorption, and degradation processes in soils. Only few stu dies examined their transport dynamics in porous media. Therefore, additional investigations are necessary to improve current understanding of how flow dynamics affect the transport behavior of antibiotics in porous media. The decomposition of antibiotics in subsurface is driven by many factors. Photo degradation has been recognized for sulfonamide group and quinolone group antibiotics, but it is not an important effect in subsurface since the light is reduced there Antibiotic degradation in subsurface is mainly through microbial activity, especially enzymatic reactions, and oxidative decarbo xylation especially in Fe or Mn enriched soils Biodegradation in soil increases when manure or sludge with high numbers of micro organisms is added S oil serves as a vast reservoir for micro organisms ( Essington 2003 ) High numbers of bacteria are important in maintaining mineral immobiliz ation and decom position processes Antibiotics are mainly affected in two ways: the microbial community can be significantly changed by antibiotic activities; these environmental bacteria can acquire antibiotic resistance genes by exposing with the antibiotics Some soil micro organisms hav e natural tolerance genes for antibiotics. However, most bacteria, antibacterial resistance genes develop only when exposed to antibiotics

PAGE 19

19 ( Riesenfeld et al. 2004 ) For example soil micro organisms became resistant after the application of tetracycline contamin ated manu re Regarding this provoked antibiotic resistance genes an overview of the antibiotic resistance genes in soil bact eria is p resented by Riesenfeld et al ( 2004 ) Furthermore, in some cases, provoked resistance did not happen in soil s but directly in the manure then spread to the soil by fertilizing In Europe the analysis of numerous manure samples revealed high bacteria resistant characters against various an tibiotics. After manure fertiliz ation, tetracycline resistant bacteria increased in soil and groundwater, but declined after the cessation of slurry fertiliz ed soil within eight months The ability to take up the special genes from the environment occurs quite commonly among soil and water samples the role of natural transformation in the ex traction of environmental bacterial resistance genes. Mobile genetic elements and transferring resista nce were detected in microbial organisms of enviro nmental samples and pig manure. Antibiotics in W ater For the continued exponential growth in human popu lation, t he limited supplies of water resources are one of the most important global issues In recent years, the occurrence and fate of antibiotics in the nature water body have been subjects to many investigations carried out globally ( Zorita et al. 2009 ; Luo et al. 2011 ; Vulliet et al. 2011 ) More than 30 antib iotic s have been found in waste water treatment plant effluent samples, in surface waters and even ground and drinking water. The main source of the antibiotics and their metabolites or degradation products is animal husbandry and administered drugs ( Kemper 2008 ) These antibiotics reach the aquatic environment by the application of manure or slurry to areas used for agricultur e or by animals excreting directly on the land. S urface run off, driftage or leaching will facilitate the antibiotics to

PAGE 20

20 transport in deeper layers of soils The solubility of m ost of the antibiotics is relatively high as a result it can be carried by water flow ( Lapworth et al. 2012 ) There are several potential pathways for antibiotics to enter the natural water bodies including groundwater aquifers. Leaching from soils (e.g., agricultural fields and landfills) is one of the most important pathways for antibiotic transmission to water bodies ( Thiele Bruhn 2003 ; Overcash et al. 2005 ) example, animal wastes containing high doses of antibiotics have been commonly applied to agricultural crop fields, most of the antibiotics are washed into the soil and eventually can be transported to surface or groundwater ( Lapworth et al. 2012 ) Although some antibiotic may be retained by the soils, the rest can leachate from soil and be released into the subsurface, but depending on soil conditions it may seep into groundwater or spread laterally until it meets a stream or other surface water ( Thiele Bruhn 2003 ; Overcash et al. 2005 ) The influence of antibiotics on the resident bacterial community in marine environment was reviewed by Nygaard et al ( 1992 ) Although t he concentrations impacting aquatic bacteria are mainly unknown antibiotics remaining active against bacteria in the soil and water system have been documented In wastewater treatment pl ants, antibiotic resist genes have been detected. These antibiotic resist bacteria enter the environment vi a biosolids application or wastewater irrigation. It has been reported that aquatic species such as algae and daphnids has also been influenced by an tibiotics. Under laboratory experiment systems, most antibiotic compounds are persistent, with only few being partially biodegraded ( Li and Zhang 2010 ) Even those

PAGE 21

21 biodegredated antibiotics could still have some antibiotic effect Several studies have been conducted t o exam the occurrence o f antibiotics in nature water body and effluents f rom wastewater treatment plant s ( Zorita et al. 2009 ) T etracycline and ciprofloxacin are not usually expected in aquatic envi ronment due to the high sorption capacity of the soil However, these substances have been detected in low levels in surface water s amples and in higher levels in overland flow water globally ( Luo et al. 2011 ; Vulliet et al. 2011 ) Antibiotic S ulf onamides, macrolides and fluroquinolone are analyzed frequently in US Tetracycline was not detected because of their strong adsorption to organic matter and low solubility The presence of tetracycline resistant bacterial has been reported by Chee Sanf ord et al. ( 2001 ) in groundwater underlying two swine production facil ities. However, the contribution of antibiotic input from agricultural use is minor in the aquatic environment. The major part of antibiotic input is from human administration via hospital effluents or municipal wastewater ( Kmmerer ( 2001 ) C haracteristics and S orption of Selected A ntibiotics in S oils Sulfamethoxazole (SMZ) and ciprofloxacin (CIP) are typical antibiotics extensively used for both human and veterinary animals. They are among the most frequently detected antibiotics in streams and groundwater in the U.S. ( Kolpin et al. 2002 ; Barnes et al. 2004 ) SMZ is characterized as a low reactive and high mobile antibiotic ( Holten Ltzhft et al. 2000 ) while CIP is known to have a high affinity to soils ( Nowara et al. 1997 ; Thiele Bruhn et al. 2004 ) The different characteristi cs of CIP and SMZ make them good candidates to represent different antibiotics in the environment. The sorption of antibiotics to soils is controlled by several mechanisms, such as cation bridge interaction, hydrophobic effect, electrostatic interaction, s urface complexation,

PAGE 22

22 and ionic exchange. Among these mechanisms, electrostatic interactions are commonly observed for most of the antibiotics including SMZ in soils. Besides electrostatic adsorption, CIP sorption to soils is affected by surface complexati on and cation bridge interaction because it is capable of forming strong complexes with multivalent metal ions on grain surfaces ( Riley et al. 1993 ; Turel et al. 1994 ) The structure has a bidentate complex between metal ions and CIP through the carboxylic group. This character may be important because the abundance of metals in the environment. Divalent metals like Cu can be accumulated to a relatively high level in contaminated soils. For example, Muchuweti et al. ( 2006 ) reported that Cu concentra tions in surface soil increased to ~230 mg kg 1 through land application of fertilizers, sewage sludge, and wastewater irrigation. In particular, Cu has been used as a feed additive to stimulate animal growth. Li et al. ( 2007 ) found that Cu concentrations were 6.86 395 mg kg 1 in swine feeds and 57.0 2,017 mg kg 1 in pig feces. In addition to Cu, Ca is one of the most abundant metals in soils. Hence it is possible that CIP may coexist with Cu or Ca in soils. The formation of complexes between divalent cations and qui nolones and its impact on CIP sorption to soils have been studied ( Pei et al. ; Wallis et al. 1996 ; Park et al. 2002 ) However, little is known about the impact of Cu and Ca on CIP fate and transport in the environment. Therefore, it is important to examine their impacts on the transport behavior of CIP in porous media. Research Objectives Several studies have been conducted to investigate the mechanisms controlling the retention of antibiotics in soils ( Heberer et al. 2004 ; Lorphensri et al. 2007 ; Gielen et al. 2009 ) Recent r e search including large both field ( Heberer et al. 2004 ; Lorphensri et al. 2007 ) and bench scale studies ( T olls 2001 ; Pedersen et al. 2003 ; Gao and

PAGE 23

23 Pedersen 2005 ; Chefetz et al. 2008 ) have been conducted t o better understand the environmental fate of these antibioitics H owever, the majority of recent research mainly focus on sorption, desorption, and degradation processes of. T heir transport dynamics in porous media has only been examined in few studies ( Kay et al. 2005 ; Wehrhan et al. 2007 ; Srivastava et al. 2009 ; Unold et al. 2009 ; Unold et al. 2009 ) Therefore, additional investigations are necessary to improve current understanding of how flow dynam ics may affect the transport behavior of antibiotics in porous media. For example, laboratory batch sorption experiments showed that the interactions between antibiotics and soil particles strongly relied on the solution chemistry such as pH and ionic stre ngth (IS). To our knowledge, however, limited investigation has been conducted to evaluate the effects of pH and IS on the transport behaviors of antibiotics in porous media. The overall objective of this study is to investigate the fate and transport of selected antibiotics in saturated porous media In Chapter 2 I hypothesized that both flow chemistry and structural fuctional group of the antibiotics were the key factors controlling the transport behavior of selected antibiotics To test the hypothesi s, column experiments were conducted to compare the transport behavior of the two antibiotics under different pH and ionic strength. The goal of this study was to determine the influence of solution chemistrial on the transport of s ulfamethoxazole (SMZ) an d ciprofloxacin (CIP) in saturated porous media. Our specific objectives were to: 1) compare the retention and transport of SMZ and CIP in saturated porous media; 2) evaluate the effects of pH and IS on SMZ and CIP transport in saturated porous media; 3) e valuate the effects of ionic strength on the SMZ and CIP transport in saturated porous media; and 4) test whether existing solute

PAGE 24

24 transport models can be used to simulate the retention and transport of SMZ and CIP in saturated porous media. In Chapter 3 I hypothesized that both sand surface metal oxides and the divalent metal cations have potential influence on CIP sorption onto sand. To test the hypothesis, laboratory batch experiments were conducted to compare the behavior of CIP sorption with two sands ( clean and native) and two divalent metals. The goal was to determine the influence of divalent metal on CIP sorption mechanism of CIP in solution. Our specific objectives were to: 1) evaluate the effects of sand surface metal oxides on CIP sorption; 2) ev aluate the effects of divalent metals Cu and Ca on CIP sorption; and 3) evaluate the contribution of carboxylic and amine functional groups on CIP sorption onto sand In Chapter 4 I hypothesized that both minerals on sand surface and the persence of metal cations have influence on CIP tansport in porous media Column experimant was designed to better understand the chemodynamic behavors of CIP in sunsurface which is essential for comprehensive assessment of their potential environmental risk. To test the hypothesis, column experiments using saturated quartz sand were conducted to compare the transport behavior of CIP in two sands and in presence of two metals. The specific objectives were to: 1) evaluate the effects of metal oxides on sand surface on CIP t ransport in porous media, 2) compare the effects of metals Cu and Ca on retention and transport of CIP in porous media;3) compare the effects of metals Cu and Ca on mobilization of CIP presorbed on porous media; and 4) apply solute transport models to simu late the retention and tra nsport of CIP in porous media.

PAGE 25

25 In Chapter 5, I hypothesized that both minerals on sand surface and colloids have influence on CIP tansport in porous media Similar to chapter 4 we tested the influence of colloids on CIP transport. C hapter 6 provides a general conclusion and emphasizes the major findings of this study, as well as further research avenues.

PAGE 26

26 CHAPTER 2 EFFECTS OF PH AND IONIC STRENGTH ON SULFAMETHOXAZOLE AND CIPROFLOXACIN TRANSPORT IN SATURATED POROUS MEDIA Introducti on In the past decade, there has been growing concern over the transport of antibiotics in the environment ( Jones et al. 2001 ) Antibioti cs are widely used in health care and agricultural industries for the treatment and prevention of human and animal diseases. High volume of manufacturing and consumption of antibiotics inevitably results in their release into the natural water bodies, whic h may impose a risk on the ecosystems. Dramatic increase of antibiotic resistance genes have already been observed in many circumstances globally ( Johnson et al. 20 08 ; Kozak et al. 2009 ; Szczepanowski et al. 2009 ) There are several potential ways for antibiotics to enter the natural water bodies such as groundwater aquifers. Leaching from soils (e.g. agricultural fields and landfills) is one of the most important pathways for antibiotic transmission to water body ( Thiele Bruhn 2003 ; Overcash et al. 2005 ) For example, large volumes of animal wastes (manure and slurry) which contain high doses of antibiotics have been commonly applied to agricultural crop fields as a source of organic fertilizer to increase crop yields. Although some antibiotic may be retained by the soils, the rest can be transported to surface and groundwater through surface runoff and infiltration ( Thiele Bruhn 2003 ) Several studies have been conducted to investigate the processes control ling the retention and release of antibiotics in soils ( Heberer et al. 2004 ; Lorphensri et al. 2007 ; Gielen et al. 2009 ) Comprehensive investigati ons ranging from large scale observations in the field ( Heberer et al. 2004 ; Lorphensri et al. 2007 ) to bench scale studies with batch experiments ( Tolls 2001 ; Pedersen et al. 2003 ; Gao and Pedersen

PAGE 27

27 2005 ; Chefetz et al. 2008 ) have advanced the knowledge of environmental fate of antibiotics. Most of the previous studies, however, mainly focused on the sorption, desor ption, and degradation processes of antibiotics in soils. Only a few studies have examined their transport dynamics in porous media ( Kay et al. 2005 ; Wehrhan et al. 2007 ; Srivastava et al. 2009 ) Therefore, additional investigations are necessary to improve current understanding of how flow dynamics may affe ct the transport behavior of antibiotics in porous media. For example, laboratory batch sorption experiments showed that the interactions between antibiotics and soil particles strongly relied on the solution chemistry such as pH and ionic strength. To ou r knowledge, however, no investigation has been conducted to evaluate the effect s of pH and ionic strength on the transport of antibiotics in porous media. SM Z and CIP are two typical antibiotics extensively used for both human and veterinary animals. T hey are among the most frequently detected antibiotics in streams and groundwater ( Kolpin et al. 2002 ; Barnes et al. 2004 ) SMZ is characterized as a low reactive and high mobile antibiotic, while CIP is known to have a high affinity to soils ( Nowara et al. 1997 ; Thiele Bruhn et al. 2004 ) The different characteristics of CIP and SMZ make them good candidates to represent a range of antibiotics in the environment. The sorption of antibiotics to soils is controlled by seve ral mechanisms, such as cation bridge interaction, hydrophobic effect, electrostatic interaction, and ionic exchange. Among these mechanisms, hydrophobic and electrostatic interactions are commonly observed for most of the antibiotics including SMZ in soil s. In addition to electrostatic interactions, CIP sorption to soil is also affected by cation bridge interaction because it is capable of forming strong complexes with multivalent metal ions on grain surfaces

PAGE 28

28 ( Riley et al. 1993 ; Turel et al. 1994 ) Nowara et al. ( 1997 ) studied the sorption of CIP to clay minerals and suggested that cation bridging was the major sorption mechanism. The overarching goal of this study was to determine the mechanisms governing the transport of SMZ and CIP in water saturated porous media. It was our hypothesis that flow chemi stry as well as the chemical structure of the antibiotics are the key factors controlling their transport in saturated porous media. To test the hypothesis, laboratory experiments were conducted to compare the breakthrough behavior of the two antibiotics i n columns packed with quartz sand under different pH and ionic strength conditions. Previous studies have demonstrated that sand columns can be effectively used in a laboratory environment to mimic porous media under different physicochemical conditions ( Gao et al. 2004 ; Bradford et al. 2005 ; Bradford and Torkzaban 2008 ) Our specific objectives were to: 1) compare the retention and transport of SMZ and CIP in saturated porous media ; 2) evaluate the effects of pH on SMZ and CIP transport in saturated porous media ; 3) evaluate the effects of ionic strength on the S MZ and CIP transport in saturated porous media ; and 4) test whether existing solute transport models can be used to simulate the retention and transport of SMZ and CIP in saturated porous media Materials and Methods Materials SMZ (Sulfamethoxazole, ACS 7 32 46 6) and CIP (Ciprofloxacin, ACS 85721 33 1) were purchased from Applichem (Germany). The chemical structure and basic information of SMZ and CIP are listed in Table 2 1. All other chemicals were analytical reagents supplied by Fisher Scientific. SMZ s tock solution was prepared in methanol at a concentration of 100 mg/L, and CIP stock solution was prepared in acetonitrile at a

