FACTORS GOVERNING SORPTION OF DISSOLVED ORGANIC MATTER AND PHARMACEUTICALS IN SOIL

by

Stephanie Clare Hofley

A thesis submitted in conformity with the requirements for the degree of Master of Science, Graduate Department of Chemistry, in the University of Toronto

© Copyright by Stephanie Clare Hofley (2012)

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Abstract “Factors governing sorption of dissolved organic matter and pharmaceuticals in soil” by Stephanie Clare Hofley (2012), for the degree of Master of Science, Graduate Department of Chemistry, in the University of Toronto.

Pharmaceuticals, personal care products and dissolved organic matter (OM) are introduced to

soil via irrigation with reclaimed wastewater. This thesis examines the basic factors that

influence sorption of these components in soil. Sorption of dissolved OM samples of varying

composition to clay surfaces was examined. Results indicate that preferential sorption is

dependent on clay type but not necessarily OM composition. Analysis of soils revealed

aliphatic components, carbohydrates and amino acids are prevalent at the soil-water interface

whereas aromatics are inaccessible at the soil-water interface. No clear relationship between

sorption affinity of 17β-estradiol, sulfamethoxazole, carbamazepine and phenanthrene and

soil OM aromaticity or aliphaticity was observed. A negative relationship between sorption

and O-alkyl content may be due to these components blocking contaminant access to high

affinity sorption sites. Therefore, application of reclaimed wastewater to soils with O-alkyl-

rich OM may result in higher mobility of contaminants.

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Acknowledgements I would like to extend my gratitude to Professor Myrna Simpson for her guidance throughout

my research. As well, I would like to thank my committee members; Professor Andre

Simpson for his assistance with the NMR and Professor Frank Wania for reviewing this

thesis.

I would like to thank my colleagues for their support. In particular, I would like to thank Dr.

Denis Courtier-Murias and Dr. Ronald Soong for their assistance in acquiring and analyzing

the NMR spectra.

Funding for this research was provided by the NSERC-BARD Canada-Israel Research

Program (CA-9114-09). I would like to acknowledge the University of Toronto for financial

support through the U of T fellowship, the Relocation Assistance Award and the SGS

Conference Grant. Financial assistance from the Ontario Graduate Scholarship Program is

also appreciated.

I would like to thank my family, boyfriend and friends for their love and encouragement. I

especially want to thank my boyfriend for his patience and assistance with proofreading as

well as my mom for everything she has done to help me achieve my goals.

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Table of Contents Abstract ..................................................................................................................................... ii 

Acknowledgements .................................................................................................................. iii 

Table of Contents ..................................................................................................................... iv 

List of Tables ........................................................................................................................... vi 

List of Figures ......................................................................................................................... vii 

List of Appendices ................................................................................................................. viii 

Chapter 1: Introduction ............................................................................................................. 1 

1.1. Reclaimed wastewater.................................................................................................... 1 

1.2. Pharmaceuticals and personal care products in sewage treatment plants ...................... 1 

1.3. Fate of pharmaceuticals and personal care products in soil ........................................... 2 

1.4. Sorption of pharmaceuticals and personal care products ............................................... 4 

1.4.1. Relationships between sorption and soil OM composition ..................................... 4 

1.4.2. Dual-mode sorption model ..................................................................................... 5 

1.5. OM in the soil environment ........................................................................................... 6 

1.6. Sorption of OM .............................................................................................................. 7 

1.6.1. Preferential sorption of OM to minerals ................................................................. 7 

1.6.2. Zonal model of organo-mineral interactions ........................................................... 9 

1.7. Sorption mechanisms ................................................................................................... 10 

1.7.1. Hydrophobic effects .............................................................................................. 10 

1.7.2. Cation and water bridging ..................................................................................... 11 

1.7.3. Cation and anion exchange ................................................................................... 12 

1.7.4. Ligand exchange ................................................................................................... 13 

1.8. Objectives .................................................................................................................... 14 

1.9. References .................................................................................................................... 17 

1.10. Figures ........................................................................................................................ 23 

Chapter 2: Nuclear Magnetic Resonance (NMR) Analysis of Soil Organic Matter at the Solid-Water Interface ............................................................................................................. 25 

2.1. Introduction .................................................................................................................. 25 

2.2. Materials and Methods ................................................................................................. 28 

2.2.1. Organo-clay complex preparation and carbon analysis ........................................ 28 

2.2.2. Soil samples and preparation ................................................................................ 30 

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2.2.3. NMR analysis of dissolved OM and organo-clay complexes ............................... 31 

2.2.4. NMR analysis of soils and humin ......................................................................... 33 

2.3. Results and Discussion ................................................................................................ 34 

2.3.1. NMR characterization of dissolved OM and organo-clay complexes .................. 34 

2.3.2. NMR characterization of soils and humin ............................................................ 42 

2.4. Conclusions .................................................................................................................. 45 

2.5. References .................................................................................................................... 47 

2.6. Tables ........................................................................................................................... 51 

2.7. Figures .......................................................................................................................... 52 

Chapter 3: Sorption of Carbamazepine, Sulfamethoxazole, 17β-Estradiol and Phenanthrene to Soils with Varying Organic Matter Composition ............................................................... 59 

3.1. Introduction .................................................................................................................. 59 

3.2. Materials and Methods ................................................................................................. 63 

3.2.1. Soil and mineral samples ...................................................................................... 63 

3.2.2. Solid-state 13C NMR analysis ............................................................................... 64 

3.2.3. Batch sorption experiments ................................................................................... 65 

3.3. Results and Discussion ................................................................................................ 68 

3.3.1. Sorbent characteristics .......................................................................................... 68 

3.3.2. Sorption coefficients ............................................................................................. 70 

3.3.3. Comparison of measured and calculated sorption coefficients ............................. 72 

3.3.4. Relationship between sorption and OM structure ................................................. 74 

3.4. Conclusions .................................................................................................................. 79 

3.5. References .................................................................................................................... 81 

3.6. Tables ........................................................................................................................... 87 

3.7. Figures .......................................................................................................................... 90 

Chapter 4: Summary and Synthesis ........................................................................................ 92 

4.1. References .................................................................................................................... 96 

Appendices .............................................................................................................................. 98 

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List of Tables

Table 2.1: Solution-state 1H NMR integration results for the dissolved OM isolates prior to sorption. .................................................................................................................................. 51 

Table 3.1: Selected chemical and physical properties of the contaminants. ........................... 87 

Table 3.2: Solid-state 13C NMR integration results for soil samples used in sorption studies. ................................................................................................................................................. 87 

Table 3.3: Freundlich and linear sorption isotherm parameters for contaminant sorption to various soils. Contaminant abbreviations are as follows: carbamazepine (CBZ), sulfamethoxazole (SMX), 17β-estradiol (E2) and phenanthrene (PHN). ............................... 88 

Table 3.4: Sorbate descriptors for the studied contaminants. Abbreviations are as follows: excess molar refractivity (E), molar volume (V), H-bond acidity (A) and basicity (B) and dipolarity/ polarizability (S). ................................................................................................... 89 

Table 3.5: Linear regression parameters for the relationship between organic carbon normalized sorption coefficients (Koc) and O-alkyl carbon content. ...................................... 89 

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List of Figures

Figure 1.1: Possible fates of contaminants introduced to the soil environment. Reprinted from: Agriculture, Ecosystems & Environment, Vol 120, Müller, K., Magesan, G.N., Bolan, N.S., A critical review of the influence of effluent irrigation on the fate of pesticides in soil., 93-116, Copyright (2007), with permission from Elsevier ..................................................... 23 

Figure 1.2: Zonal model of organo-mineral interactions. With kind permission from Springer Science + Business Media: Biogeochemistry, A conceptual model of organo-mineral interactions in soils: Self-assembly of organic molecular fragments into zonal structures on mineral surfaces, Vol 85, 2007, 9-24, Kleber, M., Sollins, P., Sutton, R., Figure 2. .............. 24 

Figure 2.1: 1H solution-state NMR spectra in D2O of the dissolved OM isolates, a) forest soil-derived dissolved OM, b) Leonardite humic acid and c) Peat humic acid, prior to sorption to clay mineral surfaces. ........................................................................................... 52 

Figure 2.2: 1H HR-MAS NMR spectra of organo-kaolinite complexes swollen in a) D2O and b) DMSO-d6. Enlargements (×16) of the aromatic regions are provided in the boxes above the spectra. .............................................................................................................................. 53 

Figure 2.3: 1H HR-MAS NMR spectra of organo-montmorillonite complexes swollen in a) D2O and b) DMSO-d6. Enlargements (×16) of the aromatic regions are provided in the boxes above the spectra. .................................................................................................................... 54 

Figure 2.4: Solid-state 13C NMR spectrum of the HF treated Northern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of the 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents. .................................................................................................................................. 55 

Figure 2.5: Solid-state 13C NMR spectrum of the HF treated Southern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of the 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents. .................................................................................................................................. 56 

Figure 2.6: Solid-state 13C NMR spectrum and 1H HR-MAS NMR spectra of the untreated Peat soil (in D2O and DMSO-d6). Aromatic regions of 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents. ...................................................... 57 

Figure 2.7: Solid-state 13C NMR spectrum of HF treated humin isolated from the Northern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents. ...................................................................................................... 58 

Figure 3.1: Solid-state 13C cross polarization-magic angle spinning (CP-MAS) NMR spectra and organic carbon content of soils. ....................................................................................... 90 

Figure 3.2: Relationships between organic carbon normalized sorption coefficients (Koc) and soil a) alkyl carbon content, b) aromatic carbon content and c) O-alkyl carbon content. ...... 91 

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List of Appendices

Appendix A: Sorption Isotherms ............................................................................................ 98 

Appendix B: Aqueous-Phase Concentrations and Equilibrium Solid-Phase Concentrations for Contaminant Sorption to Soil ................................................................................................ 102 

Appendix C: Relationships between Distribution Coefficients and Fraction of Organic Carbon ................................................................................................................................... 107 

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Chapter 1: Introduction

1.1. Reclaimed wastewater

The application of reclaimed wastewater for irrigation purposes is a way to reduce the

demand for potable water (Kinney et al., 2006; Ternes et al., 2007; Tamtam et al., 2011).

Unfortunately, concern about this practice has arisen as a variety of pharmaceuticals and

personal care products have been detected in wastewater effluents as well as fields which

have been irrigated by this method (Carballa et al., 2004; Braga et al., 2005; Castiglioni et al.,

2006; Kinney et al., 2006; Xu et al., 2007; Jelic et al., 2011; Tamtam et al., 2011; Zhang et

al., 2011). Also present in reclaimed wastewater is a considerable amount of dissolved

organic matter (OM) which is co-introduced to soil with the pharmaceuticals and personal

care products (Katsoyiannis and Samara, 2007). It is necessary to understand how these

components will interact with each other as well as soil components in order to predict their

fate and transport in the environment and to make educated decisions regarding the use of

reclaimed wastewater for irrigation.

1.2. Pharmaceuticals and personal care products in sewage treatment plants

Pharmaceuticals and personal care products enter the sewage system either through excretion

from the human body in an unmetabolized form or by flushing down toilets or drains

(Carballa et al., 2004; Braga et al., 2005; Castiglioni et al., 2006; Jelic et al., 2011). Once

contaminants enter a sewage treatment plant the possible mechanisms for removal from

wastewater include sorption to sewage sludge, volatilization and/or chemical or biological

transformation (Omil et al., 2010). However, the majority of treatment plants were built prior

to the discovery that pharmaceuticals and personal care products act as water contaminants

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and were therefore not designed for their effective removal (Kinney et al., 2006). Oftentimes,

these contaminants are resistant to degradation, have a low volatility, and are fairly water

soluble which results in low sorption to sludge (Carballa et al., 2004; Perez et al., 2005; Omil

et al., 2010). Even for compounds which are moderately biodegradable, often the hydraulic

retention time (which describes the length of time which a soluble contaminant will reside in

a sewage treatment plant) is not long enough to allow for complete removal (Perez et al.,

2005; Omil et al., 2010). Whereas additional treatments such as nitrification, enhanced

chemical treatment, oxidation and ozonation may increase the removal of more persistent

compounds, there is a significant cost involved in implementing these technologies in older

treatment plants (Perez et al., 2005; Castiglioni et al., 2006; Xu et al., 2007; Caliman and

Gavrilescu, 2009). The inefficient removal of many pharmaceuticals and personal care

products within treatment plants therefore results in their discharge to the environment

(Carballa et al., 2004; Braga et al., 2005; Castiglioni et al., 2006; Xu et al., 2007; Jelic et al.,

2011; Zhang et al., 2011).

1.3. Fate of pharmaceuticals and personal care products in soil

The hazards associated with the presence of pharmaceuticals and personal care products in

soils are still under investigation (Brooks et al., 2009; Caliman and Gavrilescu, 2009). There

is concern that constant exposure to low levels of antibiotics will cause the development of

resistant bacteria (Thiele-Bruhn et al., 2004; Hou et al., 2010). Furthermore, it has been

shown that certain pharmaceuticals can negatively impact the growth and germination of

crops which would be detrimental to agriculture (Boxall et al., 2006; D'Abrosca et al., 2008;

Aristilde et al., 2010). Current interest therefore lies in determining the typical classes and

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concentrations of pharmaceuticals and personal care products released to the soil through

reclaimed wastewater as well as their fate in the environment (Brooks et al., 2009; Caliman

and Gavrilescu, 2009; Pan et al., 2009).

Once introduced to soil, there are a variety of processes which a contaminant could undergo

(Fig. 1.1). Sorption through interactions with soil minerals and/or OM is one such process

(Semple et al., 2003; Sparks, 2003). This may reduce the bioavailability of the contaminant

and therefore its biodegradability which could result in accumulation in upper soil layers

(Semple et al., 2003; McAllister and Semple, 2010). It has been determined that certain

pharmaceuticals and personal care products are mobile in soil which leads to leaching and

detection of these contaminants in groundwater (Ternes et al., 2007; Barnes et al., 2008; Xu

et al., 2010). Pharmaceuticals have also been detected in plants such as soybeans, carrot roots

and lettuce leaves grown on wastewater irrigated fields (Boxall et al., 2006; Wu et al., 2010).

Concentrations of individual pharmaceuticals and personal care products may be low and

some may even be metabolized in the plants, but humans may still be exposed to mixtures for

which the potential health effects of long-term exposure are unknown (Boxall et al., 2006).

Plant uptake of pharmaceuticals generally occurs as a partitioning process from soil water to

the plant root and is most favorable for compounds of intermediate hydrophobicity (Wu et

al., 2010). If a contaminant is removed from solution through sorption to soil, its ability to

partition into roots is reduced (Wu et al., 2010). An understanding of the factors influencing

sorption of pharmaceuticals and personal care products is necessary as this process is central

in determining contaminant fate. Sorption limits leaching and groundwater contamination as

well as reduces the bioavailability of contaminants and therefore reduces the potential for

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biodegradation and uptake into plants (Semple et al., 2003; Ternes et al., 2007; McAllister

and Semple, 2010; Wu et al., 2010).