PAGE 29

29 use. Acetonitrile and methanol were used in this study as a cosolvent because the solubility of the two antibiotics was relatively low in deionized (DI) water. By diluting the stock solutions with DI water, the final SMZ and CIP concentrations used in this study were 200 g/L (with 0.1% cosolvent) and 50 g/L (with 0.5% cosolvent), respectively. All the glassware used in this study was acid washed before use. The ionic strength (IS) and pH of working solutions were adjusted with KBr and NaOH solutions. Quartz sand (45/30) sieved to a size range of 0.5 0.6 mm was used as the porous medium (Standard Sand & Silica Co., Davenport, FL) in the column experiments. The sand was washed sequentially by tap water and DI water, and then heated in Fisher Isotemp muffle furnace (Fisher Scientific) at 550C to rem ove trace organic impurities. Its zeta potential was determined by measuring the electrophoretic mobility of colloidal quartz sand ( Johnson et al. 1996 ) To obtain colloidal sand, a mixture containing 100 g of clean quartz sand and 200 mL DI water was ul trasonicated for 30 minutes. Aliquots of the quartz colloids were then removed and filtered through a 0.45 m filter. The filtrate was analyzed for electrophoretic mobility (U) with a ZetaPlus to convert electric ), which was 19.7 mV. Analysis of A ntibiotics Reversed phase high performance liquid chromatography (HP LC, Waters 2695, Milford, MA) equipped with a Phenomenex Gemini C18 column (150 mm 4.6 mm I.D., phase consisted of acetonitrile and sodium acetate aqueous solution (0.02 mol/L, pH

PAGE 30

30 adjusted to 4.75 with acetic acid to maximized column separation effect ) at a ratio of 25:75. A Waters 2489 ultraviolet detector was used to detect SMZ at a wavelength of was 50 R 2 ) > 0.99. The analysis of CIP was carried out using the HPLC equipped with a Nova Pak C18 column (150 mm 3.9 mm, Waters Millipore). The mobile phase consisted of acetonitrile and 0.5% phosphoric ac id at a ratio of 15:85. A fluorescence detector (Waters 2475) was used to detect CIP with the excitation and emission wavelengths at was 50 ficients ( R 2 ) > 0.99. Column E xperiments To avould air bubble t he quartz sand was wet packed into an acrylic column measuring 2.5 cm in diameter and 10 cm in height according to the procedures reported by Tian et al. ( Tian et al. 2010 ) A small amount of the quartz sand was poured gently into 8 mL of DI water standing at the bottom of the column until the sand surface was 0.5 to 1 cm below the water level. A polypropylene stir rod was used to stir the sand in the column. Approximately 8 mL DI water was then added to the column and the column were gently tapped several times to remove air bubble and ensure uniformity. This procedure was repeated several times until the column was packed to a height of 10 cm. Approximately 100 g of sand was used to pack one column with a porosity of 0.42 and a pore volume of 20.6 mL. A peristaltic pump (Masterflex L/S, Cole Parmer Instrument) was used to control the upward flow at a constant specific discharge of 0.2 cm/min. DI water was first pumped through the saturated column for about 2 h to remove impurities followed by

PAGE 31

31 work ing solutions overnight to stabilize the pore water pH and ionic strength (IS). For each antibiotic, the experiments were conducted with three different working solutions: DI water (pH=5.7, IS=0), high pH (pH=9.5, IS=0), and high IS (pH=5.7, IS=0.1 mM). On ce the outflow was stabilized, the breakthrough experiment was then initiated by switching the inflow to antibiotic solution (i.e., SMZ or CIP). For all experiments, the antibiotics were applied to the column at a 40 min pulse (i.e., 2 pore volumes PV), a nd then the column was flushed with antibiotic free solution for another 80 min or 4 PVs. The inflow concentrations of SMZ and CIP were 200 and 50 g/L, respectively, which were within the range of their typical concentrations detected in the environment ( Duong et al. 2008 ; Spongberg and Witter 2008 ; KarcI and BalcIog lu 2009 ; Zorita et al. 2009 ) Effluent samples were collected from the top of the column with a fraction collector (IS 95 Interval Sampler, Spectrum Chromatography) during sample injection and column fl ushing to analyze CIP and SMZ concentrations. All breakthrough experiments were performed in duplicate. Because all the duplications in this study had errors less than 1%, average breakthrough concentrations were reported. Two sets of breakthrough experime nts were conducted for CIP transport in the sand column under DI water conditions. After the pulse injection of CIP, the column was flushed with CIP free DI water, specific parameters for each experiment were summarized in Table 2 2. At the end of flushing the column was separated into 10 layers to determine the concentrations of retained CIP as a function of column depth. The sand was excavated under saturated conditions from top to bottom with a spatula in 1 cm increment and each increment was placed int o a small vial with 4 mL of H 3 PO 4 KH 2 PO 4 buffer and acetonitrile (ACN) solution (i.e., 27.2 g KH 2 PO 4 + 1.35 ml H 3 PO 4 in 1

PAGE 32

32 liter water with 1:1 ACN ). The excavated sand was then washed three times with the same amount of extraction solution. This extractio n method has been suggested to be the most effective way to extract fluoroquinolone antibiotics from different types of soils including sandy soils ( Uslu et al. 2008 ) It was verified by our extraction experiments, which showed 100% recovery of retained CIP from the sand CIP concentrations in the solutions were determined with HPLC. The CIP retention was then calculated for each sand section. Bromide was applied to the column as a conservative tracer for the breakthrough studies. The experimental procedures were the same as those used for S MZ and CIP. An ion chromatograph (ICS 90, Dionex Corporation) was used to determine bromide concentrations. Modeling Antibiotic Transport in Saturated Porous Media One dimensional advection dispersion equation coupled with reaction terms was used to simul ate the transport of CIP and SMZ in water saturated sand columns. I assumed that the interactions between the antibiotics and the sand grains in the column were affected by both reversible equilibrium and irreversible kinetic reactions ( Toride et al. 1995 ) The governing equation can be written as: ( 2 1) where C w is the concentration (SMZ or CIP) in pore water (g L 1 ); R is the retardation factor, which reflects the magnitude of equilibrium reactions in the sand column; D is the dispersion coefficient (cm 2 min 1 ); z is the coordi nate parallel to flow; v is the velocity of pore water (cm min 1 ); and k is the kinetic reaction rate constant (min 1 ).

PAGE 33

33 I solved the governing equation of the transport model numerically for a zero initial concentration, a pulse input boundary condition at the column inlet, and a zero concentration gradient boundary condition at the outlet. The model was first applied to the bromide breakthrough data to estimate dispersion coefficient ( D ). I assumed that the D of antibiotics was the same as that of the br omide tracer in the column. The transports of SMZ or CIP in the columns were quantified by identifying the best fit values of the R and k Results and Discussion SMZ and CIP Transport in Water Saturated Porous M edia The breakthrough curve of SMZ transport in the sand column in DI water (i.e., pH = 5.6 and IS = 0 mM) was similar to that of bromide, the conservative tracer (Figure 2 1). After the application to the sand column, SMZ was detected in the effluents around 1 PV. The breakthrough curve then quickl y moved up and reminded at a peak with further SMZ application. The SMZ concentrations decreased quickly to zero when the columns were flushed with SMZ free solution. Compared to the bromide breakthrough curve in the column, the breakthrough response of SM Z showed some delay, indicating slight retardation of the antibiotic and/or SMZ had more interation to sand than bromide in the porous media. The normalized peak concentration (C/C 0 ) of SMZ was close to unity and mass balance calculations showed that almos t all SMZ was transported through the sand column. The breakthrough experiment results demonstrated that SMZ was highly mobile in the water saturated porous media. This could be attributed to the fact that the interaction between SMZ and porous medium were repulsive because both surfaces were negatively charged under experimental conditions. The zeta potential of the quartz sand used in this experiment was 19.7 mV, confirming it was negatively

PAGE 34

34 charged. At pH of 5.6, about half of the SMZ is negatively ch arged and the other half is neutral (pKa 1 =1.6 and pKa 2 = 5.6, Table 2 1). Our data are consistent with the literature where many studies have observed the high mobility of SMZ in soils ( Delgado1 et al. 2005 ; Stoob et al. 2007 ) Simulations of the transport model matched well with the experimental breakthrough data of SMZ with R 2 of 0.99 (Figure 2 1). The best fit retardation factor R of SMZ in the column was 1.16 (Table 2 2), indicating slight retardation. The best fit kinetic reaction rate k of SMZ in the column was zero, suggesting no irreversible kinetic reactions of SMZ in the saturated porous media. Different from the transport of SMZ, the transport of CIP in the sand column showed no breakthrough response under DI water condition (pH = 5.6, IS = 0 mol/L) (Figure 2 1). After CIP was applied to the column, the column was flushed with DI water for 2 and 6 PVs (T able 2 2). For both cases, no CIP was detected in the effluents, suggesting no detectable CIP in the pore water within the column due to strong depositions on sand surfaces. At pH 5.6, most of the CIP (80%) in the solutions was positively charged (pKa 1 =6.2 and pKa 2 = 8.8, Table 2 1). The retention of CIP in the sand column therefore could be attributed to the electrostatic interactions (mainly columbic attraction) between positively charged CIP amine groups and negatively charged quartz sand surfaces and th e surface complexation as discussed before ( Gu and Karthikeyan 2005 ; Hari et al. 2005 ) Analyses of the retained CIP p rofile in the sand column after 4 PVs of flushing showed that almost all (~100%) the antibiotic was retained in the bottom 1 cm layer of the column (Figure 2 2), confirming the strong interactions between CIP and the sand grains. Additional flushing of 8 P Vs of the

PAGE 35

35 column with DI water mobilized a small portion of the retained CIP, with 95% of the CIP being retained in the bottom layer and the other 5% of CIP being found in the second 1 cm layer from the bottom (Figure 2 2). This result indicated that alth ough the sorption of CIP onto the sand surfaces was very strong, it was controlled at least partially by the reversible equilibrium reaction ( Toride et al. 1995 ) The strong CIP retention in this study was consistent with the literature examining its sorption behavior in soil systems ( Nowara et al. 1997 ; Gu and Karthikeyan 2005 ; Carrasquillo et al. 2008 ; MacKay and Seremet 2008 ) As discussed above, CIP sorption onto quartz sand in DI water systems was mainly controlled by two mechanisms: columbic attraction between cationic amine groups of CIP and negatively charged quartz sand surfaces ( Hari et al. 2005 ) and surface complexation between carboxyl groups of CIP and the trace level surfacial metal oxides (e.g., iron and aluminum oxides) ( Gu and Karthikeyan 2005 ) Effects of pH on SMZ and CIP T ransport The transport of SMZ at pH 9.5 in the porous media also exhibited high mobility (Figure 2 3a). This is because the SMZ solution at pH 9.5 was dominated by negatively charged species (SMZ ; Table 2 1). The transport of SMZ in the sand columns was almost identical for pH 9.5 and 5.6. Unlike SMZ, CIP transport at pH 9.5 showed much higher mobility than that of pH 5.6 (Figures 2 3b). Mass balance calculations showed that ~7% of CIP remained i n the sand column at pH 9.5. At pH 9.5, about 83.4% of CIP in the solution was negatively charged species (CIP Table 2 1), which limited its retention in the column. Our results indicated that pH dependent

PAGE 36

36 speciation of antibiotics play an important role in controlling their retention and transport in porous media. Simulations of the transport model matched well with the experimental breakthrough data of both SMZ and CIP with R 2 > 0.99 (Figure s 2 3a and 2 3b). Because there was no obvious difference in br eakthrough curves of SMZ between pH 5.6 and 9.5, the same parameters were used for both conditions. The best fit retardation factor R of CIP in the column at pH 9.5 was 1.37, which was higher than that of SMZ at 1.16 (Table 2 2). The best fit kinetic react ion rate k of CIP was > 0 (i.e., 0.007), suggesting part of the CIP might be retained in the sand column through irreversible kinetic reactions. The modeling results further confirmed that SMZ was more mobile than CIP in the saturated porous media under th e two pH conditions tested. Effects of I onic S trength on SMZ and CIP T ransport Solution IS showed little effect on SMZ transport in the sand columns under two pH conditions (Figure s 2 4a and 2 4b). The breakthrough responses of SMZ for two ISs were simil ar for pH 5.6 and pH 9.5. This is probably due to the fact that, under the experimental conditions, electrostatic interactions (ionic exchange and/or columbic attraction) between the SMZ and the quartz sand were negligible, and thus there was no competitio n effect from the cations or anions in the electrolyte. In a laboratory batch study, Gao and Pederson ( Gao and Pedersen 2005 ) also found that IS had no influence on SMZ sorption to clay particle s under these pH conditions. Similarly, solution pH showed almost no effect on CIP transport in the sand in the sand columns at pH 9.5 (Figure 2 5a), suggesting that electrostatic interactions between the CIP (CIP 83.4%) and the negatively charged quar tz sand were negligible. Even when the electrolyte concentration increased, there was no competition effect

PAGE 37

37 from the cations or anions. At pH 5.6, however, solution IS did show certain effect on CIP transport in the sand columns. Although no breakthrough w as observed at the end of the experiment, layer extraction results showed about 5% of the retained CIP was mobilized from front 1cm layer to the next layer (Figure 2 5b) at high IS. This result confirmed that electrostatic interactions could play an import ant role in controlling CIP transport in saturated porous media under low pH conditions, where the CIP is cationic and/or zwitterionic. Conclusions Laboratory column experiments were conducted to examine the effects of solution chemistry (i.e., IS and pH) on the retention and transport of two antibiotics in saturated porous media. Our results indicated that: 1) SMZ was more mobile in saturated porous media than CIP; 2) Solution pH played an important role in controlling the transport of CIP, but showed lit tle effect on the transport of SMZ under the experimental conditions tested; 3) Solution IS had little effects on SMZ transport under the two pH conditions (9.5 and 5.6) but slightly enhanced the mobility of the retained CIP in the sand column at pH 5.6; a nd 4) traditional solute transport model could be used to simulate the retention and transport of antibiotics in water saturated porous media.

PAGE 38

38 Table 2 1 Basic properties of sulfamethoxazole (SMZ) and ciprofloxacin (CIP), the pKa values are from Lucida et al. ( 2000 ) and Vazquez et al. ( 2001 ) respectively. structure MW pKa Speciation at pH 5.6 Speciation at pH 9.5 SMZ 253.3 pKa1=1.7 pKa2=5.6 Neutral (50%) and Anion (50%) Anion (100%) CIP 331.4 pKa,1=6.2 pKa,2=8.8 Cation (80%) and Zwitterion (20%) Anion (83.4%) and Zwitterion (16.6%)

PAGE 39

39 Table 2 2. Summary of experimental conditions and model parameters (SMZ = sulfamethoxazole and CIP = ciprofloxacin) No. Antibiotic IS pH Antibiotic apply time Flushing time R k 1 SMZ 0 5.6 40 min (2 PV) 80 min (4 PV) 1.16 0 2 SMZ 0.1 5.6 40 min (2 PV) 80min (4 PV) 1.16 0 3 SMZ 0 9.5 40 min (2 PV) 80 min (4 PV) 1.16 0 4 SMZ 0.1 9.5 40 min (2 PV) 80 min (4 PV) 1.16 0 5 CIP 0 5.6 40 min (2 PV) 80 min (4 PV) C an not be fitte d by proposed model 6 CIP 0 5.6 40 min (2 PV) 240 min (12 PV) Can not be fitted by proposed model 7 CIP 0 9.5 40 min (2 PV) 80 min (4 PV) 1.37 0.007 8 CIP 0.1 9.5 40 min (2 PV) 80 min (4 PV) 1.37 0.007

PAGE 40

40 Figure 2 1. Transport of SMZ, CIP, and bromide in sat urated sand columns under DI water conditions.

PAGE 41

41 Figure 2 2. Distribution of retained CIP in the sand columns after extended DI water flushing. The column was dissembled into 10 pieces of l cm sand layer.