1.4. Sorption of pharmaceuticals and personal care products

1.4.1. Relationships between sorption and soil OM composition

The sorption of pharmaceuticals and personal care products to soils and sediments is a

relatively recent area of research and therefore, more studies are required as many aspects are

still poorly understood (Halling-Sørensen et al., 1998; Pan et al., 2009; Pignatello et al.,

2010). Pharmaceutical and personal care product sorption and mobility has been related to

the OC content and mineral content of soils as well as properties of the contaminant such as

solubility and log Kow (Lee et al., 2003; Oppel et al., 2004; Yu et al., 2004; Drillia et al.,

2005; Williams et al., 2006; Chefetz et al., 2008; Sanders et al., 2008; Xu et al., 2009;

Karnjanapiboonwong et al., 2010). However, extensive research with other classes of

contaminants has shown that sorption is related not only to OC content but also to the

composition of the soil OM (Chin et al., 1997; Xing, 1997; Chiou et al., 1998; Perminova et

al., 1999; Chefetz et al., 2000; Mao et al., 2002; Salloum et al., 2002; Niederer et al., 2007).

Many studies have observed positive correlations between sorbent OM aromaticity and Koc

values of a contaminant in such OM (Chin et al., 1997; Xing, 1997; Chiou et al., 1998;

Perminova et al., 1999; Niederer et al., 2007). Other studies have suggested that aliphatic

components of OM have a higher sorption affinity for contaminants (Chefetz et al., 2000;

Mao et al., 2002; Salloum et al., 2002; Wang et al., 2011). A third hypothesis is that neither

OM aromaticity nor aliphaticity alone are well correlated to the degree of contaminant

sorption but that OM conformation and accessibility must also be considered (Simpson et al.,

5

2003; Chen et al., 2005; Bonin and Simpson, 2007b; Chefetz and Xing, 2009). These

observations may also apply to the sorption of pharmaceuticals and personal care products

but there are relatively few studies on the relationships between sorption of these

contaminants and OM composition (Yamamoto et al., 2003; Thiele-Bruhn et al., 2004; Bonin

and Simpson, 2007a; Sun et al., 2007; Hou et al., 2010). There have been studies involving

sorption of a large variety of polar and non-polar organic contaminants, including some with

chemical and physical properties similar to those of pharmaceuticals and personal care

products (Niederer et al., 2007; Bronner and Goss, 2011). However, these studies have also

not reached a general consensus regarding which soil OM properties govern sorption

(Niederer et al., 2007; Bronner and Goss, 2011). Therefore, which components of soil OM

are most important for sorption of pharmaceuticals and personal care products requires

further study.

1.4.2. Dual-mode sorption model

Another way of describing sorption of contaminants to OM is referred to as the dual-mode

sorption model and separates OM into two domains: glassy or condensed/rigid and rubbery

or expanded (Xing et al., 1996; Xing and Pignatello, 1997). The dual-mode sorption model

developed from observations that OM is a polymer-like species composed of

macromolecules and therefore can act in a similar fashion to polymers (Xing et al., 1996;

Xing and Pignatello, 1997). Sorption of contaminants to the rubbery domain is thought to

occur via dissolution or partitioning mechanisms (Xing et al., 1996; Xing and Pignatello,

1997; Huang et al., 2003). As sorption to this domain is nonspecific, it is concentration-

independent and results in a linear sorption isotherm (Xing et al., 1996; Xing and Pignatello,

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1997). Components of soil OM which may contribute to the rubbery domain include:

amorphous polymethylene and polysaccharides such as cellulose and chitin (Xing et al.,

1996; Mao et al., 2002). The glassy or condensed domain is thought to be associated with a

hole-filling mechanism which occurs through adsorption rather than partitioning and may

account for as much as 68% of total sorption (Xing et al., 1996; Xing and Pignatello, 1997;

Huang et al., 2003). These are site-specific interactions with a maximum amount of

contaminant that can be adsorbed to the surface (Xing et al., 1996; Xing and Pignatello,

1997). Competition between contaminants is observed and sorption levels off with increasing

contaminant concentration resulting in isotherm non-linearity (Xing et al., 1996; Xing and

Pignatello, 1997). The main contributors to the glassy domain are thought to be aromatic

compounds as they have been shown to be more rigid than aliphatic compounds and inverse

relationships between sorbent aromaticity and sorption linearity have been observed (Chien

and Bleam, 1998; Xing, 2001; Chefetz and Xing, 2009).

1.5. OM in the soil environment

OM is important for soil fertility and nutrient cycling, helps to retain water and reduces

erosion (Weil and Magdoff, 2004). In addition to dissolved OM introduced via irrigation

with reclaimed wastewater, sources of OM in soils include plant litter, microbial biomass and

organic amendments such as manures or biosolids (Kalbitz et al., 2000; Ohno et al., 2007).

Sorption to minerals can reduce the biodegradation of OM either through the formation of

strong chemical bonds or by trapping in pores which limits accessibility for microorganisms

(Kalbitz et al., 2005; Mikutta et al., 2007). In this way, molecules which are expected to

degrade rapidly such as simple carbohydrates and low molecular weight organic acids may in

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fact prove to be more recalcitrant in soils (Kalbitz et al., 2005; Mikutta et al., 2007).

Therefore, sorption greatly influences global carbon cycling and leads to soil being one of the

largest sinks for organic carbon with approximately 1600 Pg of carbon stored in the top 100

cm of soils (Eswaran et al., 1993). As well, a layer of OM can drastically alter mineral

surface properties such as surface charge density or hydrophobicity when as little as 0.1% of

the surface is covered (Schwarzenbach and Westall, 1981; Murphy et al., 1990; Day et al.,

1994). Often, the result is an increase in the sorption of organic contaminants on OM coated

minerals in comparison to uncoated minerals (Murphy et al., 1990; Wang and Xing, 2005).

OM plays a key role in determining the transport of contaminants in the soil environment

since OM retained on the mineral surface may decrease contaminant mobility by providing

additional binding sites through a cumulative sorption mechanism (Totsche and Kögel-

Knabner, 2004; Müller et al., 2007; Chefetz et al., 2008). Dissolved OM may conversely

increase the mobility of contaminants through the formation of soluble complexes or by

acting as a competitor for soil binding sites (Kan and Tomson, 1990; Totsche and Kögel-

Knabner, 2004; Müller et al., 2007; Chefetz and Xing, 2009). For these reasons, it is

necessary to understand the fundamental processes controlling sorption of OM to minerals.

1.6. Sorption of OM

1.6.1. Preferential sorption of OM to minerals

Sorption experiments of OM to mineral surfaces show saturation behaviour suggesting that

minerals have a limited number of sorption sites (Day et al., 1994; Kalbitz et al., 2000). As

well, it has been shown that previously sorbed soil OM can reduce the sorption of newly

added dissolved OM and that removal of sorbed OM may reveal additional sorption sites

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(Kaiser and Zech, 2000; Jin et al., 2008). This limited number of sorption sites in soil leads to

competition and preferential sorption of certain OM components and which components are

retained is influenced by different environmental factors (Namjesnik-Dejanovic et al., 2000;

Chorover and Amistadi, 2001; Feng et al., 2005; Wang and Xing, 2005; Majzik and

Tombacz, 2007; Ghosh et al., 2009). Many studies have shown that preferential sorption of

OM components is dependent on the types of mineral surfaces present in the soil (Chorover

and Amistadi, 2001; Feng et al., 2005; Wang and Xing, 2005; Ghosh et al., 2009). Different

components of peat humic acid were observed to sorb on three different mineral surfaces,

kaolinite, montmorillonite and goethite (Ghosh et al., 2009). Kaolinite retained more non-

polar aliphatic compounds and carbohydrates, montmorillonite preferentially retained

paraffinic fractions and OM containing many carboxylate groups and aromatic fractions were

preferentially sorbed to goethite (Ghosh et al., 2009). Another group used molar absorptivity

of an OM solution as a measure of sample aromaticity and observed a decrease upon sorption

to goethite and birnessite suggesting preferential sorption of aromatic fractions but observed

no change after sorption to montmorillonite (Chorover and Amistadi, 2001). Characterization

of OM extracted from different mineral surfaces showed that a large amount of aliphatic

compounds were associated with both smectite and kaolinite surfaces (Wattel-Koekkoek et

al., 2001; Clemente et al., 2011). However, kaolinite was associated with a large number of

polysaccharides while more aromatic compounds were extracted from the smectite surface

(Wattel-Koekkoek et al., 2001). Furthermore, it has been shown that preferential sorption can

vary with experimental conditions such as pH, ionic strength, and solution cation, as well as

the dominant exchangeable cation on the mineral surface (Namjesnik-Dejanovic et al., 2000;

Feng et al., 2005; Majzik and Tombacz, 2007; Polubesova et al., 2008). One study observed

9

that lowering the solution pH from 7 to 4 reduced the sorption of peptides to montmorillonite

(Feng et al., 2005). Another group observed preferential sorption of aromatic components to

Fe3+-enriched montmorillonite but not Cu2+-enriched or crude-montmorillonite (Polubesova

et al., 2008). It has been suggested that the observed differences in preferential sorption may

be governed by which mechanism of interaction (see section 1.7) is dominant for a particular

mineral type or for different solution conditions (Wattel-Koekkoek et al., 2001; Feng et al.,

2005; Polubesova et al., 2008).

1.6.2. Zonal model of organo-mineral interactions

Recently, a zonal model has been proposed for the sorption of OM to mineral surfaces (Fig.

1.2) which suggests that OM is retained in zones namely the contact zone, the hydrophobic

zone and the kinetic zone (Kleber et al., 2007). The contact zone is composed of the fraction

of OM which is directly attached to the mineral surface (Kleber et al., 2007). The OM in this

zone is the most strongly sorbed and is thought to account for around 65% of the OM

coverage of mineral surfaces (Kleber et al., 2007). This layer is formed relatively quickly and

may interact favorably with other organic molecules to reduce their mobility in the soil

(Kleber et al., 2007). In particular, the authors suggest that proteins play an important role in

the formation of this zone since these molecules interact strongly with certain mineral

surfaces and contain a variety of functional groups which can provide additional binding sites

for other organic molecules (Kleber et al., 2007). If the molecules in contact with mineral

surface increase the hydrophobicity of the surface, the hydrophobic zone may form (Kleber

et al., 2007). Amphiphilic moieties may orient themselves in such a way that the hydrophobic

tails interact with the molecules on the surface and the polar ends point outwards to shield the

10

hydrophobic area from water molecules (Kleber et al., 2007). This zone may take a longer

period of time to form but once sorbed should be relatively stable (Kleber et al., 2007).

Finally, molecules which are retained in the outer region form the kinetic zone (Kleber et al.,

2007). These molecules are loosely retained and are therefore in rapid exchange with the

solution (Kleber et al., 2007). This model highlights the importance of OM-OM interactions

and not simply OM-mineral interactions for the sorption and preservation of OM.

1.7. Sorption mechanisms

Multiple mechanisms have been identified through which OM can interact with mineral

surfaces and contaminants can interact with soil components. These mechanisms include

hydrophobic effects and van der Waals forces, cation or water bridging, cation or anion

exchange, ligand exchange and hydrogen bonding (Voice and Weber, 1983; Arnarson and

Keil, 2000; Sparks, 2003; von Lutzow et al., 2006). The dominant mechanism may depend

on the properties of both the sorbent (solid phase) and the sorbate (ion or molecule in

solution; Arnarson and Keil, 2000). Furthermore, certain binding mechanisms may be most

effective under different soil conditions such as pH or ionic strength as described below

(Arnarson and Keil, 2000).

1.7.1. Hydrophobic effects

Hydrophobic effects describe the combination of molecules being driven from water towards

a soil or mineral surface and the interaction between these molecules and the surface via van

der Waals forces (Voice and Weber, 1983; von Lutzow et al., 2006). These are most

important for contaminants or OM molecules which contain few functional groups and are

11

neutral at the pH of the soil (Arnarson and Keil, 2000). Water must adapt a highly ordered

structure which mimics crystalline ice to solvate a non-polar molecule (Voice and Weber,

1983). For this reason, hydrophobic effects are entropically driven as water is able to adapt a

less ordered arrangement after the OM or contaminant is driven from solution towards the

surface (Voice and Weber, 1983; Arnarson and Keil, 2000). The approach of the non-polar

molecules towards the surface causes electron density to shift creating temporary dipoles

which results in a relatively weak attraction between the non-polar molecules and the soil or

mineral (van der Waals interactions; Voice and Weber, 1983; Sparks, 2003; von Lutzow et

al., 2006). It has been suggested that sorption may occur through a combination of

hydrophobic effects and other mechanisms to increase the strength of the interaction

(Simpson et al., 2006; Kang and Xing, 2007). For example, long chain fatty acids may first

interact via the carboxylate group following which the hydrophobic tail may be able to better

displace water molecules from the surface and sorb through van der Waals interactions

(Simpson et al., 2006; Kang and Xing, 2007).

1.7.2. Cation and water bridging

Cation and water bridging are mechanisms which occur through coulombic attraction and

allow a molecule with a negatively charged functional group to sorb to a negatively charged

surface (Arnarson and Keil, 2000; von Lutzow et al., 2006). With the cation bridging

mechanism, cations on the surface reduce electrostatic repulsion and form an inner-sphere

complex with the surface and ligand, therefore acting as a bridge between the two (Sparks,

2003; von Lutzow et al., 2006; Sposito, 2008). If a water molecule inserts itself between the

cation and the functional group, a less tightly bound outer-sphere complex is formed and the

12

mechanism is termed water bridging (Arnarson and Keil, 2000; Sparks, 2003). The

importance of these two mechanisms is influenced by the valence of the cations present in

solution since polyvalent ions are able to form better bridges than monovalent cations

(Arnarson and Keil, 2000; Feng et al., 2005; von Lutzow et al., 2006; Polubesova et al.,

2008). For example, humic acid sorption to mineral surfaces has been observed to increase in

the presence of Ca2+ compared to Na+ and Fe3+ compared to Cu2+ (Feng et al., 2005;

Polubesova et al., 2008). Similarly, increasing the ionic strength of the solution will favour

sorption through a cation or water bridging mechanism (Feng et al., 2005). For interactions

with mineral oxides or hydroxyl groups of OM, the relevance of the mechanisms is a

function of pH as surface hydroxyl groups will only be deprotonated under basic conditions

(Arnarson and Keil, 2000; Sposito, 2008). Conversely, minerals such as montmorillonite

have a more permanent negative charge due to extensive isomorphic substitution during

weathering which is where a cation of similar size but lower valence displaces another in the

crystal lattice (i.e. Al3+ is replaced by Mg2+ or Si4+ is replaced by Al3+; Sparks, 2003).

Therefore, sorption to these minerals may occur through cation or water bridging regardless

of pH and indeed it has been suggested that a large amount of OM is sorbed to

montmorillonite through a cation or water bridge (Wattel-Koekkoek et al., 2001).