PAGE 42

42 Figure 2 3. Effect o f pH on the transp ort of SMZ and CIP in saturated sand columns

PAGE 43

43 Figure 2 4. Effect of IS on the transport of SMZ in satur ated sand columns at pH 5.6 and pH 9.5

PAGE 44

44 Figure 2 5. Effect of IS on the transport of CIP in saturated sand columns: (a) breakthrough curves at pH 9.5 and (b) distribution of retained CIP in the columns at pH 5.6

PAGE 45

45 CHAPTER 3 INTERACTIONS OF CU AND CA WITH CIPROFLOXACIN SORPTION AND DESORPTION ONTO SATURATED POROUS MEDIA Introduction Antibiotics are widely used in health care and agricultural indust ries for the treatment and prevention of human and animal diseases. During the past decade, there has been growing concern about the release of antibiotics in the environment as these antibiotics are designed to be refractory to biodegradation and to act e ffectively even at low doses ( Jones et al. 2001 ) Ciprofloxacin (CIP) is one of the most widely prescribed fluoquinolon antibiotic, also the main metabolite of enrofloxacin. The sources of CIP in the environment include land application of sewage sludge, wastewater irrigation, and disposing of expired pharmaceutical prescriptions ( Golet et al. 2002 ) making CIP of increasingly environmental concern. It has been detected frequently in streams and groundwater ( Kolpin et al. 2002 ; Barnes et al. 2004 ) with the concentrations ranging from ng L 1 to mg L 1 Larsson et al ( 2007 ) reported that, in the wastewater treatment plant effluents from a pharmaceutical industry, CIP concentrations are 28 to 31 mg L 1 CIP is k nown to be not readily biodegradable ( Girardi et al. 2011 ) and has high sorption affinity onto soils ( Nowara et al. 1997 ; Thiele Bruhn et al. 2004 ; Vasudevan et al. 2009 ) so up to ppm levels of CIP can be accumu lated in soils. Therefore, the soil can act as a reservoir of CIP and other antibiotics ( Zorita et al. 2009 ) Hence it is important to examine the sorption and t ransport behaviors of CIP in soils. Several mechanisms have been proposed for sorption of fluoquinolon antibiotics onto soils. These include columbic attraction (cation exchange and cation bridging) and surface complexation ( Gu and Karthikeyan 2005 ; Otker and Akmehmet BalcIoglu 2005 ;

PAGE 46

46 Trivedi and Vasudevan 2007 ) CIP sorption onto soils occurs via columbic attraction of its cationic amine moiety ( NH2 + ) to negatively charged clay surface ( Hari et al. 2005 ) Cation exchange occurs when cations in the solution replaces positively charged CIP on solid surface. Cation bridging o ccurs when CIP is sorbed onto solid surface via columbic attraction of its carboxyl group ( COO ) to sorbed cations ( Nowara et al. 1997 ; Pei et al. 2009 ) In addition, surface complexation occurs between COOH group and surficial Fe/Al oxides in soils ( Gu and Karthikeyan 2005 ; Hyun and Lee 2005 ) Hence, CIP can be sorbed onto soils via cation exchange, cation bridging and surface complexation ( Vasudevan et al. 2009 ) One appro ach to determine the sorption mechanisms for complex organic sorbate molecules with multiple functional groups such as CIP is through sorption of simple probe compounds with only one functional group ( MacKay and Seremet 2008 ) In the case of CIP, structurally similar probe compounds of similar size are readily available. Sorption of the flumequine (FQ) with carboxylic group can help discern CIP comp lexation interactions with solid phase or sorbed cations. influence of coexisting cations. Since CIP can exist as a cation, anion as well as zwitterion, it is important to consider the impact of cations. M etal cations in the environmental with relatively high concentration can impact CIP sorption by acting as competitor for its cationic moiety ( NH 2 + ) acting as a bridge connecting its carboxyl group ( COO ) to the negatively charged site, or occupying the Fe/Al oxides sorption site to reduce its surface complexation with CIP ( Wallis et al. 1996 ; Park et al. 2002 ) F ew studies focused on the environmental fate of CIP with coexisting cations. Pei et al ( 2009 ) reported that Cu

PAGE 47

47 increased CIP sorption onto both kaolinite and montmorillonite under certain pH range, but Ross and Ri ley ( 1992 ) demonstrated that Cu substantially increased the solubility and mobility of fluoquinolone antibiotic. Although it has been demons trated many cations have complexation ability to CIP ( Turel et al. 1996 ) past research tested only copper. Therefore, it is necessary to investigate the impacts of other metal cations on CIP sorption and transport in the environment. Fe/Al hydroxides are important mineral components in soils, especially in highly weathered soils ( Essington 2003 ) Even present in limited qua ntities in soils, they are the major sorption sites for both organic and inorganic contaminants as they contain highly reactive surfaces ( Huang et al. 1977 ; Violante et al. 2003 ) Fe/Al hydroxides can develop both positive and negative surface charge depending on solution pH. Hence, these minerals can exert a profound influence on the fate and behavior of contami nants in the soil. Sorption studies with pure minerals based on Fourier transform infrared (FTIR) spectroscopy showed that the carboxylic acid group participates CIP sorption onto Fe/Al oxides ( Gu and Karthikeyan 2005 ) T o our knowledge, few investigations looked at the sorption of fluoroquinolone antimicrobials on Fe/Al oxides under the influence of coexisting cations ( Pei et al. 2009 ; Guaita et al. 2011 ) In present study I hypothesized that both sand surface characteristics and the co existence of divalent metal cations were the key factors controlling CIP sorption onto quartz san d. To test the hypothesis, batch experiments were conducted to compare the sorption behavior of CIP with two sands and two divalent metals. The goal of this study was to determine the mechanisms governing the sorption mechanism of CIP when co existing with divalent metals in solution. Our specific objectives were to: 1) evaluate the

PAGE 48

48 effects of sand surface metal oxides on CIP sorption; 2) evaluate the effects of divalent metals Cu and Ca on CIP sorption; and 3) evaluate the contribution of carboxylic and am ine functional groups on CIP sorption onto sand. Materials and Methods Materials CIP (ACS 85721 33 1) was purchased from Applichem (Darmstadt, Germany). Its chemical structure and basic information are in Table 3 1. All other chemicals were of analytical grades from Fisher Scientific (Pittsburgh, PA). CIP stock solution was prepared in DI water at a concentration of 40 mg L 1 The stock solutions were stored at 4C in darkness. The solutions were prepared in deionized (DI) water and glassware was acid wash ed before use. Quartz sand (45/30) from Standard Sand & Silica Co. (Davenport, FL) was sieved to a size range of 0.5 0.6 mm. To remove metal oxides on the sand surface, the native sand was first washed with tap and DI water, and then heated in 70% nitric a cid at 90C for 5 h. The sand was then washed with DI water to remove the acid and was referred to as clean sand The sand has been used for column experiment, to better understand its interactions with CIP during transport; its sorption behaviors were in vestigated. CIP Sorption onto Sand CIP sorption isotherms onto native and clean sand were conducted using 50 mL polytetrafluoroethylene centrifuge tubes. Each vessel was filled with 3.00 g of sand and 30 mL of CIP at 9 different concentrations: 0.05, 0.1 0.2, 0.4, 0.6, 0.8, 1.0, 2.5 and 5.0 mg L 1 The vessels were shaken for 24 h. Preliminary experiment showed that it reached equilibrium within 24 h (data not shown). The suspensions were centrifuged at

PAGE 49

49 5,000 g for 10 min to separate the solids from the liquid phase. Aliquots of supernatant were withdrawn to determine CIP concentrations in solution using HPLC (Waters 2695, Milford, MA). Solid phase concentrations were obtained through mass balance calculations. The impact of preloading Cu and Ca onto sand on CIP sorption by sand was investigated by shaking 3.00 g of sand with 30 mL of 1 mM CuCl 2 or CaCl 2 in a 50 mL centrifuge tube on a shaker for 24 h at room temperature. The Cu or Ca loaded sand was washed with DI water 5 times to remove soluble Ca and Cu. The sorbed Ca or Cu on the sand was an alyzed by ICP AES (Plasma 3200, Perkin Elmer Crop, MA) after digesting with HNO 3 /H 2 O 2 hot block digestion procedure ( EPA ) The Cu/Ca pre loaded sand was then shaken with 30 mL of 1.0 mg L 1 of CIP for 24 h. The suspensions were then centrifuged and the supernatants were collected for CIP analysis using HPLC. The impact of coexisting Cu and Ca on CIP sorption by sand was investigated by shaking vessels filled with 3.00 g of sand and 30 mL of 1.0 mg L 1 CIP at 6 different Cu:CIP or Ca:CIP molar ratios: 1, 10, 100, 500, 1,000, and 1,500. Different cations are exist in soil solution so the complexation effect of these cations may add up. Some cations like Fe with extremely high complexation ab ility can be equivalent to much higher concentration of low complexation ability cations like Ca ( Turel et al. 1996 ) To better understand the total effect of all the cations h igh concentration catioins were also used The vessels were shaken for 24 h and centrifuged at 5,000 g for 10 min to separate the solids from the liquid phase. Aliquots of supernatant were withdrawn for CIP analysis using HPLC. Solid phase concentrations were obtained through mass balance calculat ions.

PAGE 50

50 Flumequine Sorption onto Sand Since flumequine (FQ) contains a carboxyl group ( COOH ) similar to CIP, it was used to investigate if the carboxyl group was involved in CIP sorption onto sand. It was done by shaking vessels filled with 3.00 g of sand and 30 mL of 3.3 mg L 1 CIP or 2.6 mg L 1 FQ (equal molar concentration at 0.01 mM ) in DI water, and 1 mM CaCl 2 or CuCl 2 solution. The vessels were shaken for 24 h and centrifuged at 5,000 g for 10 min to separate the solids from the liquid phase. Aliquot s of supernatant were withdrawn for CIP and FQ analysis using HPLC. Solid phase concentrations were obtained through mass balance calculations. Analysis of A ntibiotics and Cu and Ca The analysis of CIP and FQ was carried out using a HPLC (Waters 2695, Mil ford, MA ) equipped with a Nova Pak C18 column (150 mm 3.9 mm, Waters Millipore). The mobile phase consisted of acetonitrile and 0.5% phosphoric acid at a ratio of 15:85. A fluorescence detector (Waters 2475) was used to detect CIP and FQ with the excita tion and emission wavelengths at 278 and 445 nm, respectively. The CIP detection limit was 1 and the linear range was 50 1 with correlation coefficients ( R 2 ) > 1 and the linear range was 100 10,0 1 with correlation coefficients ( R 2 ) > 0.99. Concentrations of Ca or Cu were analyzed by ICP AES (Plasma 3200, Perkin Elmer Crop, MA). The detection limit of this method for Cu and Ca was 0.02 mg L 1 and 0.1mg L 1 respectively. Results and Discuss ion CIP Species in Solution Since CIP species impact its interactions with Cu, Ca and sand, it is important to know its speciation in solution. CIP can exist in three forms (cationic, zwitterionic, and

PAGE 51

51 anionic forms) in DI water with two proton binding si tes (carboxyl and piperazinyl groups; Table 3 1). Under experiment condition of this study at pH 5.6, CIP existed as cationic form (80%) and zwitterionic form (20%). In comparison, >99% Cu and Ca were present as free cations (Visual MINTEQ). When CIP coexi sts with metal cations in solution, it can complexe with Cu or Ca via its carboxyl and keto groups to form metal CIP or metal CIP 2 complexes. In this study, since ratios of Cu or Ca to CIP were much greater than 1, so only 1:1 metal CIP was considered ( Li et al. 1994 ; Turel et al. 1996 ) Cu 2+ + CIP CIP + (3 1 ) Ca 2+ + CIP CIP + (3 2 ) Based on the complexation stability constants of Cu and Ca with CIP (log k = 14.7 and 11.3; Table 3 1b), CIP interacted with Cu more strongly than Ca. Since fluorescence quenching ability can ( Park et al. 2007 ) the complexation ability of Cu or Ca was further tested by fluorescence quenching experiment. As suming the florescence intensity of 200 g L 1 CIP in DI water was 100, the florescence intensity in 0.1 mM Ca and 0.1 mM Cu were 67 and 42 (data not shown), indicating both Cu and Ca were effective in fluorescence quenching, with Cu being 60% stronger tha n Ca. Compared to Cu/Ca, the complexation ability of CIP to Fe/Al is much stronger, forming Fe/Al CIP 3 (log k = 46.9 and 43.6; Table 3 1b). Characteristics of Sand Compared to soils, sand is more homogenous and much simpler with limited functional groups. So sand is a good material to investigate part of the sorption mechanisms of organic contaminants. However, even with simple materials like sand,

PAGE 52

52 the sorption mechanisms of CIP onto sand are complicated. Clean sand is composed of mostly SiO 2 with its s orption capacity coming from negatively charged broken edges ( SiO ) ( Essington 2003 ) In addition to SiO native sand was coated with Fe/Al oxides (Table 3 3 ). At pH 5.6, both native and clean sand carried negative charges with PZC (point of zero charge) being 5.0 5.1 (Table 3 3 ). The PZC for Fe/Al oxides is 10 ( Es sington 2003 ) so locally the broken edges of Fe/Al oxides carried positive charges ( Fe/AlOH + ). In addition, the Fe/Al oxides on the broken edges with high Gibbs free energy are more reactive in surface complexation with CIP ( Yost et al. 1990 ; Molis et al. 2000 ; Duckworth and Martin 2001 ) Trivedi et al. ( 2001 ) demonstrated that sorption of Cu as a transition metal is more likely to bind to high affinity site on goethite, while Ca as an alkaline earth metal bind to low affinity site ( Trivedi et al. 2001 ; Violante et al. 2003 ) CIP S orption I sotherm onto C lean and N ative S and Both clean sand and native sand were able to sorb CIP, but their ability differed substantially. The sorption of CIP onto native sand was significantly higher than clean sand. CIP sorption data fit well with Langmuir equation (R 2 = 0.9986 0.9998 ; data Show as Appendi x ) suggestive of monolayer sorption The best fit value of maximum sorption capacity from Langmuir model for native sand was 50 mg kg 1 which was 10 times greater than that of clean sand 5 mg kg 1 Moreover, the affinity of binding sites calculated from Langmuir model on the native sand (1.95 L g 1 ) was 6 times greater than that on the clean sand (0.33 L g 1 ). Though clean sand mainly consists of pure silica oxides, it still sorbed some CIP, which was probably attributed to columbic interaction. At pH o f 5.6, CIP was ~80%

PAGE 53

53 positivel y charged and 20% zwitterionic (pka 1=6.2 and pka2 = 8.8; Table 3 1 ). Hence, CIP sorption onto clean sand could be attributed to columbic attraction between positively charged CIP amine groups ( NH 2 + ) and negatively charged san d surfaces ( S 3) Compared to clean sand, native sand not only provided more sorption sites but also had higher sorption affin ity. This was mainly attributed to Fe/Al oxides on the surface ( Fe/AlO + ), which amounted to 167 and 1,087 mg kg 1 (Table 3 4 ). Its surface characteristics based on SEM was consistent with the chemical analysis, EDS spectra also indicated that the native sand contained substantial amounts of Fe/Al (data Show as Appendix ). Previous study indicate surface complexation of Fe/Al oxides wit h CIP was with its carboxylic acid group ( Gu and Karthikeyan 2005 ) Assuming all the extra sorption capacity of native sand was contributed to Fe/Al oxides with equal sorption ability, their sorption capacity was calculat ed at 33g kg 1 This suggests the sorption capacity of Al/Fe oxides on sand surface is comparable to that of the pure phase ( Gu and Karthikeyan 2005 ), with sorption capacity of Al oxides at 21.8 g kg 1 and Fe ox ides 13.5 g kg 1 Based on the above discussion, CIP sorption onto sand in DI water was mainly controlled by two mechanisms (Table 3 2): 1) columbic attraction between cationic amine groups of CIP and negatively charged sand surface ( Hari et al. 2005 ) and 2) su rface complexation between anionic carboxyl groups of CIP and Fe/Al oxides on native sand ( Gu and Karthikeyan 2005 ) SiO + + NH 2 SiO NH 2 CIP ( 3 3) Fe/AlO + + COO Fe/AlO COO CIP ( 3 4)