1.7.3. Cation and anion exchange

Cation and anion exchange are additional mechanisms driven by coulombic attraction

(Arnarson and Keil, 2000; Sparks, 2003). For these mechanisms, a charged molecule

displaces an inorganic ion of the same charge from the mineral surface (Arnarson and Keil,

2000). For example, with cation exchange, the surface is negatively charged while the

13

sorbate in this case is positively charged (Arnarson and Keil, 2000). The organic cation then

displaces an inorganic cation and interacts directly with the surface (Arnarson and Keil,

2000). At acidic pH, surface hydroxyl groups may be positively charged and anionic

exchange becomes a viable mechanism (Arnarson and Keil, 2000; Sposito, 2008). If these

mechanisms are dominant then sorption should decrease with increasing ionic strength since

there will be more competition between inorganic and organic ions (Arnarson and Keil,

2000).

1.7.4. Ligand exchange

Ligand exchange involves the displacement of a surface hydroxyl group by a hydroxyl or

carboxyl group of the OM or contaminant (Arnarson and Keil, 2000; Sparks, 2003; von

Lutzow et al., 2006). This mechanism is expected to be dominant for OM-mineral sorption

when the soil contains metal oxides such as goethite which have a high density of surface

hydroxyl groups but may also occur at mineral edges of aluminosilicates (Arnarson and Keil,

2000). This mechanism is also highly pH dependent (von Lutzow et al., 2006; Sposito,

2008). Ligand exchange is much more favourable at acidic pH since the metal-oxygen bond

is weakened and surface hydroxyl groups remain protonated which makes them better

leaving groups (von Lutzow et al., 2006; Sposito, 2008). As multiple hydroxide ions can be

released, an increase in solution pH following sorption could indicate a ligand exchange

mechanism (Chorover and Amistadi, 2001). Similarly, ligand exchange is exothermic so if

this mechanism is dominant, sorption should decrease as temperature increases (Arnarson

and Keil, 2000; von Lutzow et al., 2006).

14

1.8. Objectives

For the reasons previously discussed, it is necessary to understand the factors that will

influence sorption of both dissolved OM and pharmaceuticals and personal care products that

are released to soil through irrigation with reclaimed wastewater. Sorption of OM may

reduce its mineralization and lower emissions of carbon dioxide while sorption of

pharmaceuticals and personal care products will reduce the risk of groundwater

contamination or uptake into plants (Semple et al., 2003; Ternes et al., 2007; McAllister and

Semple, 2010; Wu et al., 2010). Furthermore, sorption of OM may influence the mobility of

these contaminants in soil (Kan and Tomson, 1990; Totsche and Kögel-Knabner, 2004;

Müller et al., 2007; Chefetz and Xing, 2009). Therefore, this project consists of two parts

designed to gain a fundamental understanding of the interactions occurring between soil

minerals, OM and pharmaceuticals and personal care products. First, the sorption of

dissolved OM to mineral surfaces is analyzed and compared with the types of OM present at

the soil-water interface. As previous studies of OM sorption have focused on the sorption of

single OM samples to various mineral surfaces, in this case, the preferential sorption of OM

samples of differing composition is observed. This is done to determine whether the initial

composition of OM influences preferential sorption which is important as the composition of

OM inputs can vary based on vegetation, microbial activity and diversity, and other

environmental factors such as temperature, moisture and management practice (Baldock et

al., 1992; Guggenberger et al., 1995; Zech et al., 1997; Quideau et al., 2001). Second, the

sorption of certain pharmaceuticals and personal care products to soils and mineral samples

is studied. Soils with differing organic carbon content and composition are used as sorbents

to evaluate which soil characteristics govern interactions with these contaminants: mineral

15

content, organic carbon content or OM composition. Sorption of different pharmaceuticals

and personal care products is observed to evaluate how the chemical and physical properties

of these contaminants influence sorption.

The overall objectives of this research are to:

1. Determine which components of three different dissolved OM samples are

preferentially sorbed to clay surfaces using nuclear magnetic resonance (NMR)

spectroscopy

2. Compare OM sorbed to clay surfaces with OM present at the soil-water interface

3. Quantify the sorption of certain pharmaceutical and personal care products to clay

surfaces and soil samples of varying organic carbon content and composition to

determine which soil characteristics influence sorption

4. Study the sorption of contaminants with different chemical and physical properties to

determine the role of these properties on sorption to soil

This thesis contains two subsequent chapters and a short summary/synthesis. Chapter two

describes the sorption of three dissolved OM samples of varying composition to the clay

minerals kaolinite and montmorillonite. In addition, this chapter presents 1H high resolution-

magic angle spinning NMR spectra of whole soils to allow comparison of OM components

present at the soil-water interface to the components sorbed to mineral surfaces. Chapter

three focuses on the sorption of four contaminants, carbamazepine, sulfamethoxazole, 17β-

estradiol and phenanthrene to soils and clays. 13C cross polarization-magic angle spinning

16

NMR spectra of the soils are included as well as relationships between the soil OM

composition and sorption affinity. Sorption isotherms and relationships between distribution

coefficients (Kd) and the fraction of organic carbon (foc) in each soil are also included as

appendices. This study will improve our understanding of the interactions of wastewater

constituents with soil and assist in making informed decisions regarding the further use of

reclaimed water for irrigation.

17

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1.10. Figures

Figure 1.1: Possible fates of contaminants introduced to the soil environment. Reprinted from: Agriculture, Ecosystems & Environment, Vol 120, Müller, K., Magesan, G.N., Bolan, N.S., A critical review of the influence of effluent irrigation on the fate of pesticides in soil., 93-116, Copyright (2007), with permission from Elsevier

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Figure 1.2: Zonal model of organo-mineral interactions. With kind permission from Springer Science + Business Media: Biogeochemistry, A conceptual model of organo-mineral interactions in soils: Self-assembly of organic molecular fragments into zonal structures on mineral surfaces, Vol 85, 2007, 9-24, Kleber, M., Sollins, P., Sutton, R., Figure 2.

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Chapter 2: Nuclear Magnetic Resonance (NMR) Analysis of Soil Organic Matter at the Solid-Water Interface 1

2.1. Introduction

The sorptive preservation of organic matter (OM) on mineral surfaces is hypothesized to

govern OM dynamics in soils and sediments (Keil et al., 1994; Kaiser and Guggenberger,

2000). For example, the mineralization of dissolved OM has been observed to decrease upon

sorption to mineral surfaces which may be a result of either chemical or physical stabilization

(Kalbitz et al., 2005; Mikutta et al., 2007). As soil is one of the largest sinks for organic

carbon, with approximately 1600 Pg of carbon stored in the top 100 cm (Eswaran et al.,

1993), organo-clay interactions which may aid in protection from mineralization are

important for global biogeochemical cycling especially in a changing world (Davidson and

Janssens, 2006; Smith et al., 2008). OM sorption to clay surfaces also alters mineral surface

properties such as surface charge density or hydrophobicity (Murphy et al., 1990; Day et al.,

1994). Furthermore, the sorption of environmentally persistent organic contaminants on wet

mineral surfaces is typically low however, contaminant sorption has been observed to

increase after coating with dissolved OM or humic substances (Murphy et al., 1990; Feng et

al., 2006). Consequently, identifying the precise mechanisms involved in OM-mineral

interactions is important for understanding OM biogeochemistry in soils as well as its role in

contaminant transport.

1 Manuscript submitted to Organic Geochemistry. Stephanie C. Hofley and Myrna J. Simpson wrote the manuscript and received comments from Andre J. Simpson, Denis Courtier-Murias and Ronald Soong. Stephanie C. Hofley isolated the dissolved OM and prepared the organo-clay complexes. Stephanie C. Hofley, Myrna J. Simpson, Andre J. Simpson, Denis Courtier-Murias, Ronald Soong and David J. McNally acquired and analyzed the NMR spectra.

26

The sorption capacity of soils is limited which leads to competition for sorption sites and

preferential sorption of specific OM components (Kalbitz et al., 2000). Previous studies have

examined the sorption of OM and have determined that the characteristics of the

preferentially sorbed components are dependent on the type of mineral surface (Chorover and

Amistadi, 2001; Feng et al., 2005; Wang and Xing, 2005; Ghosh et al., 2009). For example,

studies have observed the preferential sorption of aromatic and carboxylic fractions on

goethite (Chorover and Amistadi, 2001; Ghosh et al., 2009). Other research has observed that

kaolinite typically sorbs more non-polar aliphatic compounds and carbohydrates (Feng et al.,

2005; Wang and Xing, 2005; Ghosh et al., 2009) while aliphatic, proteins and aromatic

components were selectively sorbed onto montmorillonite surfaces (Chorover and Amistadi,

2001; Feng et al., 2005; Wang and Xing, 2005; Ghosh et al., 2009). In addition, solution

conditions, such as pH, ionic strength, and solution cation, as well as the dominant

exchangeable cation on the mineral surface also play a role in the preferential sorption of OM

to mineral surfaces (Namjesnik-Dejanovic et al., 2000; Feng et al., 2005; Majzik and

Tombacz, 2007; Polubesova et al., 2008). These aforementioned studies have used a variety

of methods to investigate the nature of OM-clay interactions such as UV-Vis spectroscopy

(Namjesnik-Dejanovic et al., 2000; Chorover and Amistadi, 2001; Wang and Xing, 2005;

Majzik and Tombacz, 2007), FTIR spectroscopy (Chorover and Amistadi, 2001; Polubesova

et al., 2008), High Performance Size Exclusion Chromatography (Namjesnik-Dejanovic et

al., 2000; Chorover and Amistadi, 2001), DRIFT spectroscopy (Ghosh et al., 2009) and

solid-state 13C NMR (Wang and Xing, 2005; Ghosh et al., 2009). Recently, advanced NMR

methods have been used to characterize organo-clay complexes in detail and determine

which OM components are preferentially adsorbed (Feng et al., 2005; Simpson et al., 2006).

27

However, these advanced NMR studies have not yet been used to examine if varying OM

composition prior to sorption imparts any variation on the resulting organo-clay complex

structure. This is a key aspect to consider as the structural characteristics of dissolved OM

can differ with vegetation, microbial activity and diversity, and other environmental factors

such as temperature, moisture and management practice (Baldock et al., 1992; Guggenberger

et al., 1995; Zech et al., 1997; Quideau et al., 2001; Kögel-Knabner, 2002; Ohno et al.,

2007). One study by Alekseeva et al. (2010) observed predominant sorption of alkyl

components from three different humic acids by solid-state 13C NMR spectroscopy.

However, two of these humic acid samples had similar composition while the third was

predominantly composed of aliphatic OM; thus competition between different OM structures

(aliphatic versus aromatic versus peptides) may not have been detected.

The objective of this study was to examine the role of varying dissolved OM composition on

the preferential sorption of specific OM compounds to mineral surfaces using NMR

techniques that provide molecular-level detail on OM structure before and after sorption. The

sorption of three different dissolved OM samples that were isolated from Peat humic acid

(PHA), a forest soil and Leonardite humic acid (LHA) to Ca2+ enriched kaolinite and

montmorillonite was studied using 1H High-Resolution Magic-Angle-Spinning (HR-MAS)

NMR which enables the analysis of semi-solid phase compounds that are in contact with the

NMR solvent (Simpson et al., 2006). 1H HR-MAS NMR has been used successfully in the

past to study dissolved OM sorption to clay mineral surfaces (Feng et al., 2005; Simpson et

al., 2006). We also examined the structural composition of the dissolved OM samples prior

to sorption using solution-state 1H NMR. Lastly, to test the observed trends with organo-clay

28

complexes, we analyzed whole soils containing kaolinite and montmorillonite by 1H HR-

MAS NMR and solid-state 13C NMR. Solid-state 13C NMR was employed to analyze all

components in soil OM and used for comparison to solvent accessible OM observed using 1H

HR-MAS NMR (Simpson et al., 2011). The overall goal of this research was to test the

hypothesis of OM preferential sorption using both organo-clay complexes and whole soils

using a number of NMR techniques.

2.2. Materials and Methods

2.2.1. Organo-clay complex preparation and carbon analysis

The clay minerals kaolinite (KGa-1b Washington County, Georgia) and montmorillonite

(STx-1b Gonzales County, Texas) were purchased from The Clay Minerals Society’s Source

Clays Repository (West Layfayette, Indiana). Montmorillonite has a surface area (N2) of

83.79 ± 0.22 m2/g and a cation exchange capacity of 84.4 meq/100 g whereas kaolinite has a

surface area of 10.05 ± 0.02 m2/g and a cation exchange capacity of 2.0 meq/100 g (Van

Olphen and Fripiat, 1979). The clays were suspended in a 0.01 M solution of calcium nitrate

(Ca(NO3)2·4H2O; certified A.C.S., Fisher Chemicals) and shaken for an hour to replace any

exchangeable cations and saturate all surfaces with Ca2+. The mixture was then centrifuged

(3000 rpm, 30 min), the supernatant was decanted and the clays were rinsed with deionized

water. The whole procedure was repeated after which the clays were freeze-dried.

Leonardite humic acid standard and Pahokee Peat soil were purchased from the International

Humic Substances Society (St Paul, Minnesota). Peat humic acid was extracted from the

Pahokee Peat using the method described by Salloum et al. (2001). Briefly, the Pahokee Peat

29

was mixed with 0.1 M NaOH for approximately 20 hours. The humic acid was precipitated

from the solution by acidifying to pH 1 using 6 M HCl. The humic acid was mixed with three

cycles of 0.1 M HCl/0.3 M HF to remove clay minerals, rinsed repeatedly with deionized

water to remove chlorides and finally freeze-dried. Dissolved OM was isolated from an O

horizon of a mixed forest soil located on the campus of the University of Toronto at

Scarborough (Toronto, Canada). Approximately 20 g of soil was mixed with 200 mL of

deionized water for ~20 hours in 250 mL Nalgene® polyethylene centrifuge bottles. After

mixing, the bottles were centrifuged (4000 rpm, 30 min). The supernatant was collected and

filtered through a 0.22 μm Durapore® membrane filter (Millipore) and then freeze-dried. The

procedure was repeated until no additional soluble OM was extracted (i.e.: water phase

remained colourless).

The humic acids and forest soil-derived dissolved OM were dissolved in a 0.01 M solution of

Ca(NO3)2 and the solution pH was raised to approximately 10 using NaOH. After stirring for

an hour, HNO3 was used to readjust to pH 7 and the solution was filtered through a 0.22 μm

Durapore® membrane filter (Millipore). The filtered solution was then freeze-dried such that

dissolved OM isolates were obtained. Dissolved OM isolates (150 mg) were then re-

dissolved in 40 mL of a 0.01 M Ca(NO3)2 solution. HNO3 and NaOH were used in order to

adjust the solution to pH 7. This solution was mixed with 150 mg of homoionic kaolinite or

montmorillonite in a 45 mL Nalgene® polyethylene centrifuge tube. The tube was shaken on

an Eberbach 6010 shaker for 48 hours. The solution was centrifuged (5000 rpm, 60 min), the

supernatant was retained and the organo-clay complex was washed five times with deionized

30

water. The organo-clay complexes and the supernatants were freeze-dried and placed in the

oven at 40°C under vacuum in the presence of P2O5 for at least 24 hours.

The total carbon content of organo-clay complexes was determined using the LECO

combustion method (University of Guelph Laboratory Services, Guelph, ON). Carbon

contents were as follows: 0.50% for soil-derived dissolved OM-kaolinite; 0.71% for soil-

derived dissolved OM-montmorillonite; 0.22% for LHA-kaolinite; 0.31% for LHA-

montmorillonite; 0.56% for PHA-kaolinite; and 1.0% for PHA-montmorillonite. The carbon

data indicates that in general, montmorillonite clays sorbed greater quantities of dissolved

OM than kaolinite; likely due to its high surface area which is consistent with other studies

(Feng et al., 2005). Of the three types of dissolved OM studied, LHA sorbed the least and

PHA sorbed the most to mineral surfaces.

2.2.2. Soil samples and preparation

Two grassland soils from the Canadian Prairies were included in this study because they

contain prominent amounts of kaolinite and montmorillonite but have varying carbon

contents (Salloum et al., 2001). These two samples have also been previously studied using a

host of molecular-level organic geochemical techniques in our laboratory (Otto et al., 2005;

Otto and Simpson, 2005, 2006a, 2006b; Shunthirasingham and Simpson, 2006; Feng and

Simpson, 2007; Clemente et al., 2011). The first grassland soil (referred to as Northern

Grassland) was sampled from the University of Alberta Ellerslie Research Station. The

second grassland soil (referred to as Southern Grassland) was sampled from the Agriculture

and Agri-Food Canada Research station in Lethbridge, Alberta. Both samples were air-dried

31

after sampling and stored in glass jars in the dark. The Northern Grassland and Southern

Grassland soils have organic carbon contents of 2.8% and 5.0% respectively (Clemente et al.,

2011). For contrast to mineral soils, a standard Peat soil was analyzed (Pahokee Peat;

purchased from the International Humic Substances Society; St Paul, Minnesota). The humin

fraction of the Northern Grassland soil was included for comparison because humin

represents the fraction of OM that is tightly held by mineral surfaces (Simpson and Johnson,

2006). Humin extraction was performed at outlined in Simpson and Johnson (2006).

Prior to solid-state 13C Cross Polarization Magic Angle Spinning (CP-MAS) NMR analysis,

the two grassland soils and humin samples were repeatedly treated with 0.3 M HF to

concentrate OM (Rumpel et al., 2006). After HF treatment, all samples were rinsed five times

with deionized water to remove excess HF and then freeze dried. The Peat sample was not

HF treated for any NMR analyses. Soil and humin samples were not pre-treated prior to 1H

HR-MAS NMR analysis. Chemical shifts were assigned based on previously published data

(Malcolm, 1989).

2.2.3. NMR analysis of dissolved OM and organo-clay complexes

Dissolved OM samples prior to sorption were prepared by dissolving ~25 mg in 1 mL of

deuterium oxide (D2O; 99.9% purity, Cambridge Isotope Laboratories) and were transferred

into 5 mm High Throughputplus NMR tubes (Norell Inc.; NJ, USA). 1H solution-state spectra

of the samples were acquired with a Bruker Avance III 500 MHz spectrometer equipped with

an 1H-19F-15N-13C 5 mm broadband Quadruple Inverse (QXI) probe fitted with an actively

shielded Z gradient (Bruker BioSpin, Rheinstetten, Germany). Experiments were performed

32

with 256 scans, a recycle delay of 3 seconds and 64 K time domain points at a temperature of

25°C. Water suppression was carried out by Presaturation Utilizing Relaxation Gradients and

Echoes (PURGE; Simpson and Brown, 2005). Spectra were apodized by multiplication with

an exponential decay corresponding to 1 Hz line broadening in the transformed spectrum and

a zero filling factor of 2. Spectra were externally referenced to DSS (δ = 0.00 ppm). 1H

solution-state NMR spectra were integrated using AMIX software (v. 3.9.7; Bruker BioSpin)

and the relative percentage of each region were calculated by dividing the area by the total

proton signal (0-8 ppm minus 0.1 ppm near the water peak). The Bruker Biofluid Reference

Compound Database (version 2.0.3, Bruker BioSpin) was used for the identification of some

of the more resolved components of the forest soil-derived dissolved OM spectrum.

Organo-clay complexes were analyzed using 1H HR-MAS NMR spectroscopy. The organo-

clay complexes (40 mg) were swollen with 60 μL of deuterated dimethyl sulfoxide (DMSO-

d6; 99.9% purity, Cambridge Isotope Laboratories) or D2O (99.9% purity, Cambridge Isotope

Laboratories). Slurries were prepared in a 4 mm zirconium rotor and then sealed with a Kel-F

insert and rotor cap. Spectra were acquired with a Bruker Avance III 500 MHz spectrometer

equipped with a 4 mm 1H-13C-15N HR-MAS probe fitted with an actively shielded magic

angle gradient (Bruker BioSpin, Rheinstetten, Germany). Experiments were performed at a

spinning speed of 6666 Hz. Experiments were performed with PURGE water suppression

(Simpson and Brown, 2005), 512 scans, a recycle delay of 2 seconds and 16 K time domain

points at a temperature of 25°C. Spectra were processed with a zero filling factor of 2 and by

multiplication with an exponential decay corresponding to 2 Hz line broadening in the

transformed spectrum. Spectra of complexes swollen with D2O were externally calibrated to

33

the methyl group of DSS (δ = 0.00 ppm). Spectra of complexes swollen with DMSO-d6 were

calibrated using the solvent residual peak at 2.50 ppm. Spectral interpretations of OM

structures are based on previous work by Simpson et al. (2001, 2006) and Deshmukh et al.

(2005).

2.2.4. NMR analysis of soils and humin

Solid-state CP-MAS 13C NMR spectra were acquired on a Bruker Avance 300 MHz NMR

spectrometer equipped with a 4 mm H-X MAS probe. Samples were prepared by packing

approximately 100 mg into a 4 mm zirconium rotor which was then closed with a Kel-F cap.

The acquisition parameters were as follows: MAS spinning rate of 13 kHz, CP contact time

of 1 ms, and a 1 second recycle delay. NMR spectra were processed using a zero filling

factor of 2 and line broadening of 50 Hz. Chemical shifts were calibrated against that of

glycine as an external standard.

Untreated soil and humin samples were dried under vacuum over P2O5 to remove excess

water prior to analysis by 1H HR-MAS NMR. Samples (~30 mg) were placed in a 4 mm

zirconium oxide rotor and 60 µl of DMSO-d6 or D2O was added as a swelling solvent. After

homogenization of the sample using a stainless steel mixing rod, the rotor was sealed using a

Kel-F insert and rotor cap. 1H HR-MAS NMR spectra were acquired using a Bruker Avance

III 500 MHz spectrometer equipped with a 4 mm triply tuned 1H-13C-15N HR-MAS probe

fitted with an actively shielded magic angle gradient at a spinning speed of 6666 Hz. PURGE

was employed for water suppression (Simpson and Brown, 2005). Spectra were acquired

34

with 256 scans, a recycle delay of 2 seconds and 8 K time domain points. Spectra were

processed with a zero-filling factor of 2 and a line broadening of 1 Hz.

2.3. Results and Discussion

2.3.1. NMR characterization of dissolved OM and organo-clay complexes

The solution-state 1H NMR spectra of the dissolved OM isolates prior to sorption are shown

in Fig. 2.1. Signals in the spectra are classified in four general categories: a) aliphatic

protons, b) protons adjacent to carbonyl groups, c) protons from carbohydrates and α protons

of amino acids and d) aromatic protons (Simpson et al., 2001, 2006; Deshmukh et al., 2005).

Resonances from 0-2 ppm and 2-3 ppm correspond to aliphatic protons (such as terminal

CH3 groups or polymethylene CH2) and protons adjacent to carbonyl groups respectively

(Simpson et al., 2001, 2006; Deshmukh et al., 2005). These are likely due to the presence of

cutin and cutan from plant cuticles and suberin which can be found in roots and bark (Kögel-

Knabner, 2002; Deshmukh et al., 2005). Other lipids and proteins found in plants and

microorganisms will also contribute to signals within these regions (Simpson et al., 2006,

2007b). Signals from 3-6 ppm are mainly attributed to carbohydrates such as cellulose and

hemicelluloses but may overlap with other signals such as those from α protons of amino

acids (Simpson et al., 2006) and methoxyl from lignin. Signals within the aromatic region

(between 6-8 ppm) may be attributed to aromatic groups found in lignin or aromatic

containing amino acids such as phenylalanine and tyrosine (Simpson et al., 2001; Kögel-

Knabner, 2002).

35

The 1H NMR spectrum of the forest soil-derived dissolved OM (Fig. 2.1a) contains a large

number of intense peaks at chemical shifts corresponding to aliphatic protons and protons

adjacent to carbonyl groups (0-3 ppm). Many overlapping peaks are also observed in the 3-6

ppm region of this sample where signals from carbohydrates and α protons of amino acids

resonate. For the soil-derived dissolved OM, the relative percent of the 3-6 ppm region is

37% (Table 2.1) which indicates that these compounds are a major component of this sample

relative to the others analyzed. However, the main contribution to the spectrum is signals

from aliphatic chains, fatty acids and esters (0-3 ppm). Only a few signals appear in the

aromatic region of the spectrum of the soil-derived dissolved OM and are consistent with the

resonances of phenylacetic acid, a common degradation intermediate of phenylalanine

(Fewson, 1988). For the LHA sample (Fig. 2.1b) a large percentage (Table 2.1) is attributed

to aliphatic components and protons adjacent to carbonyls (0-3 ppm). Fewer carbohydrate

and amino acid α proton peaks (3-6 ppm) are observed in the LHA 1H NMR spectrum in

comparison to the other two samples with a relative percent of only 17% for this region.

However, the LHA sample contains the largest relative percent of aromatics of all the

dissolved OM isolates which is evident from the large broad peak observed in the aromatic

region (6-8 ppm). Lastly, a number of 1H NMR signals are also visible in the aliphatic region

(0-2 ppm) for the PHA sample (Fig. 2.1c) but they are of a lower intensity than those of the

other two dissolved OM isolates. The spectrum also contains a number of resonances in the

2-3 ppm region indicating the presence of protons adjacent to carbonyl groups. These regions

make up a lower relative percent of the PHA sample (Table 2.1) than for the other two

samples. A large peak corresponding to carbohydrates and amino acid α protons is also

observed (3-6 ppm). The relative percentage of the carbohydrates and amino acids in the

36

sample is 48% which is higher than the relative percent of aliphatic components and in

contrast to the composition of the soil-derived dissolved OM and LHA. Only a few signals

appear in the aromatic region (6-8 ppm) of the spectrum of the PHA sample.

Overall, the forest soil-derived dissolved OM is composed mainly of aliphatic components

such as alkanes and fatty acids, relative to the other samples. Carbohydrates contribute a

smaller but still considerable amount to the soil-derived dissolved OM. This is consistent

with the polar and relatively fresh nature of dissolved OM isolated from forest O horizons

(Guggenberger and Zech, 1994; Simpson et al., 2008). The LHA sample contains aliphatic

components and has the highest aromaticity of the samples studied. The organic

geochemistry is characteristic of its origin because it is formed from diagenetically altered

lignite and is more weathered than soil OM (Chang and Berner, 1998). Conversely, the PHA

spectrum contains many signals from carbohydrates and α protons from amino acids. The

relative percent of aliphatic components in the PHA is comparable to the amount of

carbohydrates and amino acid α protons in the PHA but is still lower than the amount of

aliphatic compounds in the other two samples. Peat is formed in marshes where

decomposition is typically slower than in unsaturated soils (Gardiner and Miller, 2004)

which results in relatively fresh organic materials that are enriched in carbohydrates and

other labile components (Kalbitz et al., 2000; Kögel-Knabner, 2002). The three dissolved

OM isolates used in this study, PHA, LHA and forest soil-derived dissolved OM, have

varying compositions as related to their origin and these differences are observed by

solution-state 1H NMR prior to sorption to clay mineral surfaces.

37

1H HR-MAS NMR allows the application of solution-state NMR experiments to samples that

are not fully soluble such as soil and organo-clay samples (Simpson et al., 2001). When using

this technique, it is important to note that only those species that are in contact with solvent

are visible (Simpson et al., 2011). Pure solids are not observed and thus, the use of solvents

with varying properties such as D2O and DMSO-d6 facilitates the analysis of OM

components at the solid-water interface as well as components that may be buried beneath

the OM at the solid-water interface. For example, D2O is used to mimic environmental

conditions and DMSO-d6, which is more penetrating, can show OM components that are

buried or protected (Simpson et al., 2001). Although the initial compositions of the dissolved

OM samples varied (Fig. 2.1), the 1H HR-MAS spectra of the organo-clay complexes (Figs.

2.2 and 2.3) show that similar components of each sample were preferentially sorbed to

either kaolinite or montmorillonite mineral surfaces regardless of dissolved OM composition

before sorption.

The 1H HR-MAS NMR spectra of the organo-kaolinite complexes are displayed in Fig. 2.2.

The spectra of the organo-clay complexes swollen by D2O (Fig. 2.2a) show components that

exist at the solid-water interface (Simpson et al., 2001) and peaks labeled Si are from a

natural silicate species (Simpson et al., 2007b). The 1H HR-MAS NMR spectra reveal that

the components which sorbed to kaolinite surfaces are primarily aliphatic in nature and do

not vary with dissolved OM composition. For example, the 0.9 ppm and 1.05 ppm 1H

resonances have been assigned to terminal CH3 protons and main chain CH2 protons

respectively and correspond to long-chain aliphatic compounds found in plant cuticles or

microbial lipids (Simpson et al., 2001, 2007a). The sorption of compounds containing

38

carboxylic acids and esters, such as those derived from cutin, is highlighted by signals from

2-3 ppm (Deshmukh et al., 2005). However, the weak signals observed between ~7.5-8.5

ppm also suggest a small amount of sorption of aromatic compounds (Simpson et al., 2001).

Similarly, weak signals between ~3-6 ppm indicate a small amount of sorption of

carbohydrates and/or amino acids.

1H HR-MAS NMR spectra of organo-kaolinite complexes were also acquired in DMSO-d6

(Fig. 2.2b) and were used to complement the results observed in D2O. Since DMSO-d6 is a

more penetrating solvent (relative to D2O) it is capable of breaking H bonds and penetrates

into hydrophobic domains thus exposing components which are more tightly bound to the

surface (Simpson et al., 2001). Even with this more penetrating solvent, the signals observed

are still mainly those from aliphatic components (Fig. 2.2b). However, some further

information about the nature of the sorbed aliphatic components can be obtained by

examining the resonances at 0.8 ppm and 1.2 ppm which are assigned to terminal CH3

protons and main chain CH2 protons respectively (Simpson et al., 2006). The signal arising

from the main chain CH2 protons is of a higher intensity than that of the terminal CH3

protons and suggests that most of the compounds sorbed to kaolinite are indeed from long

chain aliphatic compounds such as those found in plant cuticles or microbial lipids

(Deshmukh et al., 2005; Simpson et al., 2007a). The resonances between 2-3 ppm are from

protons adjacent to carbonyl groups and indicate that the aliphatic components sorbed are not

simply n-alkanes and n-alkenes but include fatty acids and esters (Deshmukh et al., 2005;

Simpson et al., 2006). Similarly to the D2O swollen spectra, in the DMSO-d6 swollen spectra

39

low intensity signals are observed which indicate the presence of amino acids and

carbohydrates (3-6 ppm) as well as aromatics (6-8 ppm).