PAGE 54

54 CIP S orption onto S and P reloaded with Cu and Ca To better understand the sorption mechanisms of CIP onto sand, the sand was preloaded with Cu or Ca. The amounts of Cu sorbed onto native and clean sand wer e 8 and 1 mg kg 1 whereas those of Ca were below the detection limit of ICP (data not shown). The higher sorption ability of Cu than Ca for both sands could be explained by eous solutions. The higher sorption of Cu onto native sand than clean sand was mostly attributed to its higher cation exchange capacity (CEC = 0.28 vs. 0.05 cmol kg 1 ; Table 3 1c). However, the amount of Cu sorbed by the sand was much lower than that pred icted based on CEC. Also although the net Fe/Al oxides minerals are positively charged, they still could adsorb cations to form inner sphere complexes at low pH values ( Violante et al. 2003 ) As expected, preloading Cu and Ca onto sand surface increased its CIP sorption (Table 3 4) It is known that CIP forms Cu CIP and Ca CIP complex in solution ( Turel et al. 1996 ) The increased CIP sorption onto Cu loaded sand was through cation bridging effect either forming metal complex at strong affinity sites or via Columbia interactions at weak affinity sites. This is supported by Guaita et al ( 2011 ) who examined the impact of Cu on flumequine sorption (a fluoroquinolone antibiotic). They found Cu increases flumequine accumulation onto soil via formation of Cu flumequine ternary surface complex. A separate study by Pei et al. ( 2009 ) on CIP and Cu cosorption onto clays has similar result. They attributed the stronger affinity of CIP Cu co mplexe onto sand than CIP species to Cu bridging effect between the clay surface and CIP (Eqs.3 1 and 3 2)

PAGE 55

55 After being saturated with Cu 2+ SiO or Fe/AlO was most likely present as SiO Cu + or Fe/AlO Cu + (Eqs. 3 5 & 3 7). CIP was then sorbed onto C u preloaded sand (Table 3 2) via Cu 2+ bridging (Eqs. 3 6& 3 8). Clean sand SiO + Cu 2+ SiO Cu + ( 3 5) SiO Cu + + CIP SiO Cu CIP + ( 3 6) Native sand SiO + Fe/AlO + 2 Cu 2+ SiO Cu + + Fe/AlO Cu + ( 3 7) SiO Cu + + Fe/AlO Cu + + 2 CIP SiO Cu CIP + + Fe/AlO Cu CIP + ( 3 8) Since both clean and native sand were able to sorb Cu (1 and 8 ppm), preloading Cu increased CIP sorption by 0.58 and 2.11 mg kg 1 for clean and native sand respectively (Table 3 2) Assuming Cu:CIP ratio at 1:1, then based on the Cu concentrations preloaded on sand, the projected CIP concentrations were 5.2 and 41.8 mg kg 1 However, the actual sorbed CIP was much lower. It was possible only part of the Cu sorbed was suitable for CIP sorptio n. Based on the Cu concentrations in the solution after CIP sorption onto Cu preloaded sand, 0.15 and 0.8 mg kg 1 Cu was stripped off the clean and native sand surface (data not shown). It was possible that some Cu, which was sorbed onto sites with low af finity, was stripped off the sand surface after it complexed with CIP and came into solution. Only those Cu sorbed onto sites with high affinity stayed on sand surface. This results were in agreement of Pei et al ( 2009 ) who observed ligand promoted dissolution of soil surface cations in the present of CIP.

PAGE 56

56 Compared to Cu, preloaded Ca had different impact on CIP sorption. No impact was observed on clean sand as little Ca was sorbed onto clean sand; however preloading Ca onto native sand increased its CIP sorpti on by 2.6 mg kg 1 which was higher than that of Cu (2.11 mg kg 1 ). Though under detection, some Ca was sorbed onto native sand. The additional CIP was sorbed onto sand via Ca bridging similar to Cu. However, since CIP has weaker complexation with Ca tha n Cu, less CIP Ca was stripped off sand surface than CIP Cu (log k = 14.7 and 11.3; Table 3 1b). Premixing Cu and Ca with CIP Reduced CIP Sorption onto Clean Sand To further test the above hypothesis, excess amounts of Ca / Cu was mixed with CIP and then mixed with sand. With increasing cation concentrations, both Cu and Ca decreased CIP sorption onto clean sand (Figure 3 1 A), with Cu being more effective than Ca. Since excess amounts of Ca 2+ and Cu 2+ were mixed with CIP, CIP was mostly present as Ca CI P + and Cu CIP + in solution (Table 3 2; Eqs. 3 1& 2). Since the major reactive sites on sand were SiO the increasing presence of Ca / Cu in solution would mainly compete with Ca CIP + and Cu CIP + for the negatively charged SiO sites. With increasing Ca/C u concentrations (up to 30 mg L 1 ), Ca/Cu out competed Ca CIP + and Cu CIP + for the sorption onto SiO thereby decreasing CIP sorption onto sand (Table 3 2). However, even at 30 mg L 1 Ca, still some CIP remained on the sand. In comparison, at that rati o, Cu was able to strip almost all CIP off the sand (Figure 3 1 A), reflecting Cu as a transition metal of higher charge density was more effective in CIP desorption via complexation.

PAGE 57

57 Premixing CIP with Cu and Ca changed CIP sorption onto native sand Simila r to clean sand, with increasing Cu/Ca concentrations in solution, they outcompeted Ca CIP+ and Cu CIP+ for negative sorption sites on sand surface, reducing CIP sorption onto sand. However, in native sand, besides SiO sites, Fe/AlO sites were also present. Though in limited quantity, they were more effective in CIP sorption than 1; Table 3 2). So, with increasing Cu/Ca concentrations in solution, initially they increased CIP sorption which may result from cation bridging effect of CIP and CIP (Eq. 3 5 to 3 8) (Table 3 2). However, when cation concentration continued to increase, CIP sorption onto sand significantly decreased (Figure 3 1B). Similar to the effec t of Ca and Cu on CIP sorption onto clean sand the increasing cations in the solution probably stripped CIP complexes (Cu CIP+ or Ca CIP+) off the native sand surface as they both carried positive charge, leaving Cu or Ca on sand surface. Under the expe riment condition, the greater efficiency of Cu was due to its higher affinity to CIP and to sand surface than Ca (Table 3 1b) ( Wallis et al. 1996 ; Upadhyay et al. 2006 ) Since cations sorbed on native sand surface acted as a bridge and facilitated CIP sorption while cations in the solution phase competed with CIP for sorption onto sand. The ratio of cation conc entrations on the sand surface and solution should be directly related to CIP sorption. The Kd of CIP had linear correlation with the ratio of cation in the sand to that in solution (Cu at R2=0.9566, and Ca at R2=0.9055). This confirmed that cations in th e environment could either facilitate or impede CIP sorption depending on whether they are on solid or solution phase.

PAGE 58

58 Comparison of CIP Sorption with Probe Compound Flumequine In our experiment, the carboxylic group of CIP interacted with Cu/Ca and Fe/A l oxides on sand surface. It is possible that other fluoroquinolone antibiotics with similar functional group have similar behavior. I used FQ with fundamental fluoroquinolone antibiotics structure to further test this hypothesis. The molar ratio of Cu o r Ca to CIP or FQ in the solution was 100:1, so most of the CIP or FQ should be complexed with Cu or Ca and excess amount of Cu or Ca was present in the solution. In the absence of Cu or Ca, 1.58 mg kg 1 of CIP was sorbed onto clean sand. The presence o f Ca reduced CIP sorption by ~50% to 0.83 mg kg 1 whereas Cu completely inhibited CIP sorption onto clean sand. Both Ca and Cu competed with CIP for sorption onto SiO sites as electrostatic interaction was the main driven force for CIP sorption (Figure 3 2). Compared to CIP, FQ structure lacks positive charge so reduced amount of CIP was sorbed onto clean sand at ~11% or 0.20 mg kg 1 Compared to CIP, the reduced sorption of FQ onto sand was probably due to its lack of positive charge as clean sand sur face was negatively charged. Similar to CIP sorption, both Ca and Cu reduced FQ sorption with Cu completely inhibiting FQ sorption. In the absence of Cu or Ca, native sand was more effective than clean sand, sorbing 11.7 mg kg 1 CIP. However, the prese nce of Ca increased its sorption to 16.7 mg kg 1 whereas Cu reduced its sorption to 1.80 mg kg 1 Similar data were obtained for FQ, which indicate that its carboxyl group was mostly responsible for their sorption of CIP or FQ. The data also imply that n ative sand was more effective in sorbing CIP or FQ than CIP Cu or FQ Cu complexes. Since Ca has weak complexation ability with

PAGE 59

59 CIP or FQ, it helped their sorption via cation bridging ( Park et al. 2000 ) This is consistent with the study of enrofloxacin and other important quinolone antibiotic containing carboxylic groups ( Nowara et al. 1997 ) They also obse rved the high affinity between quinolone antibiotics and metal oxides, further demonstrating the importance of the carboxyl group in quinolone antibiotic in controlling its behaviors in the environment. Environmental I mplication Although CIP has been demon strated to have complexation ability with many cations, only the environmental influence of Cu was investigated in previous study ( Pei et al. 2009 ; Guaita et al. 2011 ) Also cation concentration as high as several hundred mg kg 1 has been frequently reported in surface soil ( Muchuweti et al. 2006 ) hence, the impact of cations at high concentrations should also be included to better understand the environmental be havior of CIP. In this study I demonstrated both Ca and Cu impacted CIP sorption onto sa nd. This indicated other metals with similar complexation ability with CIP should not be overlooked when present in the system. These cations sorbed on sand surface can decrease CIP mobility by increase CIP sorption while cations in solution can promote C IP mobility by reducing CIP sorption. In addition, due stripping CIP off sand surface. Recent study indicated after complexation with cations, CIP can be active over time and inhibit m icrobial activities in both solution and solid phase ( Girardi et al. 2011 ) For CIP sorption experiment us ing Ca to adjust ionic strength ( Zhang and Dong 2008 ; Wang et al. 2009 ) it may not reflect the CIP sorption beh avior at equilibrium as Ca can not only complex with CIP but also compete with CIP on sorption sites in soils.

PAGE 60

60 Conclusions I examined the effects of divalent metals Ca and Cu on CIP sorption on clean sand native sand. Our results indicated that: 1) Fe/Al o xides on native sand surface was responsible for CIP sorption; 2) Fe/Al oxides played an important role in cation bridging effect of Cu and Ca; 3) Both Cu and Ca decreased CIP sorption onto clean sand; and 4) Both Cu and Ca promoted CIP sorption onto nativ e sand, but at higher concentrations, they decreased CIP sorption. Our research demonstrated the importance of cations such as Cu and Ca in controlling the fate and transport of CIP in t he environment.

PAGE 61

61 Table 3 1. Basic properties of ciprofloxacin (CIP) Structure MW pKa Species at pH 5.6 331.4 pKa1=6.2 pKa2=8.8 Cation (80%) Zwitteron (20%)

PAGE 62

62 Table 3 2. Stability constant (log k) of CIP with Cu, Ca, Fe and Al ( 1996 ) Ca Cu Fe Al M(CIP) 11.3 14.7 M(CIP) 2 28.5 29.3 M(CIP) 3 46.9 43.3 Table 3 3 Basic property of sand Particle size (m) CEC (cmol kg 1 ) ZPC* Total Fe (mg kg 1 ) Total Al (mg kg 1 ) Clean Sand 4.5 5.5 0.05 5.0 Native Sand 4.5 5.5 0.28 5.1 167 1, 087 ZPC = zero point charge

PAGE 63

63 Table 3 4 The impact of Ca or Cu on CIP sorption onto clean and native sand (3 g sand with 1 mg L CIP for 24 h) CIP sorbed (mg kg 1 ) CIP in solution (g L 1 ) Potential sorption mechanisms Treatment Clean Native Clean N ative Clean Native DI water 3.050.10 7.000.17 695 300 SiO CIP Fe/AlO CIP SiO CIP Si/Fe/AlO Ca 2+ + CIP 2.980.06 9.600.82 702 40 SiO CIP SiO Ca CIP SiO CIP SiO Ca CIP Fe/AlO CIP Fe/AlO Ca/Cu CIP Si/Fe/AlO Cu 2+ + CIP 3.630.03 9.110.22 637 89 SiO CIP SiO Cu CIP SiO CIP SiO Cu CIP Fe/AlO CIP Fe/AlO Cu CIP Si/Fe/AlO Ca 2+ + Ca 2+ CIP 2.900.07 9.571.4 710 44 SiO CIP SiO Ca CIP SiO CIP SiO Ca CIP Fe/AlO CIP Fe/AlO Ca CIP Si/Fe/AlO Cu 2+ + Cu 2+ CIP 4.090.08 9.250.84 591 79 SiO CIP SiO Cu CIP SiO CIP SiO Cu CIP Fe/AlO CIP Fe/AlO Cu CIP

PAGE 64

64 Figure 3 1. CIP sorption coefficient Kd* by clean and native sand at 1 mg L CIP different cation concentrations. A) clean sand, and B) native sand *Kd = Cs/Cw; wher e Cs is solid phase CIP concentration; Cw is the solution phase CIP concentration. B A

PAGE 65

65 Figure 3 2. Sorption ability of sand of 0.01 mmol CIP or FQ in DI water, 1 mmol CaCl 2 or CuCl 2 solution A) clean sand, and B) native sand A) Clean sand B) Native sand

PAGE 66

66 Figure 3 3. Poss ible interactions in the system

PAGE 67

67 CHAPTER 4 INFLUENCE OF CU AND CA ON CIPROFLOXACIN TRANSPORT IN SATURATED POROUS MEDIA Introduction Ciprofloxacin (CIP), one of the most widely prescribed fluoroquinolones (FQs) antibiotics, is widely used in health care an d agricultural industries to treat and prevent human and animal diseases. It is designed to be refractory to biodegradation and to act effectively even at low doses ( Jones et al. 2001 ) so it is persistent in the environment and imposes a risk to the ecosystems. CIP has been frequently detected in streams and groundwater ( Kolpin et al. 2002 ; Barnes et al. 2004 ) with the concentrations ranging from ng L 1 to mg L 1 ( Larsson et al. 2007 ) CIP is known to be not readily biodegradable ( Girardi et al. 2011 ) and has high sorption affinity onto soils ( Nowara et al. 1997 ; Thiele Bruhn et al. 2004 ; Vasudevan et al. 2009 ) so up to ppm levels of CIP can be accumulated in soils. Therefore, the soil can act as a reservoir of CIP and oth er antibiotics ( Zorita et al. 2009 ) Hence, it is important to examine the fate and transport behaviors of CIP in soil and water system. The ability of FQs to for m complexes with various metal cations is important for their antibiotic activity ( Turel et al. 1996 ) so metal cations are expected to influence their fate and transport in the environment. The high affinity of soil m atrix for FQs has been attributed to their complexation with either exposed structural or exchangeable cations ( Drakopoulos and Ioannou 1997 ; Nowara et al. 1997 ; Trivedi et al. 2001 ; Gu and Karthikeyan 2005 ; Zhan g and Huang 2005 ) and the attraction of their protonized piperazine moiety (NH 2 + ) to the negatively charged mineral surface ( Goyne et al. 2005 ) Interaction between FQs and soil minerals are universal even silica shows substantal sorbtion ability for FQs ( Goyne et al. 2005 ) However, it should be noted that the

PAGE 68

68 inter action between FQs and metal cations are reversable and controled by the density of charge of metal cations ( Turel et al. 1996 ; Aristilde and Sposito 2008 ) The impacts of metal cations on CIP sorption onto minerals have been studied in bacth experiments ( 1992 ; Pei et al. 2009 ) however, little information is available on their impacts on CIP transport. Minerals like Fe/Al oxides can develop both positive and negative surface charge ( Huang et al. 1977 ; Violante et al. 2003 ) even present in limited quantities in soils, they are the major sorption sites for CIP. Hence, metals in both aqueous and mineral phase greatly influece the fate and transport of CIP in porous media and soils. This study was to provide insights into how metals influence CIP tansport in porous media to better understand the chemodynamic behavors of CIP in soils an d receiving waters, which is essential for comprehensive assessment of their potential environmental risk. I hypothesized that both minerals on sand surface and the persence of metal cations were critial factors controlling CIP tranpsort in saturated porou s media. To test the hypothesis, column experiments using saturated quartz sand were conducted to compare the transport behavior of CIP in two sands and in presence of two metals.The specific objectives were to: 1) evaluate the effects of metal oxides on s and surface on CIP transport in porous media, 2) compare the effects of metals Cu and Ca on retention and transport of CIP in porous media;3) compare the effects of metals Cu and Ca on mobilization of CIP presorbed on porous media; and 4) apply solute tran sport models to simulate the retention and tra nsport of CIP in porous media.

PAGE 69

69 Materials and Methods Materials Ciprofloxacin (ACS 85721 33 1) was purchased from Applichem (Darmstadt, Germany). Its chemical structure and basic information are presented in Ch apter 3 Quartz sand treatment and basic information was same as chapter 3 Column E xperiments See Chapter 2 Modeling Antibiotic Transport in Saturated Porous Media section Impacts of Aqueous Ca and Cu on CIP Transport in Sand M edia To test the impacts of aqueous Ca and Cu on CIP transport in sand media, 0.2 mg L 1 CIP (0.7 M) was mixed with DI water, 1 mM CaCl2, or 1 mM CuCl2. Once the outflow was stabilized, the experiment was then terminated. Effluent samples were collected from the top of the column with a fraction collector (IS 95 Interval Sampler, Spectrum Chromatography) during sample injection and column flushing to analyze CIP concentrations. All experiments were performed in duplicate. To determine the impacts of Ca and Cu on CIP sorption onto sand, separate batch study with same CIP concentration and Ca or Cu solution was conducted using 50 mL polytetrafluoroethylene centrifuge tubes. Each vessel was filled with 3.00 g of sand and 30 mL of 0.7 M CIP with DI water, 1 mM CaCl2 or 1mM CuCl2. The vessels were shaken for 24 h. Preliminary experiment showe d that it reached equilibrium within 24 h (data not shown). The suspensions were centrifuged at 5,000 g for 10 min to separate the solids from the liquid phase. Aliquots of supernatant were withdrawn to determine CIP concentrations, solid phase concentrat ions were obtained through mass balance calculations. The CIP concentration was determined by HPLC (Waters 2695, Milford, MA) equipped with a fluorescence detector (Waters 2475, Milford, MA). The

PAGE 70

70 sorbed Ca or Cu on the sand was analyzed by ICP MS (NexION 3 00, Perkin Elmer Crop, MA) after digesting with HNO3/H2O2 hot block digestion procedure ( EPA ) The detection limits for Ca and Cu were 0.5 g kg 1 a nd 0.05 g Kg 1, respectively Batch experiments were performed in triplicate. Impacts of A queous Ca and Cu on M obilization of CIP P resorbed onto S and M edia In the previous experiment, CIP and Cu or Ca was transported simultaneously in the sand column. T o determine the ability of aqueous Ca and Cu in mobilizing CIP in sand media, CIP was applied to the column at a 5 PV pulse, and then the column was flushed with DI water, 1 mM CaCl 2 or 1 mM CuCl 2 Experiment was terminated when out flow stabilized or aft er 100 PV. Effluent samples were collected from the top of the column with a fraction collector during flushing with Ca or Cu solution after CIP injection pulse to analyze CIP concentrations. All experiments were performed in duplicate. Since Ca/Cu was in effective in mobilizing CIP presorbed onto sand column, a separate column study was conducted. For each set of experiment, CIP was applied to the columns at a 5 PV pulse, and then the columns was flushed 5 PV of DI water, 1 mM CaCl 2 or 1 mM CuCl 2 respect ively. At the end of flushing, the column was separated into 10 sections to determine the concentrations of retained CIP as a function of column depth. The sand was excavated under saturated conditions from top to bottom with a spatula in 1 cm increment an d each increment was placed into a small vial with 4 mL of H 3 PO 4 KH 2 PO 4 buffer and acetonitrile solution (i.e., 27.2 g KH 2 PO 4 + 1.35 ml H 3 PO 4 in 1 liter water with 1:1 acetonitrile ) ( Uslu et al. 2008 ) The excavated sand was then washed three times with the same amount of extraction solution. CIP conc entrations in

PAGE 71

71 the solutions were determined with HPLC. The CIP retention was then calculated for each sand section. Modeling CIP Transport in Saturated Porous Media One dimensional advection dispersion equation coupled with reaction terms was used to simu late the transport. See Chapter 2 modeling section. Results and Discussion CIP Species in Solution and Characteristics of Sand interactions with Cu, Ca and sand. CIP can exist in three forms (cationic, zwitterionic, and anionic forms) in DI water with two active function groups including carboxyl and piperazinyl groups. At pH 5.6, 100% of piperazinyl groups existed as NH 2 + and 20% of carboxyl groups presented as COO ( Chapter 3, Table 3 1). In comparison, >99% Cu and Ca were present as free cations based on Visual MINTEQ (data not shown). When coexisting with metal cations in solution, CIP can complex with Cu or Ca via its carboxyl and keto groups to form metal CIP or metal CI P 2 complexes. In this study, since the amounts of Cu or Ca (1 mM) used were much greater than CIP (0.7 M), only 1:1 metal CIP was important ( Li et al. 1994 ; Turel et al. 1996 ) It should be noted that CIP cation compl exation reactions in solution were reversible, especially when cations with low complexation ability (e.g., Ca) compete with cations with strong ability (e.g., Fe/Al) ( Turel et al. 1996 ; Aristilde and Sposito 2008 ) In this experiment, the net surface charge for clean sand at pH 5.6 was negative with its PZC (point of zero charge) being 5.0 5.1 ( Chapter 3, Table 3 3 ), with the majority of negative sites being from broken edges of silicon oxides ( SiO ), which sorb

PAGE 72

72 CIP via electrostatic attraction of its piperazinyl groups. Different from clean sand, native sand was coated with small amount of Fe/Al oxides (167 a nd 1,087 mg/kg Fe and Al) ( See Chapter 3 Table 3 3 ). The PZC for Fe/Al oxides is 10 ( Essington 2003 ) so locally the broken edges of Fe/Al oxides carried positi ve charges ( FeOH + and AlOH + ). The broken edges of Fe/Al oxides with high energy are highly reactive and can complex with CIP, forming surface complexation ( Yost et a l. 1990 ; Molis et al. 2000 ; Duckworth and Martin 2001 ) (Figure 4 1C). In addition, Fe/Al oxides can also sorb Cu/Ca ( Fe/AlO -Ca 2+ or Fe/AlO --Cu 2 + ) (Figure 4 1C). Trivedi et al. ( 2001 ) demonstrated that Cu as a transition metal is more likely to bind to high affinity site on goethite whereas Ca as an alkaline earth metal bind to low affinity site ( Trivedi et al. 2001 ; Violante et al. 2003 ) CIP Transport in Saturated Sand Porous M edia The breakthrough curve (BTC) of CIP transport in clean sand column in DI water (i .e., pH = 5.6 and IS = 0 mM) was significantly delayed compared to that of bromide, the conservative tracer (Figure 4 2A). After applied to the clean sand column, CIP was first detected in the effluents ~30 PV. The BTC then slowly climbed to a peak value a t ~60 PV and stayed there during further CIP injection (Figure 4 2B). Compared to the bromide BTC in the column, the delayed CIP transport indicated significant CIP retardation in sand media, which was consistent with the result in Chapter 2 The normalize d peak concentration (C/C0) of CIP was ~70%. The breakthrough experiment demonstrated that CIP interacted with the clean sand. This interaction was attributed to the electrostatic attraction between positively charged CIP and negatively charged clean sand ( SiO --CIP+) ( Taboada Serrano et al. 2005 ; Huang et al. 2012 ) (Figure 4 1B). At pH of 5.6, CIP carri ed both positively charged piperazine group (100% NH2+,

PAGE 73

73 Table 3 1) and negatively charged carboxyl group (20% COO Table 3 1). Our data were consistent with the literature where many studies have observed the low mobility of CIP in soils ( Stoob et al. 2007 ; Unold et al. 2009 ) Simulations of the transport model matched well with the BTC of CIP transport in clean sand with R2 =0.99 (Figure 4 1). The best fit retardation factor R of CIP in the clean sand column was 21.7 (Table 4 1 ). The best fit kinetic reaction rate k of CIP was 0.018 min 1, suggesting part of the CIP was retained in the sand column through kineti c reactions. Different from CIP transport in clean sand, its transport in native sand column showed no breakthrough response under DI water condition (data not shown). No CIP was detected in the effluents even after prolonged injection as much as 100 PV. T his indicated strong retardation/deposition of CIP onto native sand. In native sand, in addition to surface charge from clean sand, it was also coated with Fe/Al oxides. Besides the electrostatic attraction between CIP and sand, the retention of CIP onto t he sand column was primarily attributed to strong complexation between CIP and Fe/Al oxides ( Riley et al. 1993 ; Tu rel et al. 1994 ) ( Chapter 3 Figure 3 3 ). The log k of complexation constant of Fe(CIP)3 and Al(CIP)3 was 46.9 and 43.6 ( Turel et al. 1996 ) The fact that no CIP came out from native sand indicated that the complexa tion of CIP with Fe/Al on the sand surface was much stronger than sorption of CIP by SiO The strong CIP retention in native sand column was consistent with the literature examining its sorption behaviors in soils ( Nowara et al. 1997 ; Gu and Karthikeyan 2005 ; Carrasquillo et al. 2008 ; MacKay and Seremet 2008 ) As discussed above, CIP sorption onto sand in DI water systems was mainly controlled by two mechanisms: 1) weak

PAGE 74

74 electrostatic attraction between CIP functional groups and heterogeneously charged sand surfa ces ( Hari et al. 2005 ) and 2) strong complexation between CIP carboxyl groups and Fe/Al oxides on sand surface ( Nowara et al. 1997 ) (Figure 4 1C). These two mechanisms were confirmed in the literature through batch and spectroscopic studies of CIP sorption onto Fe/Al oxides and aluminosilicates ( Nowara e t al. 1997 ; Goyne et al. 2005 ; Gu and Karthikeyan 2005 ; Oker and Akmehmet BalcIoglu 2005 ) Presence of Cu an d Ca promoted CIP transport in Clean S and When CIP was mixed with Cu 2+ and Ca 2+ it existed primarily as Cu CIP + and Ca CIP + with little free CIP in solution (Figure 4 1A). In addition, excess amounts of Cu 2+ and Ca 2+ in comparison to Cu CIP + and Ca CI P + were also present in the solution. Compared to CIP alone, presence of Cu or Ca significantly promoted CIP transport in clean sand (Figure 4 3). The BTC of CIP in the presence of Ca or Cu was significantly advanced. The BTC of CIP transport in the clean sand column in presence of Cu (i.e., pH = 5.6, CIP = 0.7 M and Cu = 1 mM) was only slightly retarded than the BTC of bromide, the conservative tracer (Figure 4 2A). A fter applied to the sand column, CIP was detected in the effluents ~1.5 PV. The BTC then quickly climbed to a peak value at ~4 PV and stayed there during further injection (Figure 4 3A). Compared to the CIP transport in clean sand under DI water condition (30 PV), the BTC of CIP in the presence of Cu showed significant advance, indicating muc h reduced attraction between CIP and sand surface. Similar to Cu, the presence of Ca also significantly increased CIP transport in clean sand (Figure 4 3B). CIP was detected in the effluents ~1.5 PV, then quickly climbed and reached peak value at ~5 PV. T he normalized peak concentration (C/C0) of CIP was ~ 0.7 (Figure 4 3B), which was 25% less than that in presence of Cu (C/C0 ~ 0.95) (Figure 4 3A), indicating that Ca was less effective in

PAGE 75

75 mobilizing CIP from clean sand than Cu. This was consistent with t he fact that clean sand had greater sorption capacity of Cu than Ca (2.99 vs. 0.03 mg/kg; Chapter 3 Table 3 3). At pH 5.6, 100% of piperazinyl groups of CIP existed as NH 2 + and 20% of carboxyl groups presented as COO ( Chapter 3 Table 3 1). When CIP was tr ansported in clean sand, it was effectively sorbed onto negatively charged sand surface ( SiO ; Figure 4 1B). However, when CIP was mixed with Cu or Ca, the sand was ineffective in sorbing CIP as it existed as Cu CIP + or Ca CIP + This was probably becaus e aqueous Cu 2+ and Ca 2+ outcompeted Cu CIP + or Ca CIP + to be sorbed onto sand surface, lea ving more CIP in the solution. This hypothesis was supported by batch experiment where 1 mM of Cu or Ca was mixed with 0.7 M CIP (Table 4 2 ) The ability of clean sa nd in sorbing CIP was reduced from 0.5 to 0.03 mg kg 1 in the presence of Cu and from 0.5 to 0.38 mg kg 1 in the presence of Ca (Table 4 2 ). In this aspect, clean sand was more effective in sorbing Cu than Ca, with CIP being the least effectiv e among the three (Figure 4 1B). Though Cu/Ca effectively increased CIP transport, the BTC of CIP in the presence of either Ca or Cu showed some delay, indicating slight retardation of CIP in the porous media. This was probably due to the attraction of CuCIP + or CaCIP + onto negatively charged sand surface. Simulations of the transport model matched well with the BTCs of CIP in the presence of Ca or Cu, with R 2 of 0.93 and 0.97, respectively (Table 4 1 ). The best fit retardation factor R of CIP in the column was 1.71 (C a) and 1.17 (Cu) The best fit kinetic reaction rate k of CIP in the column was 0.016 min 1 (Ca) and 0.0031 min 1 (Cu)

PAGE 76

76 compared to bromide tracer with 0 min 1 suggesting some of the CIP was retained in the sand column through kinetic reactions with more in the presence of Cu than Ca. The modeling results further confirmed that Cu was more effective than Ca in increasing CIP mobility in clean sand. Presence of Cu P romoted CIP Transport in Native S and In the presence of 1 mM Cu, CIP was detected in the effl uents ~30 PV after applied to the native sand column. The BTC then slowly climbed to a peak value and stayed there during further CIP injection. The normalized peak concentration (C/C0) of CIP was ~ 1 after 60 PV compared to ~0.7 after 5 PV in clean sand. Clearly, compared to clean sand, it was much easier for Cu CIP+ to be sorbed onto native sand. In native sand, besides quartz sand, Fe/Al oxides were also present. With stronger complexation of CIP to Fe/Al than Cu ( Gu and Karthikeyan 2005 ) Fe/Al oxide s on native sand surface enhanced CIP sorption onto sand (Figure 4 1C). Since native sand had stronger sorption capacity for CIP than clean sand (1.4 vs. 0.5 mg/kg; Table 4 3), it took much longer for CIP to come out from clean sand tha n native sand (60 v s. 5 PV). Unlike Cu, when Ca was present in the solution, CIP could not be detected even at 100 PV, indicating much weaker complexation of CIP with Ca. The log k of Ca CIP+ and Cu CIP+ complexation constant was 11.3 and 14.7 ( Turel et al. 1996 ) so Cu CIP+ was 3 order magnitude stronger than Ca CIP+. As a result, it was easier for CIP to disassociate from Ca CIP+, leaving CIP to be sorbed onto sand via strong complexation with Fe/Al on the sand surface. This was supp orted by the batch data where 1.7 and 1.0 mg/kg CIP was sorbed onto native sand in the presence of Ca and Cu (Table 4 2 ). Our result indicated that the presence of metals in aqueous phase and surface

PAGE 77

77 chemistry of porous media were both important in contro lling CIP retention and transport in sand media. Aqueous Cu and Ca Increased Transport of Presorbed CIP in Clean S and In the previous experiment, CIP was mixed with Cu or Ca. In this experiment, CIP was presorbed onto sand and then leached with aqueous Cu or Ca. Both Cu and Ca mobilized CIP presorbed on sand surface, with Cu being more effective than Ca (Figure 4 4). After Cu was applied to sand column presorbed with CIP, CIP was detected around 1.5 PV, then quickly climbed to peak around 2 PV. The normal ized peak concentration (C/C 0 ) of CIP was ~ 1.8. This number was greater than unity because presorbed CIP onto sand column was released in short period when Cu was eluted in the column. Mass balance calculations showed that 90% of CIP was remobilized from sand column. The results demonstrated that Cu was effective in mobilizing CIP presorbed in clean sand. This could be attributed to the fact that the interaction between CIP and clean sand was mostly via electrostatic attraction. When significant amount Cu (1 mM) solution flushed the column, Cu 2+ effectively replaced the CIP + on the sand surface since they were both positively charged (Figure 4 1B). This was similar to the experiment where Cu 2+ was mixed with CIP, with Cu being more effective being sorbed onto clean sand surface than Cu CIP + Similar to Cu, after Ca was applied to sand column with presorbed CIP, CIP was detected ~1.5 PV, then quickly climbed to peak ~3 PV. The normalized peak concentration (C/C 0 ) of CIP was ~ 0.25. Mass balance calculations showed that 30% of CIP was remobilized from sand column compared to 90% by Cu, reflecting Ca was less effective in desorbing CIP from sand surface. This was consistent with the fact that clean sand was more effective i n sorbing Cu than Ca (Table 4 2 ).