The 1H HR-MAS NMR spectra of the organo-montmorillonite complexes (Fig. 2.3) show

similar features as the organo-kaolinite complexes. The largest signals are from main chain

CH2 and CH3 protons followed by those from protons α to carbonyl groups. In the DMSO-d6

swollen 1H HR-MAS NMR spectra, the intensity of the main chain CH2 resonance is

comparable to that of the terminal CH3 resonance. This ratio of CH2 to CH3 intensity

suggests that the compounds sorbed to the montmorillonite surface include a mixture of

chain lengths and suggests that proteins as well as cutin-derived waxes have been sorbed

because short aliphatic side chains from amino acids such as valine and leucine will

contribute a greater number of CH3 groups than CH2 groups (Feng et al., 2005; Simpson et

al., 2007b). In the D2O swollen 1H HR-MAS NMR spectra, the resonance at ~3.7 ppm may

be assigned to either amino acid α protons or carbohydrate protons. In the DMSO-d6 swollen

1H HR-MAS NMR spectra the intensity of these signals increase slightly which also supports

the finding that some amino acids/peptides sorbed to montmorillonite. The aromatic region

of the D2O swollen 1H HR-MAS NMR spectra is relatively free of resonances. In the DMSO-

d6 swollen 1H HR-MAS NMR spectra, the intensity of the aromatic region is small in

comparison to the aliphatic region however; a broad resonance is observed which may be

from lignin or aromatic side chains of proteins. Because these signals are not observed in the

D2O spectra, it suggests that aromatic structures are buried and more tightly bound to

montmorillonite surfaces.

40

While the primary components sorbed to kaolinite and montmorillonite are aliphatic in

nature, there are observable differences between the organo-kaolinite and organo-

montmorillonite complexes. For example, kaolinite preferentially sorbed long chain aliphatic

compounds (likely derived from plant cuticles) whereas montmorillonite preferentially

sorbed aliphatic compounds of various chain lengths including cutin-derived waxes and

peptides. The 1H HR-MAS NMR spectra of both the kaolinite and montmorillonite

complexes contain signals between 3-6 ppm attributable to either carbohydrates or α protons

of amino acids. Further differences are found in the aromatic region. For the kaolinite

complexes, distinct resonances are observable both in D2O and DMSO-d6 whereas, for

montmorillonite, only broad aromatic signals are detected in the DMSO-d6 spectra. This

suggests that the aromatic moieties on montmorillonite are either more tightly bound to the

surface or are attributed to larger molecules whereas those on kaolinite are more mobile

(Simpson et al., 2001, 2006). The results obtained here suggest that preferential sorption is

primarily controlled by the type of minerals present in soil and not the characteristics of the

dissolved OM.

Sorption is believed to be controlled by the different mechanisms responsible for the binding

of OM to different mineral types (Chorover and Amistadi, 2001; Wattel-Koekkoek et al.,

2001). For example, montmorillonite has a higher cation exchange capacity than kaolinite

and as such, cation-bridging is believed to be a more active mechanism for sorption to

montmorillonite (Van Olphen and Fripiat, 1979; Wattel-Koekkoek et al., 2001). The

observation that mineral type determines preferential sorption is supported by Ghosh et al.

(2009) who studied sorption of Peat humic acid on kaolinite, montmorillonite and goethite

41

(pH of 5 and Na+ as the dominant cation). Non-polar aliphatic compounds and carbohydrates

were preferentially sorbed on kaolinite and paraffinic fractions were sorbed on

montmorillonite whereas compounds containing carboxylic functional groups were

preferentially sorbed on goethite. Similar trends in the relative intensities of the CH2 and CH3

groups as those observed in this study were detected by Feng et al. (2005) who also

examined the sorption of Peat humic acid on kaolinite and montmorillonite using 1H HR-

MAS NMR. However, it should be noted that the results of Feng et al. (2005) were obtained

using a pH of 7 but with Na+ as the dominant cation. As in this study, those that have

employed Ca2+ enriched clays likewise observed the preferential sorption of aliphatic

compounds on kaolinite and montmorillonite. For example, Chorover and Amistadi (2001)

saw no change in molar absorptivity of a natural OM solution (pH of 4) following mixing

with montmorillonite which suggests the selective sorption of predominantly aliphatic

components. Wang and Xing (2005) also observed the sorption of aliphatic fractions of a

humic acid on Ca2+ enriched kaolinite and montmorillonite by solid-state 13C NMR at pH 5.

Finally, Alekseeva et al. (2010) observed that organo-montmorillonite complexes showed

relatively lower intensity aromatic peaks than the original humic solutions (pH of 7)

indicating that alkyl groups and oxygen containing groups were preferentially sorbed.

Recently, Clemente et al. (2011) separated grassland soils into sand-, silt- and clay-sized

fractions to examine the structure of the OM associated with each fraction. Aliphatic and

polymethylene components were observed to accumulate in the clay fraction whereas more

aromatic components were observed in the silt- and sand-sized fractions. This suggests that

the organo-clay complexes prepared in this study do indeed mimic the processes occurring in

the environment which was tested further with the analysis of whole soils.

42

2.3.2. NMR characterization of soils and humin

The solid-state 13C CP-MAS and the 1H HR-MAS NMR spectra of soil samples are shown in

Figs. 2.4-2.6 and the NMR spectra of soil humin are displayed in Fig. 2.7. The solid-state 13C

CP-MAS NMR spectra for all samples show the typical range of components observed in soil

OM. Alkyl carbon (0-50 ppm) is prominent in all samples and is generally attributed to

lipids, waxes and polymethylene chain carbon from plant cuticles (Malcolm, 1989). The O-

alkyl region contains resonances from carbohydrates and methoxyl groups found in lignin

and peptides (Malcolm, 1989). Aromatic signatures (110-165 ppm) stem from lignin, black

carbon and aromatic amino acids (Malcolm, 1989). Lastly, the carboxylic and carbonyl

region (165 -210 ppm) show a clear peak at ~170 ppm which is attributed to COOH groups

that are found in fatty acids and peptides (Malcolm, 1989). The inclusion of the solid-state

13C CP-MAS NMR data in this study is for comparison to the 1H HR-MAS NMR data and it

is important to note that all samples show prominent signals that can be attributed to aromatic

compounds found in soil OM (Figs. 2.4-2.7).

As discussed previously, 1H HR-MAS NMR allows the application of solution-state NMR

experiments to samples that are not fully soluble and which also contain pure solid domains.

Only those species that are in contact with solvent are observed and the use of different

swelling solvents allows for a comparison of OM structures that are visible at different

interfaces (Simpson et al., 2001; Feng et al., 2005, 2006). In this study, D2O was employed

as one of the swelling solvents to examine which OM components are in direct contact with

soil water. The 1H HR-MAS NMR spectra for all four samples in D2O were similar in that

nearly identical signals were observed albeit in different amounts (Figs. 2.4-2.7). In D2O, we

43

would expect to observe signals from compounds that are found primarily at the solid-

aqueous interface and are therefore readily accessible to water (Simpson et al., 2001). Signals

at approximately 0.8, 1.4, 2.1 and 2.2 ppm can be attributed to aliphatic waxy/lipid

compounds such as CH3-CH2-CH2-CH2-R, CH3-CH2-R-CH2-CH2-CH2-OH, R-CH2-CH2-

CO2H and R-CH2-CH2-CO2H, respectively (Simpson et al., 2001; Deshmukh et al., 2005).

Several overlapped signals observed at 3.5-4.3 ppm indicated the presence of different

carbohydrates, amino acids, aromatic methoxyl groups and CH2 units adjacent to ether and

ester groups (Simpson et al., 2001; Deshmukh et al., 2005). A signal was observed which

overlapped with water at approximately 4.8 ppm and was attributed to an ester of a mid-chain

hydroxyl compound found in the cutin/cutan of plants (Deshmukh et al., 2005). In each

sample, a sharp formic acid signal was observed at approximately 8.5 ppm.

1H HR-MAS NMR analysis using DMSO-d6 as the swelling solvent revealed signals

originating from a wider range of components (Figs. 2.4-2.7). DMSO-d6 is a more

penetrating solvent than D2O because it can break hydrogen bonds and in 1H HR-MAS

NMR reveals additional information about OM components that may be buried and

inaccessible to D2O (Simpson et al., 2001). In this study, many signals can be attributed to

cutinaceous and cutanaceous compounds from the leaves of plants or microbially derived

lipids (Deshmukh et al., 2005; Simpson et al., 2007a). The 1H NMR spectrum can be sub-

divided based on the types of compounds that resonate in different regions. For example,

from δ=0-1 ppm, one would expect to find terminal CH3 groups; 1-2 ppm main chain

methylene CH2 units from lipids, waxes/cuticle/lipids; 2-3 ppm substituted methylenes and

methines α to a functionality in hydrocarbons; 3-4 ppm carbohydrates; 4-6 ppm the

44

anomeric protons of carbohydrates, esters, double bonds, tannins and; 6-8 ppm mainly

aromatic protons (Kelleher et al., 2006).

It is important to note that the aromatic protons were not observed when the soil samples

were swollen in D2O which implies that these constituents are not prevalent at the soil-water

interface and confirms the results observed with the organo-clay complexes. Further

evidence for this is based on the clear aromatic signal in all of the solid-state 13C NMR

spectra. In particular, signals from aromatic compounds at approximately 7 ppm were only

visible in samples after swelling in DMSO-d6. Compared to D2O, DMSO-d6 is a highly

penetrating solvent that can enter both hydrophobic and hydrophilic domains (Simpson et

al., 2001). In addition, samples swollen in DMSO-d6 have enhanced molecular mobility and

when combined with MAS (that minimizes the effects of chemical shift anisotropy, dipole-

dipole interactions and magnetic susceptibility line broadening), less mobile soil

components become visible (Keifer et al., 1996; Millis et al., 1997; Stark et al., 2000; Fang

et al., 2001). Swelling in DMSO-d6 was also shown to disrupt soil aggregates and provide

information on hydrophobic structures that are physically protected by the arrangement of

the OM under aqueous conditions (Hayes and Swift, 1978). Our results for soil and humin

samples swollen in DMSO-d6 are consistent with a study that showed aromatic moieties in

soil are protected within hydrophobic regions that are not accessible by water (Simpson et

al., 2001). These aromatic signals were not observed in D2O and indicate that these

compounds are not easily accessible at the soil-water interface. It can thus be hypothesized

that aromatic moieties in soil may exist within a hydrophobic environment (Piccolo et al.,

1996; Simpson et al., 2001), or within the layers of phyllosilicate clay minerals, for which

45

aromatic species are known to have high affinity (Jaynes and Vance, 1999; Simpson et al.,

2001). These environments would be protected from water, but would be accessible to a

more penetrating solvent such as DMSO-d6. These results are also consistent with

hypotheses developed from contaminant-OM studies which found that OM accessibility to

the contaminant is equally as important as the structure of the OM (Murphy et al., 1990;

Salloum et al., 2001; Feng et al., 2006; Simpson and Johnson, 2006; Bonin and Simpson,

2007). Our results also highlight the importance of OM-OM interactions which protect

moieties from water penetration and perhaps biodegradation. Consequently, both OM-clay

interactions and OM-OM interactions are important for the short-term and long-term fate of

OM in soil environments.

2.4. Conclusions

The use of 1H HR-MAS NMR with contrasting solvents revealed that kaolinite sorbed more

long chain aliphatic species while montmorillonite sorbed aliphatic compounds of various

chain lengths, including cutin–derived waxes and peptides. The types of compounds

observed on the clay surfaces were independent of initial OM composition which varied

considerably prior to sorption. These results with organo-clay complexes demonstrate that

preferential sorption is more influenced by mineral type and experimental conditions (pH,

dominant cation and ionic strength) than the structural characteristics of OM prior to

sorption. This observation is consistent with studies on soil humin structure which reported

that humin contains high concentrations of aliphatic components (Simpson and Johnson,

2006). The presence of these compounds in humin fractions suggests that they are persistent

and non-extractable.

46

The observations with organo-clay samples were confirmed using whole soil and a

corresponding soil humin sample. Aliphatic compounds, carbohydrates and amino acids were

prevalent at the soil-aqueous interface of soils swollen in D2O. Aromatic compounds were

detected in the 13C CP-MAS NMR spectra but were not visible by 1H HR-MAS NMR

spectroscopy until DMSO-d6, a more penetrating solvent than D2O, was used as a swelling

solvent. This suggests that aromatic compounds in whole soils are not available at the soil-

aqueous interface but instead exist in more hydrophobic domains. Therefore, aromatic

species are likely less available for interaction with contaminant species and will have less of

an influence on their transport through soils. Since these results were likewise found with a

Peat soil which is low in minerals, our study also highlights the importance of OM-OM

interactions which may play a role in the protection and preservation of specific OM

components.

47

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Baldock, J.A., Oades, J.M., Waters, A.G., Peng, X., Vassallo, A.M., Wilson, M.A., 1992. Aspects of the chemical-structure of soil organic materials as revealed by solid-state 13C NMR spectroscopy. Biogeochemistry 16, 1-42.

Bonin, J.L., Simpson, M.J., 2007. Variation in phenanthrene sorption coefficients with soil organic matter fractionation: The result of structure or conformation? Environmental Science & Technology 41, 153-159.

Chang, S., Berner, R.A., 1998. Humic substance formation via the oxidative weathering of coal. Environmental Science & Technology 32, 2883-2886.

Chorover, J., Amistadi, M.K., 2001. Reaction of forest floor organic matter at goethite, birnessite and smectite surfaces. Geochimica et Cosmochimica Acta 65, 95-109.

Clemente, J.S., Simpson, A.J., Simpson, M.J., 2011. Association of specific organic matter compounds in size fractions of soils under different environmental controls. Organic Geochemistry 42, 1169-1180.

Davidson, E.A., Janssens, I.A., 2006. Temperature sensitivity of soil carbon decomposition and feedbacks to climate change. Nature 440, 165-173.

Day, G.M., Hart, B.T., Mckelvie, I.D., Beckett, R., 1994. Adsorption of natural organic matter onto goethite. Colloids and Surfaces A-Physicochemical and Engineering Aspects 89, 1-13.

Deshmukh, A.P., Simpson, A.J., Hadad, C.M., Hatcher, P.G., 2005. Insights into the structure of cutin and cutan from Agave americana leaf cuticle using HRMAS NMR spectroscopy. Organic Geochemistry 36, 1072-1085.

Eswaran, H., Vandenberg, E., Reich, P., 1993. Organic carbon in soils of the world. Soil Science Society of America Journal 57, 192-194.

Fang, X., Qiu, F., Yan, B., Wang, H., Mort, A.J., Stark, R.E., 2001. NMR studies of molecular structure in fruit cuticle polyesters. Phytochemistry 57, 1035-1042.