PAGE 78

78 Aqu eous C nhance T ransport of P resorbed CIP in N ative S and Since no CIP was detected after 100 PV of DI water, 1 mM Ca or 1 mM Cu, a separate profile study was conducted. Analysis of the retained CIP profile in the sand column after 5 PV of DI water flushing showed that almost all (~95%) CIP was retained in the bottom 1 cm layer of the column and the rest of 5% of CIP being in the second 1 cm layer from the bottom (Figure 4 5), suggesting strong interactions between CIP and the native sand grai ns (Figu re 4 1C). Similar to DI water, analysis of the retained CIP profile in the sand column after flushing with Ca solution showed that 94% the CIP was retained in the bottom 1 cm layer and the rest of 6% of CIP being found in the second 1 cm layer from the bottom (Figure 4 5). This suggested the limited effect of Ca on mobilizing the CIP presorbed on the native sand surface. Unlike Ca, analysis of the retained CIP profile in the sand column after 5 PV of 1 mM Cu solution flushing showed that 88%, 10%, a nd 2% CIP was retained in the first, second, third 1 cm layer from the b ottom, respectively (Figure 4 5). The results were consistent with the hypothesis that sorption of CIP onto native sand via complexation with Fe/Al was stronger and more effective tha n that via electron static attraction onto SiO Once sorbed onto Fe/Al surface, Ca or Cu was probably unable to replace CIP. However, CIP sorbed onto SiO surface can be replaced by Ca (~1%) or Cu (~7%). So more likely, those mobilized CIP was from SiO surface. Though limited in quantity, Fe/Al was more effective sorbing CIP than SiO as supported by the batch data where 1.4 mg/kg CIP was sorbed by native sand compared to 0.5 mg/kg by clean sand (Table 4 2 ).

PAGE 79

79 Conclusions CIP can interact with meta ls in aqueous and solid phase, which has not been studied in details. This study showed that while surface metal oxides was effective in immobilizing CIP in soils, solution phase Ca and Cu facilitated CIP transport. Critical factors in facilitating or im peteing CIP transport included: 1) the type of metal cation, which controls the stability of the complex formed; 2) the concentration of metals present; and 3) wether the cations are in aquous or solid phase. Although only aqueous Cu and Ca were investigat ed in present study, CIP could form complexes with majority of metal cations in soils ( Ture l 2002 ) suggesting other metals could have simliar effect on CIP transport. In addition, almost all FQs share the same carboxyl, keto and piperazine reactive functional group like CIP, similar chemodynamics process was expectd for other FQs. Therefore, other major cations in aqueous and soil environment should also be evaluated to better understand the chemodynamics of FQs in soils and receiving waters, which is critial for comprehensive assessment of the potential environmental risk of FQs.

PAGE 80

80 Table 4 1 Transport model parameters and experiment conditions for CIP transport in sand columns. No. Sand Influent R k R 2 1 Clean DI water 21.7 0.018 0.99 2 Native DI water ---------3 Clean 1 mM Ca 1.71 0.016 0.93 4 Native 1 mM Ca ----------5 C lean 1 mM Cu 1.17 0.0031 0.97 6 Native 1 mM Cu 22.0 0.0054 0.99 The influent consisted of 0.7 M CIP with DI water, 1 mM Ca or 1 mM Cu. When Fe/Al oxides from native sand were removed, they were referred to as clean sand R is the retardation factor, whi ch reflects the magnitude of equilibrium reactions in the sand column; k is the kinetic reaction rate constant (min 1 ) and D is the dispersion coefficient (cm 2 min 1 ). Table 4 2 Partition coefficient of CIP onto clean and native sand after shaki ng 3 g o f sand with 0.7 M CIP plus water, 1 mM Ca or 1 mM Cu for 24 h. Cu/Ca sorbed on sand (mg kg 1 ) CIP sorbed on sand (mg kg 1 ) Treatment Clean Native Clean Native DI water under detection limit under detection limit 0.50.06 1.40.01 1 mM Cu 2.990.02 210 .40 0.030.00 1.00.01 1 mM Ca 0.030.03 0.210.06 0.380.09 1.70.11

PAGE 81

81 A) CIP forms stronger complex with Cu than Ca in solution B) Electrostatic attraction of CIP, Ca and Cu onto clean sand surface C) Complex of CIP with Fe/Al and electrostatic attraction of Ca and Cu onto Fe/Al oxides on native sand surface Figure 4 1. Possible interactions of CIP with clean and native sand.

PAGE 82

82 Figure 4 2. Transport of bromi de in native and clean sand and CIP in clean sand media under DI water condition. When Fe/Al oxides from native sand were removed, they were referred to as clean sand A= Br B= CIP

PAGE 83

83 Figure 4 3. CIP transport in saturate d sand me dia where it was mixed with:1 mM Cu in clean sand, 1 mM Ca in clean sand and Cu in native sand. When Fe/Al oxides from native sand were removed, they were referred to as clean sand. C = Cu / native sand Cu/native sand A= Cu/clean sand B=Ca/clean sand

PAGE 84

84 Figure 4 4. Mobilization of presorbed CIP onto sand after co lumn was flushed with DI water, 1 mM Ca, and 1 mM Cu for 5 pore volumes

PAGE 85

85 Figure 4 5. Distributio n of sorbed CIP in different layers of the saturated native sand column after flushed with DI water, 1 mM Ca or 1 mM Cu

PAGE 86

86 CHAPTER 5 COLLOID FACILITATED CIP TR ANSPORT IN SATURATED POROUS MEDIA Introduction Small particles in subsurface have been characterized as mobile sorbent can facilitate containments transport. Colloids, defined as soil particles with diameters less than 10 micrometers ( Essington 2003 ) are widely recognized as mobile particles in the subsurface. These colloids have gain growing concern due to their high affinity to contaminants and risks to human health ( Roy and Dzombak 1997 ) Many organic contaminants, generally considered to be highly retarded due to strong interactions with immobile aquifer material, also have a high affinity for the mobile colloidal material. Consequently, the association between the contaminant and colloid ultimately affects the tran sport of the contaminant ( Roy and Dzombak 1997 ) .The mobili ty potential predicted by batch adsorption experiment maybe misleading in consider of the presence of colloids as mobile sorbents ( Grolimund and Borkovec 2005 ) The significance of colloidal facilitated transport depends on the follow ing factors: (a) the identity and concentration of colloids; (b) the nature of the interaction between the contaminants and the colloids; and (c) the mobility of the colloids in an aquifer. Although it is known that mobile colloids can enhance the mobility of contaminants, there are not many experimental studies focusing on colloid facilitated contaminant transport in water saturated porous media Sen and Khilar ( 2006 ) After entering water satura ted media, most of the chemical contaminants and natural colloids interact with each other and with the surrounding media. These interactions may alter the dynamic behavior of contaminants, and increase the difficulty to understand the fate and transport o f contaminants in the media.

PAGE 87

87 Inorganic colloids include clay, metal oxides, and inorganic precipitates in the sub micrometer size range ( Essington 2003 ) These colloids occur both naturally and from anthropogenic sources. These colloids were detected in the ground water down gradient from the disposal site indicating that the colloids were mobile in the aquifer system. Ciprofloxacin (CIP) is one of the most widel y prescribed fluoquinolon antibiotics, also the main metabolite of enrofloxacin. The sources of CIP in the environment include land application of sewage sludge, wastewater irrigation, and disposing of expired pharmaceutical prescriptions (Golet et al. 200 2), making CIP of increasingly environmental concern. With highly reactive functional group, CIP has been demonstrated can highly associated with certain natural colloids like Fe/Mn oxides ( Gu and Karthikeyan 2005 ) geophite and montmorillonite ( Carrasquillo et al. 2008 ) In particular, no research has yet been published on the influence of natural clay on CIP mobility in saturate porous media. Little information is currently available concerning the association between inor ganic colloids and organic contaminants. Due to both the existing data base concerning organic contaminant sorption to typical subsurface mineral surfaces and to inorganic colloidal transport, the scientific framework suggests facilitated transport is a vi able transport mechanism. It is therefore crucial to examine the enhanced CIP transport in water saturated porous media. This study was designed to improve the current understanding of colloid facilitated organic contaminant transport in saturated porous media. Clay colloids and CIP were used in laboratory experiments to examine their transport and cotransport behaviors in water saturated sand columns. Our objectives were 1) to compare the

PAGE 88

88 transport of kaolinite, montmorillonite and CIP in water saturated porous media; 2) to examine the colloid facilitated CIP transport in saturated porous media; 3) to examine the effects of kaolinite and montmorillonite on the mobility of CIP in saturated porous media. Materials and Methods Materials Ciprofloxacin (ACS 857 21 33 1) was purchased from Applichem (Darmstadt, Germany). Its chemical structure and basic information are presented in Chapter 3. Quartz sand treatment and basic information was same as chapter 3. Kaolinite and montmorillonite powders (EM Science, Gibb stown, N.J.) were used to make colloids according to the procedures reported by Gao et al. 2004. The mean sizes of the colloids, as determined by photon correlation spectroscopy, did not vary significantly during the experiments and equaled 0.80 m for kao linite and 0.65 m for montmorillonite Batch Adsorption Experiments Adsorption experiments of CIP to kaolinite and montmorillonite were conducted using polytetrafluoroethylene centrifuge tubes as completely mixed batch vessels. Each vessel was filled wit h a predetermined amount of sorbent about 500 ppm kaolinite or 200 montmorillonite and 50 mL target CIP solutions 0, 0.5, 1, 1.5, 2, 5, 10ppm. The vessels were shaken for more than 24 h (predetermined) to reach apparent sorption equilibrium, and then centr ifuged to separate the solids from the liquid phase. Aliquots of supernatant were withdrawn to determine liquid phase CIP concentrations using HPLC (Waters 2695, Milford, MA) equipped with a fluorescence detector (Waters 2475, Milford, MA).

PAGE 89

89 Column Experime nts Column packing method was same as Chapter 2 ( Modeling Antibiotic Transport in Saturated Porous Media section ) Bromide was applied to the column as a conservative tracer for the breakthrough studies, see chapter 2, Column Experiments section. CIP Collo id Co transport in Sand Media Four types of colloid suspensions were used: pure kaolinite, kaolinite CIP suspension, pure montmorillonite and montmorillitnite CIP suspension. The CIP colloid suspension was applied to the column as a 2 pore volume pulse, a nd then the column was flushed with DI water for 4 pore volume. Colloid concentrations were determined by measuring the total extinction of light with UV visible spectrophotometry wavelength 350 nm for Kaolinite and 275 nm for montmorillitnite The CIP conc entration was determined by HPLC. The experiments were conducted in duplicates. Colloids Mobilization of CIP Presorbed onto Sand Media In the previous experiment, CIP and colloids were transported simultaneously in the sand column. To determine the abilit y of colloids in mobilizing CIP presorbed in sand media, CIP was applied to the column at a 2 PV pulse, and then the column was flushed with 500 ppm Kaolinite or 200 ppm montmorillitnite Experiment was terminated when out flow stabilized or after 6 PV. Ef fluent samples were collected from the top of the column with a fraction collector during flushing with kaolinite or montmorillitnite solution after CIP injection pulse to analyze CIP concentrations. All experiments were performed in duplicate.

PAGE 90

90 Results an d Discussion CIP Sorption Isotherm onto Sand, Kaolinite and Montmorillonite Both Kaolinite and montmorillonite were able to sorb CIP, but their ability differed substantially. The sorption of CIP onto native sand was significantly higher than clean sand. C IP sorption data fit well wit h Langmuir equation (R 2 = 0.998 0.999 ; data not show ). The best fit value of maximum CIP sorption capacity from Langmuir model were 60mg kg 1 for Kaolinite and 30 g kg 1 for montmorillonite, CIP sorbed to montmorillonite signif icantly higher than Kaolinite Montmorillonite possesses 95% permanent charge, 5% pH dependent charge, and a higher CEC, than kaolinite, which possesses 95% pH dependent charge, 5% permanent charge, and lower CEC ( Essington 2003 ) Previous research suggested the CEC is one of the key factors for CIP sorption to soil minerals ( Carrasquillo et al. 2008 ; Vasudevan et al. 2009 ) At pH of 5.6, amine groups ( NH 2 + ) of CIP was ~10 0% positively charged hence, high CIP sorption onto montmorillonite can be attributed t o columbic attraction between positively charged CIP amine groups ( NH 2 + ) and negatively charged site in mont morillonite. To better understand the mechanism of CIP sorption to montmorillonite, one experiment was conducted to exam the sodium released form m ontmorillonite while CIP was sorbed. The result showed sodium released is directly linear proportional to CIP sorbed to montmorillonite further confirmed cation exchanged happened when CIP sorbed. Although lack of CEC (0.28 cmol kg 1 ) native sand also sho wed relatively high sorption ability to CIP. This was mainly attributed t o Fe/Al oxides on the surface ( Fe/AlO + ), which amounted to 167 and 1,087 mg kg 1 (Table 3 4, chapter 3). Previous study indicate d surface complexation of Fe/Al oxides with CIP was wit h its carboxylic

PAGE 91

91 acid group (Gu and Karthikeyan 2005). The method using both native sand and clean sand (chapter3) also confirmed the surface complexation process happened on native sand surface. CIP, Kaolinite and Montmorillonite Transport in Saturated Sa nd Column Transport of kaolinite and montmorillonite was performed in saturated sand column. Based on the breakthrough curve of montmorillonite and kaolinite were both same to that of conservative tracer bromide, indicating that no interaction between thes e two colloids and sand media (Fig. 5 1). This could be attributed to the fact that the interaction between colloids and porous medium were repulsive because both surfaces were negatively charged under experimental conditions. The zeta potential of the qu artz sand used in this experiment was 19.7 mV, confirming it was negatively charged. Different from colloids, CIP show no breakthrough curve, see chapter 2, CIP transport part. CIP Colloid Co transport in Sand Media Kaolinite has similar breakthrough cu rve like bromide tracer, CIP have no effect on k aolinite transport ( Fig. 5 3 ) However, there was no break th ough for CIP co transport with k aol inite indicating much weaker complexation of CIP with kaolinite. As a result, it was easier for CIP to disassoc iate from CIP Kaolinite complexion, leaving CIP to be sorbed onto sand via strong complexation with Fe/Al on the sand surface. In the presence of montmorillonite, CIP was detected in the effluents about 1 PV after applied to the native sand column. The BT C then quickly climbed to a peak value and stayed there during further CIP montmorillonite injection ( Fig. 5 2 ) The CIP concentrations decreased quickly to zero when the columns were flushed with CIP free