Feng, X., Simpson, M.J., 2007. The distribution and degradation of biomarkers in Alberta grassland soil profiles. Organic Geochemistry 38, 1558-1570.

Feng, X.J., Simpson, A.J., Simpson, M.J., 2005. Chemical and mineralogical controls on humic acid sorption to clay mineral surfaces. Organic Geochemistry 36, 1553-1566.

Feng, X.J., Simpson, A.J., Simpson, M.J., 2006. Investigating the role of mineral-bound humic acid in phenanthrene sorption. Environmental Science & Technology 40, 3260-3266.

Fewson, C.A., 1988. Microbial metabolism of mandelate: A microcosm of diversity. FEMS Microbiology Reviews 54, 85-110.

Gardiner, D.T., Miller, R.W., 2004. Soils in our Environment, 10th Edition. Prentice Hall, New Jersey.

Ghosh, S., Wang, Z.Y., Kang, S., Bhowmik, P.C., Xing, B.S., 2009. Sorption and fractionation of a peat derived humic acid by kaolinite, montmorillonite, and goethite. Pedosphere 19, 21-30.

48

Guggenberger, G., Zech, W., 1994. Dissolved organic carbon in forest floor leachates - simple degradation products or humic substances. Science of the Total Environment 152, 37-47.

Guggenberger, G., Zech, W., Haumaier, L., Christensen, B.T., 1995. Land-use effects on the composition of organic matter in particle-size separates of soils: II. CPMAS and solution 13C NMR analysis. European Journal of Soil Science 46, 147-158.

Hayes, M.H.B., Swift, R.S., 1978. The Chemistry of Soil Constituents. Wiley, New York. Jaynes, W.F., Vance, G.F., 1999. Sorption of benzene, toluene, ethylbenzene, and xylene

(BTEX) compounds by hectorite clays exchanged with aromatic organic cations. Clays and Clay Minerals 47, 358-365.

Kaiser, K., Guggenberger, G., 2000. The role of DOM sorption to mineral surfaces in the preservation of organic matter in soils. Organic Geochemistry 31, 711-725.

Kalbitz, K., Solinger, S., Park, J.H., Michalzik, B., Matzner, E., 2000. Controls on the dynamics of dissolved organic matter in soils: A review. Soil Science 165, 277-304.

Kalbitz, K., Schwesig, D., Rethemeyer, J., Matzner, E., 2005. Stabilization of dissolved organic matter by sorption to the mineral soil. Soil Biology & Biochemistry 37, 1319-1331.

Keifer, P.A., Baltusis, L., Rice, D.M., Tymiak, A.A., Shoolery, J.N., 1996. A comparison of NMR spectra obtained for solid-phase-synthesis resins using conventional high-resolution, magic-angle-spinning, and high-resolution magic-angle-spinning probes. Journal of Magnetic Resonance Series A 119, 65-75.

Keil, R.G., Montlucon, D.B., Prahl, F.G., Hedges, J.I., 1994. Sorptive preservation of labile organic matter in marine sediments. Nature 370, 549-552.

Kelleher, B.P., Simpson, M.J., Simpson, A.J., 2006. Assessing the fate and transformation of plant residues in the terrestrial environment using HR-MAS NMR spectroscopy. Geochimica et Cosmochimica Acta 70, 4080-4094.

Kögel-Knabner, I., 2002. The macromolecular organic composition of plant and microbial residues as inputs to soil organic matter. Soil Biology & Biochemistry 34, 139-162.

Majzik, A., Tombacz, E., 2007. Interaction between humic acid and montmorillonite in the presence of calcium ions I. Interfacial and aqueous phase equilibria: Adsorption and complexation. Organic Geochemistry 38, 1319-1329.

Malcolm, R.L., 1989. Applications of solid-state 13C NMR spectroscopy to geochemical studies of humic substances. In: Hayes, M.H.B., MacCarthy, P., Malcolm, R.L., Swift, R.S. (Eds.), Humic Substances II. In Search of Structure. John Wiley, New York, pp. 340-372.

Mikutta, R., Mikutta, C., Kalbitz, K., Scheel, T., Kaiser, K., Jahn, R., 2007. Biodegradation of forest floor organic matter bound to minerals via different binding mechanisms. Geochimica et Cosmochimica Acta 71, 2569-2590.

Millis, K.K., Maas, W.E., Cory, D.G., Singer, S., 1997. Gradient, high-resolution, magic-angle spinning nuclear magnetic resonance spectroscopy of human adipocyte tissue. Magnetic Resonance in Medicine 38, 399-403.

Murphy, E.M., Zachara, J.M., Smith, S.C., 1990. Influence of mineral-bound humic substances on the sorption of hydrophobic organic compounds. Environmental Science & Technology 24, 1507-1516.

49

Namjesnik-Dejanovic, K., Maurice, P., Aiken, G., Cabaniss, S., Chin, Y., Pullin, M., 2000. Adsorption and fractionation of a muck fulvic acid on kaolinite and goethite at pH 3.7, 6, and 8. Soil Science 165, 545-559.

Ohno, T., Chorover, J., Omoike, A., Hunt, J., 2007. Molecular weight and humification index as predictors of adsorption for plant- and manure-derived dissolved organic matter to goethite. European Journal of Soil Science 58, 125-132.

Otto, A., Shunthirasingham, C., Simpson, M.J., 2005. A comparison of plant and microbial biomarkers in grassland soils from the prairie ecozone of Canada. Organic Geochemistry 36, 425-448.

Otto, A., Simpson, M., 2005. Degradation and preservation of vascular plant-derived biomarkers in grassland and forest soils from western Canada. Biogeochemistry 74, 377-409.

Otto, A., Simpson, M.J., 2006a. Evaluation of CuO oxidation parameters for determining the source and stage of lignin degradation in soil. Biogeochemistry 80, 121-142.

Otto, A., Simpson, M., 2006b. Sources and composition of hydrolysable aliphatic lipids and phenols in soils from western Canada. Organic Geochemistry 37, 385-407.

Piccolo, A., Nardi, S., Concheri, G., 1996. Macromolecular changes of humic substances induced by interaction with organic acids. European Journal of Soil Science 47, 319-328.

Polubesova, T., Chen, Y., Navon, R., Chefetz, B., 2008. Interactions of hydrophobic fractions of dissolved organic matter with Fe3+- and Cu2+-montmorillonite. Environmental Science & Technology 42, 4797-4803.

Quideau, S.A., Chadwick, O.A., Benesi, A., Graham, R.C., Anderson, M.A., 2001. A direct link between forest vegetation type and soil organic matter composition. Geoderma 104, 41-60.

Rumpel, C., Rabia, N., Derenne, S., Quenea, K., Eusterhues, K., Kögel-Knabner, I., Mariotti, A., 2006. Alteration of soil organic matter following treatment with hydrofluoric acid (HF). Organic Geochemistry 37, 1437-1451.

Salloum, M.J., Dudas, M.J., McGill, W.B., 2001. Variation of 1-naphthol sorption with organic matter fractionation: The role of physical conformation. Organic Geochemistry 32, 709-719.

Shunthirasingham, C., Simpson, M.J., 2006. Investigation of bacterial hopanoid inputs to soils from western Canada. Applied Geochemistry 21, 964-976.

Simpson, A.J., Kingery, W.L., Shaw, D.R., Spraul, M., Humpfer, E., Dvortsak, P., 2001. The application of H-1 HR-MAS NMR spectroscopy for the study of structures and associations of organic components at the solid - aqueous interface of a whole soil. Environmental Science & Technology 35, 3321-3325.

Simpson, A.J., Brown, S.A., 2005. Purge NMR: Effective and easy solvent suppression. Journal of Magnetic Resonance 175, 340-346.

Simpson, A.J., Simpson, M.J., Kingery, W.L., Lefebvre, B.A., Moser, A., Williams, A.J., Kvasha, M., Kelleher, B.P., 2006. The application of 1H high-resolution magic-angle spinning NMR for the study of clay-organic associations in natural and synthetic complexes. Langmuir 22, 4498-4503.

Simpson, A.J., Simpson, M.J., Smith, E., Kelleher, B.P., 2007a. Microbially derived inputs to soil organic matter: Are current estimates too low? Environmental Science & Technology 41, 8070-8076.

50

Simpson, A.J., Song, G., Smith, E., Lam, B., Novotny, E.H., Hayes, M.H.B., 2007b. Unraveling the structural components of soil humin by use of solution-state nuclear magnetic resonance spectroscopy. Environmental Science & Technology 41, 876-883.

Simpson, A.J., McNally, D.J., Simpson, M.J., 2011. NMR spectroscopy in environmental research: From molecular interactions to global processes. Progress in Nuclear Magnetic Resonance Spectroscopy 58, 97-175.

Simpson, M.J., Johnson, P.C.E., 2006. Identification of mobile aliphatic sorptive domains in soil humin by solid-state 13C nuclear magnetic resonance. Environmental Toxicology and Chemistry 25, 52-57.

Simpson, M.J., Otto, A., Feng, X.J., 2008. Comparison of solid-state carbon-13 nuclear magnetic resonance and organic matter biomarkers for assessing soil organic matter degradation. Soil Science Society of America Journal 72, 268-276.

Smith, P., Fang, C., Dawson, J.J.C., Moncrieff, J.B., 2008. Impact of global warming on soil organic carbon. Advances in Agronomy, Vol 97 97, 1-43.

Stark, R.E., Yan, B., Ray, A.K., Chen, Z., Fang, X., Garbow, J.R., 2000. NMR studies of structure and dynamics in fruit cuticle polyesters. Solid State Nuclear Magnetic Resonance 16, 37-45.

Van Olphen, H., Fripiat, J.J., (Eds.), 1979. Data Handbook for Clay Minerals and Other Non-Metallic Materials, Pergamon Press, Oxford.

Wang, K.J., Xing, B.S., 2005. Structural and sorption characteristics of adsorbed humic acid on clay minerals. Journal of Environmental Quality 34, 342-349.

Wattel-Koekkoek, E.J.W., van Genuchten, P.P.L., Buurman, P., van Lagen, B., 2001. Amount and composition of clay-associated soil organic matter in a range of kaolinitic and smectitic soils. Geoderma 99, 27-49.

Zech, W., Senesi, N., Guggenberger, G., Kaiser, K., Lehmann, J., Miano, T.M., Miltner, A., Schroth, G., 1997. Factors controlling humification and mineralization of soil organic matter in the tropics. Geoderma 79, 117-161.

51

2.6. Tables

Table 2.1: Solution-state 1H NMR integration results for the dissolved OM isolates prior to sorption.

Relative Percentage of Total 1H NMR Signal (0-8 ppm) Sample Aliphatic Protons

(0-2 ppm) Protons α to a

Carbonyl (2-3 ppm)

Carbohydrates and Amino Acid

α Protons (3-6 ppm)

Aromatic Protons (6-8 ppm)

DOM 40 17 37 6 LHA 41 26 17 16 PHA 26 17 48 9

52

2.7. Figures

Figure 2.1: 1H solution-state NMR spectra in D2O of the dissolved OM isolates, a) forest soil-derived dissolved OM, b) Leonardite humic acid and c) Peat humic acid, prior to sorption to clay mineral surfaces.

8.5 8.0 7.5 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 ppm

aromatic compounds

c) PHA

b) LHA

a) DOM

c) PHA

b) LHA

a) DOM

amino acids & carbohydrates

water

aliphatic compounds

acetic acid

protons αto carbonyls

1H Chemical Shift (ppm)

53

Figure 2.2: 1H HR-MAS NMR spectra of organo-kaolinite complexes swollen in a) D2O and b) DMSO-d6. Enlargements (×16) of the aromatic regions are provided in the boxes above the spectra.

8.5 8.0 7.5 7.0 6.5 6

8.5 8.0 7.5 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 ppm

8.5 8.0 7.5 7.0 6.5 6.

PHA-kaolinite

LHA-kaolinite

8.5 8.0 7.5 7.0 6.5 6.

8.5 8.0 7.5 7.0 6.5 6.

b) 1H HR-MAS NMR Swollen in DMSO-d6

8.5 8.0 7.5 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 ppm

PHA-kaolinite

DOM-kaolinite

aromatic

CH3

acetic acid

protons αto carbonyl

main chain CH2

1H Chemical Shift (ppm)

water aromatic

CH3protons αto carbonylprotons αto carbonyl

DMSO

1H Chemical Shift (ppm)

amino acids & carbohydrates

LHA-kaolinite

DOM-kaolinite

a) 1H HR-MAS NMR Swollen in D2O

SiSi

wateramino acids & carbohydrates

x 16

x 16

x 16

x 16

x 16

x 16

8.5 8.0 7.5 7.0 6.5 6

8.5 8.0 7.5 7.0 6.5 6

main chain CH2

54

Figure 2.3: 1H HR-MAS NMR spectra of organo-montmorillonite complexes swollen in a) D2O and b) DMSO-d6. Enlargements (×16) of the aromatic regions are provided in the boxes above the spectra.

1H Chemical Shift (ppm)1H Chemical Shift (ppm)8.5 8.0 7.5 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 ppm 8.5 8.0 7.5 7.0 6.5 6.0 5.5 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0 0.5 ppm

DOM-montmorillonite

LHA-montmorillonite

PHA-montmorillonite

protons αto carbonyl

DMSO

amino acids & carbohydrates

protons αto carbonyl

water

LHA-montmorillonite

PHA-montmorillonite

a) 1H HR-MAS NMR Swollen in D2O

Si

Si

b) 1H HR-MAS NMR Swollen in DMSO-d6

acetic acid

DOM-montmorillonite

wateramino acids & carbohydrates

x 16

x 16

x 16

x 16

x 16

x 16

8.5 8.0 7.5 7.0 6.5 6.

8.5 8.0 7.5 7.0 6.5 6.

8.5 8.0 7.5 7.0 6.5 6.

5 8.0 7.5 7.0 6.5 6.0 5.

5 8.0 7.5 7.0 6.5 6.0 5.

5 8.0 7.5 7.0 6.5 6.0 5.

aromatic aromatic

CH3

main chain CH2

CH3

main chain CH2

55

Figure 2.4: Solid-state 13C NMR spectrum of the HF treated Northern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of the 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents.

23456789 ppm

in DMSO-d6

in D2O13C CP-MAS NMR of

whole soil treated with HF

13C Chemical Shift (ppm)

1H HR-MAS NMR

x 10

x 10

DM

SO

1H Chemical Shift (ppm)

56

Figure 2.5: Solid-state 13C NMR spectrum of the HF treated Southern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of the 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents.

23456789 ppm

x 8

x 8

in D2O

in DMSO-d6

H OH

O

DM

SO

200 150 100 50 0

in D2O13C CP-MAS NMR of whole soil treated with HF

13C Chemical Shift (ppm)

1H HR-MAS NMR

1H Chemical Shift (ppm)

in DMSO-d6

57

Figure 2.6: Solid-state 13C NMR spectrum and 1H HR-MAS NMR spectra of the untreated Peat soil (in D2O and DMSO-d6). Aromatic regions of 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents.

1H HR-MAS NMR @ 500MHz

in D2O

23456789 ppm

x 6

x 6

23456789 ppm

x 6x 6

x 6x 6

H OH

O

DM

SO

13C CP-MAS NMR of untreated Peat

13C Chemical Shift (ppm)

1H Chemical Shift (ppm)

in DMSO-d6050100150200

in D2O

1H HR-MAS NMR

58

Figure 2.7: Solid-state 13C NMR spectrum of HF treated humin isolated from the Northern Grassland Soil. 1H HR-MAS NMR spectra of the untreated soil (in D2O and DMSO-d6) are also shown. Aromatic regions of 1H HR-MAS NMR spectra are expanded to highlight results with the different solvents.

23456789 ppm

x 6

x 6

23456789 ppm

x 6x 6

x 6x 6

H OH

O

DM

SO

050100150200

Chemical Shift 050100150200

Chemical Shift

13C CP-MAS NMR of HF treated Humin

13C Chemical Shift (ppm)

1H Chemical Shift (ppm)

in DMSO-d6

in D2O

1H HR-MAS NMR

59

Chapter 3: Sorption of Carbamazepine, Sulfamethoxazole, 17β-Estradiol and Phenanthrene to Soils with Varying Organic Matter Composition

3.1. Introduction

The use of reclaimed wastewater for irrigation purposes has been considered to reduce the

demand for potable water (Kinney et al., 2006; Ternes et al., 2007; Tamtam et al., 2011).

However, there is concern about some of the constituents that are found in wastewater, namely

pharmaceuticals and personal care products, which are not entirely removed during the

wastewater treatment process (Carballa et al., 2004; Braga et al., 2005; Castiglioni et al., 2006;

Jelic et al., 2011; Zhang et al., 2011). Consequently, the application of wastewater to soil can

result in the concomitant introduction of these contaminants in low concentrations. The analysis

of soils irrigated with reclaimed wastewater detected various pharmaceuticals and personal care

products, even months after cessation of irrigation, which suggests that once applied to soils

these contaminants have the potential to persist (Kinney et al., 2006; Tamtam et al., 2011).

Therefore, a fundamental understanding of the interaction mechanisms of pharmaceuticals and

personal care products with soil components will assist in determining their fate in the soil

environment and will help elucidate whether these contaminants will accumulate in upper soil

layers, leach to groundwater, undergo degradation and/or be taken up by plants.

Investigation into the fate of pharmaceuticals and personal care products in the environment only

began recently such that the factors governing interactions between these contaminants and soil

are still not fully understood (Halling-Sørensen et al., 1998; Pan et al., 2009; Pignatello et al.,

2010). Previous studies have determined the sorption affinity and relative mobility of

pharmaceuticals and personal care products for soils and have related this to soil properties such

60

as organic carbon (OC) content and mineral content as well as the physicochemical properties of

the contaminants (Lee et al., 2003; Oppel et al., 2004; Yu et al., 2004; Drillia et al., 2005;

Williams et al., 2006; Chefetz et al., 2008; Sanders et al., 2008; Xu et al., 2009;

Karnjanapiboonwong et al., 2010). In addition, studies have investigated sorption of

pharmaceuticals and personal care products with respect to varying soil organic matter (OM)

characteristics (Yamamoto et al., 2003; Thiele-Bruhn et al., 2004; Bonin and Simpson, 2007a;

Sun et al., 2007; Hou et al., 2010). For example, Thiele-Bruhn et al. (2004) characterized soil

OM by pyrolysis mass spectrometry and observed a positive correlation between sulfonamide

sorption to soil and polar moieties. Similar conclusions that sulfamethoxazole could be

interacting with polar groups of sediment fractions were drawn by Hou et al. (2010). Other

studies on the sorption of endocrine disruptors to humic acids and dissolved OM surrogates have

suggested that an important sorption mechanism may be π-interactions with aromatic

components (Yamamoto et al., 2003; Sun et al., 2007; Hou et al., 2010). Conversely, Bonin and

Simpson (2007a) did not observe any clear trend between sorption of steroid estrogens to whole

soil and the OM characteristics as measured by solid-state 13C NMR. Therefore, further

investigation into the sorption of pharmaceutical and personal care products is required to

understand which soil characteristics control sorption and environmental mobility.

The sorption behaviour of polycyclic aromatic hydrocarbons has been extensively studied for

several decades and it is widely accepted that soil OM is the primary sorption domain for these

contaminants in soils, provided the OC content is greater than 0.1% (Schwarzenbach and

Westall, 1981; Murphy et al., 1990). However, the precise characteristics of soil OM that govern

sorption are still under investigation. For example, a number of studies have suggested that

61

sorption is proportional to the amount of aromatic carbon in soil OM (Chin et al., 1997; Xing,

1997; Chiou et al., 1998; Perminova et al., 1999). Xing (1997) observed an increase in the OC

normalized sorption coefficients (Koc) of naphthalene with increasing soil OM aromaticity.

Conversely, other studies proposed that alkyl carbon is a high affinity sorption domain for

polycyclic aromatic hydrocarbons (Chefetz et al., 2000; Mao et al., 2002; Salloum et al., 2002;

Wang et al., 2011). Salloum et al. (2002) observed that aliphatic-rich natural OM sorbed equal or

greater amounts of phenanthrene than aromatic-rich samples. Similarly, Chefetz et al. (2000)

observed higher Koc values for sorption of pyrene to cuticle and humin, two natural OM samples

with high aliphaticity, than to highly aromatic natural OM samples. A third group of studies have

found poor correlations between Koc values and both aromaticity and aliphaticity and have

proposed that these properties alone are not adequate predictors for sorption (Simpson et al.,

2003; Chen et al., 2005; Bonin and Simpson, 2007b; Chefetz and Xing, 2009). Instead, one

hypothesis that has emerged is that the accessibility and conformation of OM determines

whether it interacts with contaminants (Murphy et al., 1990; Salloum et al., 2002; Gunasekara

and Xing, 2003; Simpson et al., 2003; Chen et al., 2005; Bonin and Simpson, 2007b; Chefetz and

Xing, 2009). This suggests that investigations involving sorption to whole soils and not just soil

fractions such as humic substances are required as sorption domains may be exposed or altered

during extraction processes. Other recent studied examined sorption of a large variety of polar

and non-polar organic contaminants, including some with chemical and physical properties more

similar to those of pharmaceuticals and personal care products than polycyclic aromatic

hydrocarbons but no general consensus on which soil properties influence sorption has been

reached (Niederer et al., 2007; Bronner and Goss, 2011a). For example, Niederer et al. (2007)

observed large differences between air-OM sorption coefficients for OM samples of different

62

origins and found a correlation between this variability and the sorbent aromaticity. Conversely,

Bronner and Goss (2011a) suggested that Pahokee Peat could be used as a surrogate for sorption

to soil OM of varied origin and composition. Thus, further studies of the interactions of

pharmaceuticals and personal care products in soil are necessary to determine whether soil OM

characteristics will influence the sorption to soils of this new and emerging group of soil organic

contaminants.

This study examines the sorption of four contaminants with varying chemical and physical

properties (carbamazepine, sulfamethoxazole, 17β-estradiol and phenanthrene) to soil and two

common soil minerals (kaolinite and montmorillonite). Carbamazepine is an anti-epileptic drug

which has low biodegradability and therefore has the potential to accumulate in soils (Castiglioni

et al., 2006; Kinney et al., 2006; Williams et al., 2006). Sulfamethoxazole is a common

antimicrobial agent and its presence in the environment raises concern over the possibility of

bacteria developing resistance through long-term exposure (Thiele-Bruhn et al., 2004; Hou et al.,

2010) whereas 17β-estradiol is a naturally occurring estrogen with a high endocrine-disrupting

potential (Lee et al., 2003; Bonin and Simpson, 2007a; Zhang et al., 2011). Various

concentrations of carbamazepine, sulfamethoxazole and 17β-estradiol have been detected in

wastewater effluents so these are examples of contaminants which may be introduced to soil via

reclaimed wastewater irrigation (Carballa et al., 2004; Braga et al., 2005; Castiglioni et al., 2006;

Jelic et al., 2011; Zhang et al., 2011). Phenanthrene is a commonly studied polycyclic aromatic

hydrocarbon which is included in this study for comparison purposes. Sorption to soils,

especially those which contain a low OC content and a high mineral content, may be influenced

by contaminant-mineral interactions. Therefore, sorption of carbamazepine, sulfamethoxazole

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and 17β-estradiol to montmorillonite and kaolinite was studied to determine whether sorption to

minerals was considerable or whether sorption was influenced most by the soil OM content.

Furthermore, contaminants were sorbed to five soils chosen for their varying soil OM

composition, which was analyzed by solid-state 13C Cross Polarization-Magic Angle Spinning

Nuclear Magnetic Resonance (CP-MAS NMR) spectroscopy. This was done to test whether the

hypotheses for polycyclic aromatic hydrocarbon sorption, that OM structure and conformation

influence contaminant sorption affinity, were also applicable to pharmaceutical and personal care

product sorption. The overall objective of this research was to determine which soil properties

govern sorption of pharmaceuticals and personal care products: mineral content, OC content or

OM composition.

3.2. Materials and Methods

3.2.1. Soil and mineral samples

Surface soil samples from varying locations and which have different OM properties (quantity

and composition) were selected for this study. The Pahokee Peat soil was purchased from the

International Humic Substances Society (St Paul, Minnesota) as an example of a soil with a low

mineral content and a high OC content (48.35%; Bonin and Simpson, 2007a). A forest fire

impacted surface soil (0-10 cm) was collected in October of 2002 from a site near the town of

Nestow, Alberta. The area was dominated by Jack pine (Pinus banksiana) and experienced a

wildfire in the summer of 2001 (Otto et al., 2006). The Charred soil has an OC content of 14.3%

and has high aromaticity (Otto et al., 2006). A pine forest soil (O horizon) was collected from a

forest near Hinton, Alberta with predominantly Lodgepole pine (Pinus contorta) (Otto and

Simpson, 2005). This soil has an OC content of 23.1% and contains OM that is of a more polar

64

and fresh nature (Otto and Simpson, 2005). A grassland soil was collected from the University of

Alberta Ellerslie Research Station in Edmonton where the predominant vegetation was western

wheatgrass (Agropyron smithii; Otto and Simpson, 2006). This soil has an OC content of 5.26%

and the dominant mineral is montmorillonite but the mineralogy also includes kaolinite, illite and

chlorite (Dudas and Pawluk, 1969; Otto and Simpson, 2006). An agricultural Brandon loam soil

(0-15 cm depth) was sampled in the fall of 2007 from a plot at the Central Experimental Farm in

Ottawa, Ontario and provides an example of a soil to which reclaimed water may potentially be

applied to for irrigation purposes (Ma et al., 2003). The OC content of this soil is 1.66% (LECO

combustion method) and the minerals present are feldspar, amphibole, chlorite, illite and mixed-

layer minerals (MacLean and Brydon, 1963). Soil samples were air dried and passed through a 2

mm sieve. Before use in sorption experiments, soils were finely ground using a mortar and

pestle.

The clay minerals kaolinite (KGa-1b Washington County, Georgia) and montmorillonite (STx-

1b Gonzales County, Texas) were purchased from The Clay Minerals Society’s Source Clays

Repository (West Layfayette, Indiana) and were used as received. Montmorillonite has a surface

area (N2) of 83.79 ± 0.22 m2/g and a cation exchange capacity of 84.4 meq/100 g whereas

kaolinite has a surface area of 10.05 ± 0.02 m2/g and a cation exchange capacity of 2.0 meq/100

g (Van Olphen and Fripiat, 1979).

3.2.2. Solid-state 13C NMR analysis

Soil samples were analyzed by solid-state 13C CP-MAS NMR to characterize differences in the

OM composition. Solid-state 13C NMR provides structural information about all of the

65

components present in the soil OM (Simpson et al., 2011). Prior to NMR analysis, the soils with

the lowest OC content (Grassland and Agricultural) were treated with 10% HF acid to

concentrate the OC and subsequently increase the signal to noise ratio without significantly

altering the 13C distribution of the samples (Schmidt et al., 1997; Rumpel et al., 2006). The

treated samples were rinsed repeatedly with deionized water to remove excess HF and freeze-

dried. NMR spectra were acquired with a Bruker BioSpin Avance III 500 MHz spectrometer

equipped with a 4 mm H-X MAS probe (Bruker BioSpin, Rheinstetten, Germany).

Approximately 100 mg of soil was packed in a 4 mm zirconium rotor and sealed with a Kel-F

cap. The spectra were acquired with a spinning rate of 13 kHz, a ramp-CP contact time of 1 ms

and a 1 s recycle delay. Over an acquisition time of 14 ms, the number of scans to obtain the

spectra ranged from 23 K to 68 K (1 K=1024 transient) and 1024 data points were collected.

Spectra were processed with a zero filling factor of 2 and 50 Hz line broadening. Chemical shifts

were calibrated using the carboxyl signal of glycine (176.03 ppm) as an external standard and

were assigned according to published studies (Baldock et al., 1992; Guggenberger et al., 1995;

Preston et al., 1997; Mao et al., 2000). Spectra were integrated using AMIX software (v. 3.9.7;

Bruker BioSpin) and were divided into four regions corresponding to: alkyl carbon (0-50 ppm),

O-alkyl carbon (50-110 ppm), aromatic carbon (110-160 ppm) and carbonyl carbon (160-200

ppm). Relative C percentages were calculated by dividing the area of each region by the total

carbon signal (0-200 ppm).

3.2.3. Batch sorption experiments

Carbamazepine (≥98% purity), phenanthrene (98% purity) and 17β-estradiol (≥98% purity) were

purchased from Sigma-Aldrich (St Louis, Missouri) and sulfamethoxazole (≥98% purity) was

66

purchased from Fluka Analytical (St Louis, Missouri). Chemical and physical properties of the

contaminants are listed in Table 3.1. Sorption experiments were performed in an aqueous

background solution of 0.01 M calcium chloride (Fisher Chemicals, Fair Lawn, New Jersey)

with 10-4 M mercury (II) chloride (> 99.5% purity, Sigma-Aldrich, St Louis, Missouri) added as

a biocide (Wolf et al., 1989). The solution was adjusted to pH 7 with drops of dilute NaOH and

the pH was verified using an Accumet® Basic pH meter. Stock solutions of the contaminants

were prepared in methanol and then diluted with the background solution to obtain the desired

concentrations. Methanol concentrations were less than 0.5% of the total volume to reduce

cosolvent interference with sorption behaviour (Feng et al., 2006). Contaminant concentration

ranges for sorption experiments were based on aqueous solubility properties (i.e.: higher

concentration ranges were used for water soluble contaminants). Final solution concentrations

for carbamazepine and sulfamethoxazole ranged from 2-10 mg/L and for phenanthrene and 17β-

estradiol from 0.2-1.0 mg/L. Contaminant solutions were added to pre-weighed soil samples in

13 mL Kimax glass test tubes with Teflon-lined screw caps containing 5 glass beads to aid with

mixing. Preliminary experiments were conducted to ensure that sorption was between 20-80% to

minimize analytical error. For each concentration, test tubes containing no soil served as control

samples. Test tubes were shaken for 48 hours (preliminary experiments indicated that apparent

equilibrium was reached before this time) on an Eberbach 6010 shaker at room temperature. The

samples were then centrifuged (1000 rpm, 1 hour) and 2 mL of the supernatants were placed in

amber vials for analysis by high-performance liquid chromatography (HPLC). Sorbed amounts

were calculated from the difference between solution concentrations of controls and equilibrium

aqueous-phase concentrations. Sorption isotherms were constructed from three replicates of five

concentration points.