PAGE 92

92 solution. The normalized peak concentration (C/C0) of CIP was ~ 1. Exactly same breakthrough curve was detected for montmorillonite, the correlation between CIP and montmorillonite concentrations revealed a strong linear relationship with R 2 of 0.99(data not show). The linear relationship can be used to si mplify the models by assuming that the CIP and montmorilinite are always directly proportiona l during the transport process. The high ability of montmorillonite as CIP carrior could be explained by the high sorption ability to CIP of montmorillonite. Also previous research demonstrated CIP could be sorbed to montmorillonite interelayer ( Wang et al. 2011 ) The interaction between sand and CIP sorbed between the montm orillonte layers can be limited since directly. Transport of Presorbed CIP in Native Sand under the Influence of Colloids In the previous experiment, CIP was mixed with colloid. In this experiment, CIP was presorbed onto sand and then leached with colloid s suspension. After CIP presorbed into sand, both Kaolinite and montmorillonite have similar breakthrough curve to colloid transport in DI water. That indicate although small amount CIP was adsorbed to the sand may alert the surface character of sand, it is not significant enough to change the interaction between colloids and sand. Kaolinite could not mobilize CIP presorbed on sand surface, no CIP could be detected in the effluence while flushing column with kaolinite suspension (Figure 5 4) However, mont morillonite was much more effective to mobilize CIP. After montmorillonite was applied to sand column presorbed with CIP, CIP was detected around 0.8 PV, then quickly climbed to peak around 2 PV. The normalized peak concentration (C/C0) of CIP was ~ 1.6 (Fi gure 5 5) This number was greater than unity because presorbed CIP onto sand column was released in short period when montmorillonite was eluted in the column. Mass balance calculations showed that 99%

PAGE 93

93 of CIP was remobilized from sand column. The results demonstrated that montmorillonite was effective in mobilizing CIP presorbed in sand. The higher affinity of montmorillonite for CIP over sand resulted in CIP to be stripped from the sand and sorbed onto the montmorillonite. When montmorillonite solution f lushed the column, CIP was detached to sand surface and relocated to montmorillonite particles. The much higher mobilization ability of montmorillonite to CIP presorbed in sand column, indicating the sorption ability of colloids was the key factor to cont rol the mobilization capacity. Environmental I mplication In this experiment only montmorillonite was tested, other colloids like goethite also has been reported have high sorption ability, as a result they could also be effective to facilitate CIP transpor t. Besides, soil minerals like iron or aluminum oxides have also been demonstrated to have strong sorption to CIP ( Gu and Karthikeyan 2005 ) they could also perform as CIP carrior for CIP transport. Therefore the conclusion that CIP is low mobility chemical in soil system based on the strong CIP sorption to soil minerals is misleading. In consider of high concentration of CIP in water body has been reported colloid facil itated CIP transport should not be overlooked. Conclusions I examined the effects of colloid kaolinite and montmorillonite on C IP transport in saturate porous media Our results indicated that: 1) montmorillonite has higher sorption ability to CIP than Ka olinite; 2) premixed montmorillonite can facilitate CIP transport, Premixed Kaolinite could not facilitated CIP transport; 3)montmorillonite can remobilize presorbed CIP, kaolinite cannot remobilize presorbed CIP. Our research demonstrated the importance of colloids in controlling the fate and transport of CIP in the environment.

PAGE 94

94 Figure 5 1 Transport of kaolinite and montmorillonite in sand media under DI water.

PAGE 95

95 Figure 5 2 Premixed CIP and kaolinite co transport in saturated sand media

PAGE 96

96 Fig ure 5 3 Premixed CIP and montmorillonite co transport in saturated sand media

PAGE 97

97 Figure 5 4. Mobilization of presorbed CIP onto sand after column was flushed with montmorillonite.

PAGE 98

98 Figure 5 5 Mobilization o f presorbed CIP onto sand after column wa s flushed with Kaolinite.

PAGE 99

99 CHAPTER 6 CONCLUSIONS AND FUTU RE DIRECTIONS Laboratory batch and column experiments were conducted to examine the effects of solution chemistry (i.e., ionic strength(IS) and pH) on the retention and transport of two antibiotics i n saturated porous media. Our results indicated that: 1) s ulfamethoxazole ( SMZ ) was more mobile in saturated porous media than c iprofloxacin ( CIP ) ; 2) Solution pH played an important role in controlling the transport of CIP, but showed little effect on the transport of SMZ under the experimental conditions tested; 3) Solution IS had little effects on SMZ transport under the two pH conditions (9.5 and 5.6) but slightly enhanced the mobility of the retained CIP in the sand column at pH 5.6; and 4) traditional solute transport model could be used to simulate the retention and transport of antibiotics in water saturated porous media. In addition to solution IS and pH, CIP can interact with metals in aqueous and solid phase, which has not been studied in details. Our study showed that while surface metal oxides was effective in immobilizing CIP in soils, solution phase Ca and Cu increase the mobility of CIP Our results indicated that: 1) Fe/Al oxides on native sand surface was responsible for CIP sorption; 2) F e/Al oxides played an important role in cation bridging effect of Cu and Ca; 3) b oth Cu and Ca decreased CIP sorption onto clean sand; and 4) b oth Cu and Ca promoted CIP sorption onto native sand, but at higher concentrations, they decreased CIP sorption. Our research demonstrated the importance of cations such as Cu and Ca in controlling the fate and transport of CIP in the environment. Column experiements were designed, in consider of t he primary disadvantage of the batch technique for measuring retardati on factor is that it does not necessarily

PAGE 100

100 reproduce the chemical reaction conditions that take place in the real environment. T o better understand the chemodynamic behavors of CIP in soils and receiving waters column experiment was designed to investigate c ritical factors in facilitating CIP transport or remobilize CIP presorbed in the sand column. These critical factors included: 1) cation with stronger complxation ability with CIP have better ability to facilitate CIP transport ; 2) cations with stronge r complxation ability also more effectively remobilize presorbed CIP; 3) cations can not effectively mobilized CIP presorbed to native sand. CIP h as been reported with great sorption ability in batch experiment in different soil s, indicating low mobility. However, widely present divalent cations like Cu or Ca could significant ly influence CIP behavior in CIP sorption and transport For CIP sorption batch research using Ca to adjust ionic strength ( Zhang and Dong 2008 ; Wang et al. 2009 ) it may not be appropriate to precisely reflect the equilibrium CIP sorption behavior. For the dynamic transport system, Ca also has significant ability to promote CIP transport. In respect of the widely exist Ca in soil system, this promoted mobility of CIP should not be overlooked. Although only aqueous Cu and Ca were investigated in present study, CIP could form complexes with majo rity of metal cations in soils ( Turel 2002 ) suggesting other metals could have simliar e ffect on CIP transport. Some cations like Fe with extremely high complexation ability can be equivalent to much higher concentration of low complexation ability cations like Ca ( Turel et al. 1996 ) To better understand the total effect of all the cations high concentration catioins were also used, since the effect of these metals can add up. In addition, almost all FQs share the same carboxyl, keto and

PAGE 101

101 piperazine reactive functional group like CIP, similar chemodynamic s process was expectd for other FQs. Therefore, other major cations in aqueous and soil environment should also be evaluated to better understand the chemodynamics of FQs in soils and receiving waters, which is critial for comprehensive assessment of the p otential environmental risk of FQs. Colloids with strong sorption ability to CIP can also facilitate CIP transport. In this experiment only montmorillonite was tested, other colloids like goethite also has been reported have high sorption ability, as a re sult they could also be effective to facilitate CIP transport. Besides, soil minerals like iron or aluminum oxides have also been demonstrated to have strong sorption to CIP ( Gu and Karthikeyan 2005 ) they could also perform as CIP carrior for CIP transport. Therefore the conclusion that CIP is low mobility chemical in soil system based on the strong CIP sorption to soil minerals is misleading. In consider of high concen tration of CIP in water body has been reported colloid facilitated CIP transport should not be overlooked. To conclude, the chemical property of antibiotics greatly influence on the ir fate and transport in the environment Although SMZ and CIP were invest igated in present study, similar process was expected for other antibiotics with similar functional group. Therefore, other major solution chemistry properties in aqueous and soil environment should also be evaluated to better understand the fate and trans port of these antibiotics in soils and receiving waters, which is critical for comprehensive assessment of the potential environmental risk of antibiotics

PAGE 102

102 APPENDIX A ADSORPTION ISOTHERMS OF CIP TO SAND : A 1) NATIVE SAND, A 2) CLEAN SAND Figure A 1 Na tive sand isotherm CIP sorption isotherm

PAGE 103

103 Figure A 2 Clean sand isotherm CIP sorption isotherm

PAGE 104

104 APPENDIX B S CANNING ELECTRON MICROSCOPE IMAGE OF NATIVE SAND Figure B 1 S canning electron microscope Image of native sand

PAGE 105

105 LIST OF REFERENCES (FEDESA), E. F. o. A. H. (1997). "Antibiotics and Animals." FEDESA/FEFANA Press release 8 Aristilde, L. and G. Sposito (2008). "Molecular modeling of metal complexation by a fluoroquinolone antibiotic." Environ. Toxicol. Chem. 27 (11): 2304 2310 Barnes, K. K., S. C. Christenson, et al. (2004). "Pharmaceuticals and other organic waste water contaminants within a leachate plume downgradient of a municipal landfill." Ground Water Monitoring And Remediation 24 (2): 119 126. Barnes, K. K., S. C. Chris tenson, et al. (2004). "Pharmaceuticals and other organic waste water contaminants within a leachate plume downgradient of a municipal landfill." Ground Water Monit. Rem. 24 (2): 119 126. Bradford, S. A., J. Simunek, et al. (2005). "Straining of colloids at textural interfaces." Water Resources Research 41 (10): W10404, doi:10410.11029/12004WR003675. Brown, K. D., J. Kulis, et al. (2006). "Occurrence of antibiotics in hospital, residential, and dairy, effluent, municipal wastewater, and the Rio Grande in New Mexico." Science Of The Total Environment 366 (2 3): 772 783. Carrasquillo, A. J., G. L. Bruland, et al. (2008). "Sorption of Ciprofloxacin and Oxytetracycline Zwitterions to Soils and Soil Minerals: Influence of Compound Structure." Environ. Sci. Technol. 42 (20): 7634 7642. Carrasquillo, A. J., G. L. Bruland, et al. (2008). "Sorption of Ciprofloxacin and Oxytetracycline Zwitterions to Soils and Soil Minerals: Influence of Compound Structure." Environmental Science & Technology 42 (20): 7634 7642. Chander, Y. K. Kumar, et al. (2005). "Antibacterial activity of soil bound antibiotics." Journal of Environmental Quality 34 (6): 1952 1957. Chee Sanford, J. C., R. I. Aminov, et al. (2001). "Occurrence and diversity of tetracycline resistance genes in lagoons and gr oundwater underlying two swine production facilities." Applied and Environmental Microbiology 67 (4): 1494 1502. Chefetz, B., T. Mualem, et al. (2008). "Sorption and mobility of pharmaceutical compounds in soil irrigated with reclaimed wastewater." Chemosph ere 73 (8): 1335 1343. Cheng, T. and J. E. Saiers (2010). "Colloid Facilitated Transport of Cesium in Vadose Zone Sediments: The Importance of Flow Transients." Environmental Science & Technology 44 (19): 7443 7449.

PAGE 106

106 Delgado1, A. V., F. Gonzalez Caballero, et al. (2005). "Measurement and Interpretation of Electrokinetic Phenomena." Pure and Applied Chemistry 77 (10). Drakopoulos, A. I. and P. C. Ioannou (1997). "Spectrofluorimetric study of the acid base equilibria and complexation behavior of the fluoroquinolo ne antibiotics ofloxacin, norfloxacin, ciprofloxacin and pefloxacin in aqueous solution." Anal. Chim. Acta 354 (1 3): 197 204. Duckworth, O. W. and S. T. Martin (2001). "Surface complexation and dissolution of hematite by C1 C6 dicarboxylic acids at pH = 5. 0." Geochim. Cosmochim. Acta 65 (23): 4289 4301. Duckworth, O. W. and S. T. Martin (2001). "Surface complexation and dissolution of hematite by C1 C6 dicarboxylic acids at pH = 5.0." Geochimica et Cosmochimica Acta 65 (23): 4289 4301. Duong, H. A., N. H. Pha m, et al. (2008). "Occurrence, fate and antibiotic resistance of fluoroquinolone antibacterials in hospital wastewaters in Hanoi, Vietnam." Chemosphere 72 (6): 968 973. EPA "Test methods for evaluating solid waste, Laboratory Manual Physical/Chemical Method s, U.S.G.P. Office, Washington, DC, 1986.". Essington, M. E. (2003). "Soil and water chemistry an intergrative approach." Gao, B., J. E. Saiers, et al. (2004). "Deposition and mobilization of clay colloids in unsaturated porous media." Water Resources Rese arch 40 (8): W08602, doi:08610.01029/02004WR003189. Gao, J. A. and J. A. Pedersen (2005). "Adsorption of sulfonamide antimicrobial agents to clay minerals." Environmental Science & Technology 39 (24): 9509 9516. Gielen, G., M. R. van den Heuvel, et al. (2009 ). "Factors impacting on pharmaceutical leaching following sewage application to land." Chemosphere 74 (4): 537 542. Golet, E. M., A. Strehler, et al. (2002). "Determination of Fluoroquinolone Antibacterial Agents in Sewage Sludge and Sludge Treated Soil Us ing Accelerated Solvent Extraction Followed by Solid Phase Extraction." Analytical Chemistry 74 (21): 5455 5462. Goyne, K. W., J. Chorover, et al. (2005). "Sorption of the antibiotic ofloxacin to mesoporous and nonporous alumina and silica." J. Colloid Inte rface Sci. 283 (1): 160 170. Grolimund, D. and M. Borkovec (2005). "Col loid Facilitated Transport of Strongly 39 (17): 6378 6386.

PAGE 107

107 Gu, C. and K. G. Karthikeyan (2005). "Sorption of the Ant imicrobial Ciprofloxacin To Aluminum and Iron Hydrous Oxides." Environmental Science & Technology 39 (23): 9166 9173. Gu, C. and K. G. Karthikeyan (2005). "Sorption of the Antimicrobial Ciprofloxacin To Aluminum and Iron Hydrous Oxides." Environ. Sci. Techn ol. 39 (23): 9166 9173. Guaita, D. P., S. Sayen, et al. (2011). "Copper(II) influence on flumequine retention in soils: Macroscopic and molecular investigations." Journal of Colloid and Interface Science 357 (2): 453 459. Hari, A. C., R. A. Paruchuri, et al. (2005). "Effects of pH and cationic and nonionic surfactants on the adsorption of pharmaceuticals to a natural aquifer material." Environmental Science & Technology 39 (8): 2592 2598. Hari, A. C., R. A. Paruchuri, et al. (2005). "Effects of pH and Cationic and Nonionic Surfactants on the Adsorption of Pharmaceuticals to a Natural Aquifer Material." Environ. Sci. Technol. 39 (8): 2592 2598. Heberer, T., A. Mechlinski, et al. (2004). "Field studies on the fate and transport of pharmaceutical residues in bank f iltration." Ground Water Monitoring and Remediation 24 (2): 70 77. Holten Ltzhft, H. C., W. H. J. Vaes, et al. (2000). "Influence of pH and Other Modifying Factors on the Distribution Behavior of 4 Quinolones to Solid Phases and Humic Acids Studied by "Ne gligible Depletion" SPME HPLC." Environmental Science & Technology 34 (23): 4989 4994. Huang, L., H. Fang, et al. (2012). "Experiment on surface charge distribution of fine sediment." SCI CHINA SER E 55 (4): 1146 1152. Huang, P. M., T. S. C. Wang, et al. (19 77). "Retention of Phenolic Acids By Noncrystalline Hydroxy Aluminum and Iron Compounds and Clay Minerals of Soils." Soil Science 123 (4): 213 219. Hyun, S. and L. S. Lee (2005). "Quantifying the Contribution of Different Sorption Mechanisms for 2,4 Dichlo rophenoxyacetic Acid Sorption by Several Variable Charge Soils." Environmental Science & Technology 39 (8): 2522 2528. Johnson, L., A. Sabel, et al. (2008). "Emergence of fluoroquinolone resistance in outpatient urinary Escherichia coli isolates." American Journal of Medicine 121 (10): 876 884. Johnson, P. R., N. Sun, et al. (1996). "Colloid transport in geochemically heterogeneous porous media: Modeling and measurements." Environmental Science & Technology 30 (11): 3284 3293.

PAGE 108

108 Jones, O. A. H., N. Voulvoulis, e t al. (2001). "Human pharmaceuticals in the aquatic environment A review." Environ. Technol. 22 (12): 1383 1394. Kanti Sen, T. and K. C. Khilar (2006). "Review on subsurface colloids and colloid associated contaminant transport in saturated porous media." Advances in Colloid and Interface Science 119 (2 3): 71 96. KarcI, A. and I. A. BalcIoglu (2009). "Investigation of the tetracycline, sulfonamide, and fluoroquinolone antimicrobial compounds in animal manure and agricultural soils in Turkey." Science Of Th e Total Environment 407 (16): 4652 4664. Kay, P., P. A. Blackwell, et al. (2005). "Column studies to investigate the fate of veterinary antibiotics in clay soils following slurry application to agricultural land." Chemosphere 60 (4): 497 507. Kemper, N. (200 8). "Veterinary antibiotics in the aquatic and terrestrial environment." Ecological Indicators 8 (1): 1 13. Kolpin, D. W., E. T. Furlong, et al. (2002). "Pharmaceuticals, hormones, and other organic wastewater contaminants in US streams, 1999 2000: A nation al reconnaissance." Environ. Sci. Technol. 36 (6): 1202 1211. Kozak, G. K., D. L. Pearl, et al. (2009). "Distribution of sulfonamide resistance genes in Escherichia coli and Salmonella isolates from swine and chickens at abattoirs in Ontario and Quebec, Can ada." Applied and Environmental Microbiology 75 (18): 5999 6001. Kmmerer, K. (2001). "Drugs in the environment: emission of drugs, diagnostic aids and disinfectants into wastewater by hospitals in relation to other sources a review." Chemosphere 45 (6 7): 957 969. Lapworth, D. J., N. Baran, et al. (2012). "Emerging organic contaminants in groundwater: A review of sources, fate and occurrence." Environmental Pollution 163 : 287 303. Le Corre, K. S., C. Ort, et al. (2012). "Consumption based approach for asse ssing the contribution of hospitals towards the load of pharmaceutical residues in municipal wastewater." Environment International 45 : 99 111. Li, B. and T. Zhang (2010). "Biodegradation and Adsorption of Antibiotics in the Activated Sludge Process." Envi ronmental Science & Technology 44 (9): 3468 3473. Li, R. C., D. E. Nix, et al. (1994). "Interaction Between Ciprofloxacin and Metal Cations: Its Influence on Physicochemical Characteristics and Antibacterial Activity." Pharm. Res 11 (6): 917 920.

PAGE 109

109 Li, Y. x., W. Li, et al. (2007). "Contribution of additives Cu to its accumulation in pig feces: study in Beijing and Fuxin of China." Journal of Environmental Sciences 19 (5): 610 615. Lorphensri, O., D. A. Sabatini, et al. (2007). "Sorption and transport of acetamin ophen, 17 alpha ethynyl estradiol, nalidixic acid with low organic content aquifer sand." Water Research 41 (10): 2180 2188. Lucida, H., J. E. Parkin, et al. (2000). "Kinetic study of the reaction of sulfamethoxazole and glucose under acidic conditions I. Effect of pH and temperature." International Journal Of Pharmaceutics 202 (1 2): 47 61. Luo, Y., L. Xu, et al. (2011). "Occurrence and Transport of Tetracycline, Sulfonamide, Quinolone, and Macrolide Antibiotics in the Haihe River Basin, China." Environmen tal Science & Technology 45 (5): 1827 1833. MacKay, A. A. and D. E. Seremet (2008). "Probe Compounds to Quantify Cation Exchange and Complexation Interactions of Ciprofloxacin with Soils." Environ. Sci. Technol. 42 (22): 8270 8276. Martins, A. F., T. G. Vasc oncelos, et al. (2008). "Concentration of ciprofloxacin in Brazilian hospital effluent and preliminary risk aassessment: A case study." Clean Soil Air Water 36 (3): 264 269. Molis, E., O. Barr¨s, et al. (2000). "Initial steps of ligand promoted dissolution of gibbsite." Colloids Surf., A 163 (2 3): 283 292. Muchuweti, M., J. W. Birkett, et al. (2006). "Heavy metal content of vegetables irrigated with mixtures of wastewater and sewage sludge in Zimbabwe: Implications for human health." Agriculture, Ecosystems & Environment 112 (1): 41 48. Nowara, A., J. Burhenne, et al. (1997). "Binding of Fluoroquinolone Carboxylic Acid Derivatives to Clay Minerals." J. Agric. Food. Chem. 45 (4): 1459 1463. Nygaard, K., B. T. Lunestad, et al. (1992). "Resistance to oxytetracycl ine, oxolinic acid and furazolidone in bacteria from marine sediments." Aquaculture 104 (1 2): 31 36. Oker, H. M. and I. Akmehmet BalcIoglu (2005). "Adsorption and degradation of enrofloxacin, a veterinary antibiotic on natural zeolite." J Hazard Mater. 122 (3): 251 258. Overcash, M., R. C. Sims, et al. (2005). "Beneficial reuse and sustainability: The fate of organic compounds in land applied waste." Journal of Environmental Quality 34 (1): 29 41.

PAGE 110

110 Park, H. R., T. H. Kim, et al. (2002). "Physicochemical proper ties of quinolone antibiotics in various environments." European Journal of Medicinal Chemistry 37 (6): 443 460. Park, H. R., K. Y. Chung, et al. (2000). "Ionization and divalent cation complexation of quinolone antibiotics in aqueous solution." Bulletin of the Korean Chemical Society 21 (9): 849 854. Park, H. R., J. J. Seo, et al. (2007). "Fluorescence quenching of norfloxacin by divalent transition metal cations." Bulletin of the Korean Chemical Society 28 (9): 1573 1578. Pedersen, J. A., J. Gao, et al. (200 3). "Sorption of sulfonamide antimicrobial agents to soil constituents." Abstracts of Papers of the American Chemical Society 226 : U493 U494. Pei, Z., X. Q. Shan, et al. (2009). "Coadsorption of Ciprofloxacin and Cu(II) on Montmorillonite and Kaolinite as Affected by Solution pH." Environ. Sci. Technol. 44 (3): 915 920. Pei, Z. G., X. Q. Shan, et al. "Coadsorption of Ciprofloxacin and Cu(II) on Montmorillonite and Kaolinite as Affected by Solution pH." Environmental Science & Technology 44 (3): 915 920. Riese nfeld, C. S., R. M. Goodman, et al. (2004). "Uncultured soil bacteria are a reservoir of new antibiotic resistance genes." Environmental Microbiology 6 (9): 981 989. Riley, C. M., D. L. Ross, et al. (1993). "Characterization of the complexation of fluoroqui nolone antimicrobials with metal ions by nuclear magnetic resonance spectroscopy." Journal of Pharmaceutical and Biomedical Analysis 11 (1): 49 59. Ross, D. L. and C. M. Riley (1992). "Physicochemical properties of the fluoroquinolone antimicrobials. III. C omplexation of lomefloxacin with various metal ions and the effect of metal ion complexation on aqueous solubility." Int. J. Pharm. 87 (1 3): 203 213. Roy, S. B. and D. A. Dzombak (1997). "Chemical Factors Influencing Colloid Facilitated Transport of Contam inants in Porous Media." Environmental Science & Technology 31 (3): 656 664. Spongberg, A. L. and J. D. Witter (2008). "Pharmaceutical compounds in the wastewater process stream in Northwest Ohio." Science Of The Total Environment 397 (1 3): 148 157. Srivast ava, P., S. M. Sanders, et al. (2009). "Fate and Transport of Sulfadimethoxine and Ormetoprim in Two Southeastern United States Soils." Vadose Zone Journal 8 (1): 32 41.

PAGE 111

111 Stoob, K., H. P. Singer, et al. (2007). "Dissipation and transport of veterinary sulfon amide antibiotics after manure application to grassland in a small catchment." Environ. Sci. Technol. 41 (21): 7349 7355. Szczepanowski, R., B. Linke, et al. (2009). "Detection of 140 clinically relevant antibiotic resistance genes in the plasmid metagenome of wastewater treatment plant bacteria showing reduced susceptibility to selected antibiotics." Microbiology Sgm 155 : 2306 2319. Taboada Serrano, P., V. Vithayaveroj, et al. (2005). "Surface Charge Heterogeneities viron. Sci. Technol. 39 (17): 6352 6360. Thiele Bruhn, S. (2003). "Pharmaceutical antibiotic compounds in soils a review." Journal of Plant Nutrition and Soil Science Zeitschrift Fur Pflanzenernahrung Und Bodenkunde 166 (2): 145 167. Thiele Bruhn, S., T. S eibicke, et al. (2004). "Sorption of sulfonamide pharmaceutical antibiotics on whole soils and particle size fractions." Journal Of Environmental Quality 33 (4): 1331 1342. Threats, F. o. M. (2009). Microbial Evolution and Co Adaptation:A Tribute to the Lif e and Scientific Legacies of Joshua Lederberg, The National Academies Press. Tian, Y., B. Gao, et al. (2010). "Transport of Engineered Nanoparticles in Saturated Porous Media." Journal of Nanoparticle Research: DOI: 10.1007/s11051 11010 19912 11057. Tolls, J. (2001). "Sorption of veterinary pharmaceuticals in soils: A review." Environmental Science & Technology 35 (17): 3397 3406. Toride, N., F. J. Leij, et al. (1995). "The CXTFIT code for estimating transport parameters from laboratory or field tracer exper iments, Version 2.0. Research Report No. 137, 121 p., U.S. Salinity Laboratory, USDA, ARS, Riverside, California." Trivedi, P., L. Axe, et al. (2001). "Adsorption of metal ions onto goethite: single adsorbate and competitive systems." Colloids Surf., A 191 (1 2): 107 121. Trivedi, P. and D. Vasudevan (2007). "Spectroscopic Investigation of Ciprofloxacin Speciation at the Goethite water Interface." Environmental Science & Technology 41 (9): 3153 3158. Turel, I. (2002). "The interactions of metal ions with quin olone antibacterial agents." Coord. Chem. Rev. 232 (1 2): 27 47. Turel, I., N. Bukovec, et al. (1996). "Complex formation between some metals and a quinolone family member (ciprofloxacin)." Polyhedron 15 (2): 269 275.

PAGE 112

112 Turel, I., I. Leban, et al. (1994). "Syn thesis, characterization, and crystal structure of a copper(II) complex with quinolone family member (ciprofloxacin): bis(1) cyclopropyl 6 fluoro 1,4 dihydro 4 oxo 7 piperazin 1ylquinoline 3 carboxylate) copper(II) chloride hexahydrate." J. Inorg. Biochem. 56 (4): 273 282. Turner, N. B., J. N. Ryan, et al. (2006). "Effect of desorption kinetics on colloid facilitated transport of contaminants: Cesium, strontium, and illite colloids." Water Resour. Res. 42 (12): W12S09. Unold, M., J. Simunek, et al. (2009). "T ransport of Manure Based Applied Sulfadiazine and Its Main Transformation Products in Soil Columns." VADOSE ZONE J 8 (3): 677 689. Upadhyay, S., P. Kumar, et al. (2006). "Complexes of quinolone drugs norfloxacin and ciprofloxacin with alkaline earth metal p erchlorates." Journal of Structural Chemistry 47 (6): 1078 1083. Uslu, M., A. Yediler, et al. (2008). "Analysis and Sorption Behavior of Fluoroquinolones in Solid Matrices." Water, Air, Soil Pollut. 190 (1): 55 63. Vasudevan, D., G. L. Bruland, et al. (2009) "pH dependent ciprofloxacin sorption to soils: Interaction mechanisms and soil factors influencing sorption." Geoderma 151 (3 4): 68 76. Vazquez, J. L., M. Berlanga, et al. (2001). "Determination by fluorimetric titration of the ionization constants of ci profloxacin in solution and in the presence of liposomes." Photochemistry And Photobiology 73 (1): 14 19. Violante, A., M. Ricciardella, et al. (2003). "Adsorption of Heavy Metals on Mixed Fe Al Oxides in the Absence or Presence of Organic Ligands." Water, Air, Soil Pollut. 145 (1): 289 306. Vulliet, E., C. Cren Olive, et al. (2011). "Occurrence of pharmaceuticals and hormones in drinking water treated from surface waters." Environmental Chemistry Letters 9 (1): 103 114. Wallis, S. C., L. R. Gahan, et al. (199 6). "Copper(II) complexes of the fluoroquinolone antimicrobial ciprofloxacin. Synthesis, X ray structural characterization, and potentiometric study." Journal of Inorganic Biochemistry 62 (1): 1 16. Wang, C. J., Z. H. Li, et al. (2011). "Adsorption of cipro floxacin on 2:1 dioctahedral clay minerals." Applied Clay Science 53 (4): 723 728. Wang, Z., X. Yu, et al. (2009). "Norfloxacin Sorption and Its Thermodynamics on Surface Modified Carbon Nanotubes." Environmental Science & Technology 44 (3): 978 984.

PAGE 113

11 3 Wehrhan A., R. Kasteel, et al. (2007). "Transport of sulfadiazine in soil columns Experiments and modelling approaches." Journal of Contaminant Hydrology 89 (1 2): 107 135. Yost, E. C., M. I. Tejedor Tejedor, et al. (1990). "In situ CIR FTIR characterization of salicylate complexes at the goethite/aqueous solution interface." Environmental Science & Technology 24 (6): 822 828. Zhang, H. and C. H. Huang (2005). "Oxidative Transformation of Fluoroquinolone Antibacterial Agents and Structurally Related Amines by Man ganese Oxide." Environ. Sci. Technol. 39 (12): 4474 4483. Zhang, J. Q. and Y. H. Dong (2008). "Effect of low molecular weight organic acids on the adsorption of norfloxacin in typical variable charge soils of China." Journal of Hazardous Materials 151 (2 3): 833 839. Zorita, S., L. Martensson, et al. (2009). "Occurrence and removal of pharmaceuticals in a municipal sewage treatment system in the south of Sweden." Science of the Total Environment 407 (8): 2760 2770.

PAGE 114

114 BIOGRAPHICAL SKETCH Hao Chen, the only ch ild of her parents, was born and brought up in Yan Cheng city Jiangsu province China. After graduated from high school she join ed the Nanjing University of Technology for her environmental engineering in 2004 Following this she ob tained a environmental science at the Institution of Soil Science Chinese Academy of Sciences Sulfonamid es antibiotic residual in Jiangsu province and sorption on Taihu paddy s oil In 2008 she joined the University of Florida, Gainesville, Florida to pursue her PhD in S oil and W ater S cience Department Here she was a graduate research assistant and worked on the antibiotic sorption and transport in porous media