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Contaminant concentrations were measured using an Agilent 1100 HPLC system equipped with

an autosampler, a 5μm PrevailTM C18 4.6 × 250 mm column, a diode array detector and a

fluorescence detector. Solvent A, 0.1% glacial acetic acid in water, was prepared by mixing 500

μL of acetic acid (ACS grade, Fischer Scientific, Fair Lawn, New Jersey) with water (Millipore

Synergy® UV) in a 500 mL volumetric flask and filtering through a 0.45 μm mixed cellulose

ester membrane filter (Millipore). Solvent B was acetonitrile (Optima® grade, Fisher Scientific,

Fair Lawn, New Jersey). Carbamazepine analysis parameters were: 5 μL injection volume, a

mobile phase of 25% A and 75% B, a flow rate of 1 mL min-1, column temperature of 25°C,

diode array detection at an absorbance wavelength of 286 nm and a retention time of ~4 min

(Chefetz et al., 2008). Sulfamethoxazole analysis parameters were: 5 μL injection volume, a

mobile phase of 60% A and 40% B, a flow rate of 1 mL min-1, column temperature of 25°C,

diode array detection at an absorbance wavelength of 265 nm and a retention time of ~5.9 min

(Hou et al., 2010). 17β-estradiol analysis parameters were: 20 μL injection volume, a mobile

phase of 40% A and 60% B, a flow rate of 1 mL min-1, column temperature of 25°C,

fluorescence detection with an excitation wavelength of 226 nm and an emission wavelength of

310 nm and a retention time of ~7.6 min (Sun et al., 2010). Phenanthrene analysis parameters

were: 20 μL injection volume, a mobile phase of 10% A and 90% B, a flow rate of 1 mL min-1,

column temperature of 25°C, diode array detection at an absorbance wavelength of 254 nm and a

retention time of ~5.9 min (Bonin and Simpson, 2007b).

Origin Version 7.0 (Origin Lab, Northampton, Massachusetts) was used to calculate isotherm

coefficients. Sorption isotherms were modelled using the Freundlich equation: x/m = KFCen

where x/m is the equilibrium solid-phase solute concentration in mg/g, KF is the Freundlich

68

sorption coefficient with units of (mg/g)/(mg/L)n, Ce is the aqueous-phase solute concentration

with units of mg/L and n is the isotherm nonlinearity index (unitless). Isotherms were also

modelled using a linear sorption isotherm: x/m = KdCe where Kd is the distribution coefficient in

units of L/g. Where appropriate, OC normalized coefficients were calculated using Koc = Kd/foc

where foc is the percent normalized fraction of OC in the soil (Feng et al., 2006; Bonin and

Simpson, 2007a). Origin Version 7.0 (Origin Lab, Northampton, Massachusetts) was also used

to obtain linear regression parameters for the relationship between Koc and soil O-alkyl carbon

content.

3.3. Results and Discussion

3.3.1. Sorbent characteristics

The solid-state 13C NMR spectra of the soil samples are shown in Fig. 3.1 and the relative

integration results for the different structures are listed in Table 3.2. Resonances within the solid-

state 13C NMR spectra (Fig. 3.1) are classified into four general categories: a) alkyl carbon, b) O-

alkyl carbon, c) aromatic carbon and d) carboxyl and carbonyl carbon. Resonances from 0-50

ppm are from alkyl carbon such as methyl groups (0-25 ppm) and methylene from simple

aliphatic chains (25-35 ppm) which arise from lipids and waxes from plant cuticles as well as the

polyesters cutin and suberin (Guggenberger et al., 1995; Kögel-Knabner, 1997; Preston et al.,

1997; Mao et al., 2000). Resonances from 50-110 ppm can be assigned to O-alkyl carbon such as

alcohols, carbohydrates and methoxy carbon (56 ppm) from proteins or lignin and ethers

(Preston et al., 1997; Mao et al., 2000). Within the O-alkyl region, between 90-110 ppm,

resonances from anomeric carbon of carbohydrates are observed (Kögel-Knabner, 1997; Preston

et al., 1997; Mao et al., 2000). Aromatic resonances appear between 110-160 ppm and are from

69

aromatic constituents found in lignin, tannins and aromatic side chains of proteins such as

phenylalanine and tyrosine as well as black carbon (Baldock et al., 1992; Guggenberger et al.,

1995). Finally, carboxyl and carbonyl carbon resonances appear between 160-200 ppm (Baldock

et al., 1992; Guggenberger et al., 1995).

Of the soil samples analyzed, the Peat contains the highest relative percentage of alkyl carbon

and the least aromatic carbon. The Charred soil is enriched in aromatic carbon likely in the form

of black carbon which is a remnant of incomplete combustion and can account for as much as

30-45% of total OC in charred soils (Skjemstad et al., 1996; Schmidt et al., 1999; Cornelissen et

al., 2005). Carbohydrates and alcohols are consumed as fuel during forest fires resulting in a

relatively small contribution of O-alkyl components which is consistent with a high degree of

soil OM degradation (Baldock et al., 1992; Fernandez et al., 1997; Kavdir et al., 2005).

Conversely, the Pine Forest soil contains the largest percent of O-alkyl carbon of all the soils

which is consistent with fresh OM and a relatively low amount of aromatic carbon (Baldock et

al., 1992). The Grassland sample is the second most aromatic of the samples. Lastly, the

Agricultural soil contains a large amount of O-alkyl carbon and is of intermediate aromaticity.

Therefore, NMR analysis reveals that the OM composition of the soils varies and is likely the

result of many factors including overlying vegetation, climate, mineral content, microbial

activity, recent wildfires and land-use (Baldock et al., 1992; Guggenberger et al., 1995; Zech et

al., 1997; Quideau et al., 2001; Kögel-Knabner, 2002; Ohno et al., 2007).

70

3.3.2. Sorption coefficients

Isotherm parameters for contaminant sorption to the soils are listed in Table 3.3. All isotherms

were nonlinear; with Freundlich n indices ranging from 0.41 to 0.89 (see Figs. A.1-A.8).

Although sorption isotherms were better modelled using the Freundlich equation, the Freundlich

sorption coefficients (KF) could not be directly compared due to the differing n values and

therefore linear sorption coefficients (Kd) were also calculated. Each contaminant exhibited a

different sorption affinity for the soils. The highest sorption coefficients (Kd values) were

observed for phenanthrene with sorption of the remaining compounds decreasing in the order:

17β-estradiol >> carbamazepine > sulfamethoxazole which inversely correlates to the aqueous

solubilities and positively correlates to the log Kow values of the contaminants. This suggests that

van der Waals forces are likely an important, if not the only, sorption mechanism and that the

contaminants sorbed mostly to OM components rather than water-coated soil minerals (Lee et

al., 2003; Yamamoto et al., 2003; Yu et al., 2004).

To test the role of soil minerals in the sorption of pharmaceuticals and personal care products of

interest, sorption of 17β-estradiol, sulfamethoxazole and carbamazepine to kaolinite and

montmorillonite was measured. Previous studies in our laboratory found that phenanthrene

sorption to kaolinite and montmorillonite was not detectable so this was not repeated in this

study (Feng et al., 2006; Bonin and Simpson, 2007b). Sorption of carbamazepine, 17β-estradiol

and sulfamethoxazole to kaolinite was not detected. This is consistent with other reports that

suggest contaminant affinity for kaolinite is lower than for montmorillonite likely due to its

decreased surface area and nonexpanding nature (Van Emmerik et al., 2003; Bonin and Simpson,

2007b). Carbamazepine and 17β-estradiol sorption to montmorillonite was detected but

71

sulfamethoxazole sorption was not observed. Sulfamethoxazole has a pKa value of 5.29 and is

negatively charged at the experimental pH whereas the other contaminants are neutral (Hou et

al., 2010). Decreasing sulfonamide sorption with increasing pH due to sulfonamide

deprotonation has been observed previously (Gao and Pedersen, 2005). Sorption of

carbamazepine on montmorillonite resulted in an isotherm with a KF that is orders of magnitude

lower than the KF values for carbamazepine sorption to the majority of the soils (Table 3.3).

Carbamazepine sorption isotherms (see Figs. A.1 and A.2) showed that similar amounts of the

contaminant sorbed to montmorillonite and the Agricultural soil in the low concentration range.

However, montmorillonite was not one of the main minerals found in the Agricultural soil and

the mixture of minerals present may have sorbed lower amounts of carbamazepine. Furthermore,

the sorption isotherm for montmorillonite had an n index of greater than 1 which implies that

sorbate-sorbate interactions were stronger than sorbate-sorbent interactions and that

carbamazepine would rather remain in solution at low concentrations than interact with the

mineral surface (Schwarzenbach et al., 2003). Therefore, the shape of the isotherms suggests that

in this experiment, carbamazepine is more likely to sorb to OM than to mineral surfaces. 17β-

estradiol sorption to montmorillonite also gave an isotherm with a KF which is an order of

magnitude lower than any of the values for 17β-estradiol sorption to soil (Table 3.3). These

results suggest that for the soils studied, sorption of the pharmaceuticals and personal care

products to the mineral phase is not the dominant sorbent property and that OC is the main

sorbent characteristic that drives sorption of these compounds to soil provided the soils are not

dry.

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Sorption affinity of the contaminants for the soils also generally increased with OC content

which further supports the hypothesis that sorption occurred mainly through interactions with the

soil OM (Schwarzenbach and Westall, 1981; Hou et al., 2010). The notable exception to this

trend is the Charred soil which had Kd coefficients of equal or greater value than those of the

Peat even though the Charred soil has a much lower OC content. This may relate to the OC

quality of this sample because the Charred soil contains black carbon which has a high sorption

affinity for contaminants as compared to other types of OM found in soils and sediments (Jonker

and Koelmans, 2002; Cornelissen and Gustafsson, 2004; Cornelissen et al., 2005; Sun et al.,

2010). Sorption to black carbon would also account for the observed nonlinearity of the sorption

isotherms and lower n indices (Table 3.3) for the Charred soil (Accardi-Dey and Gschwend,

2002; Cornelissen and Gustafsson, 2004; Cornelissen et al., 2005; Sun et al., 2010) because of

the condensed and highly aromatic nature of black carbon which behaves as a “glassy” domain

(Xing and Pignatello, 1997; Cornelissen et al., 2005). Although there was an increase in sorption

with increasing OC content, this was not a linear relationship (see Figs. C.1 and C.2) indicating

that other factors such as OM characteristics must be considered.

3.3.3. Comparison of measured and calculated sorption coefficients

Poly parameter linear free energy relationships (pp-LFERs) which predict the log Koc values for

contaminant sorption to Pahokee Peat are found in the literature (Endo et al., 2009; Bronner and

Goss, 2011b). These relationships predict the sorption of a compound based on sorbate

descriptors including: excess molar refractivity (E), molar volume (V), H-bond acidity (A) and

basicity (B) and dipolarity/polarizability (S; Endo et al., 2009; Bronner and Goss, 2011b). Log

Koc values were calculated from some of these pp-LFER equations and were compared to those

73

measured in this study. Sorbate descriptors for carbamazepine, sulfamethoxazole, 17β-estradiol

and phenanthrene are given in Table 3.4.

The log Koc values for sorption to Peat measured in this study are as follows: 2.38 ± 0.04 for

carbamazepine, 2.30 ± 0.02 for sulfamethoxazole, 3.49 ± 0.04 for 17β-estradiol and 4.20 ± 0.02

for phenanthrene. Based on the pp-LFER equation for sorption to Peat at low concentrations

published by Endo et al. (2009), the calculated log Koc values are: 4 ± 1 for carbamazepine, 4 ± 1

for sulfamethoxazole, 8 ± 1 for 17β-estradiol and 5.6 ± 0.7 for phenanthrene. This pp-LFER

consistently overestimates the log Koc measured for sorption of the contaminants. This is

expected as low contaminant concentrations were defined as orders of magnitude lower (or

<10%) than the aqueous solubility limits whereas the concentrations used in this experiment

were generally above these limits (Endo et al., 2009). The log Koc values calculated based on the

pp-LFER equation for sorption to Peat at high concentrations published by Endo et al. (2009)

are: 2.6 ± 0.9 for carbamazepine, 2 ± 1 for sulfamethoxazole, 5 ± 1 for 17β-estradiol and 4.4 ±

0.7 for phenanthrene. The measured log Koc values of all contaminants fall within error of these

predicted values which suggests that this equation can be used to accurately predict log Koc for

sorption to Peat. Finally, based on the pp-LFER equation for sorption to Peat published by

Bronner and Goss (2011b), the calculated log Koc values are: 1.9 ± 0.5 for carbamazepine, 0.7 ±

0.4 for sulfamethoxazole, 2.3 ± 0.5 for 17β-estradiol and 4.0 ± 0.3 for phenanthrene. The

measured log Koc for phenanthrene agrees with the calculated value within error. However, this

pp-LFER underestimates the log Koc values measured for carbamazepine, 17β-estradiol and

sulfamethoxazole. This suggests that this equation does not accurately predict sorption of all

74

contaminants to Pahokee Peat. The log Koc values measured in this study are better modelled by

the pp-LFER for sorption to Peat at high sorbate concentrations published by Endo et al. (2009).

3.3.4. Relationship between sorption and OM structure

Comparisons between the OM structure of the soils and the Koc values were made to determine

whether the soil OM composition influenced sorption affinity. To account for the higher sorption

affinity of contaminants for black carbon than other types of soil OM, the use of a black carbon

normalized sorption coefficients (KBC) for describing sorption to soils has been proposed

(Gustafsson et al., 1997; Accardi-Dey and Gschwend, 2002). However, accurate quantification

of the fraction of black carbon in a soil can be difficult (Cornelissen et al., 2005) and in this

experiment, the fraction of black carbon in the soils was unknown. Other studies have calculated

Koc values for soils known to contain black carbon and other condensed OM fractions, such as

kerogen and nonhydrolyzable carbon, and have compared sorption of contaminants to these

fractions with sorption to other OM fractions (Salloum et al., 2002; Yu et al., 2006; Sun et al.,

2010; Zhang et al., 2010). Therefore, while the Charred soil in this study is assumed to contain

black carbon, the sorption coefficients are normalized to OC content for comparison with the

other soils.

Specifically, the relationships of Koc with aliphaticity and aromaticity (Figs. 3.2a and 3.2b) were

tested as previous research with polycyclic aromatic hydrocarbons and other organic molecules

has highlighted the importance of these two components (Chin et al., 1997; Xing, 1997; Chiou et

al., 1998; Perminova et al., 1999; Chefetz et al., 2000; Mao et al., 2002; Salloum et al., 2002;

Niederer et al., 2007; Wang et al., 2011). No clear trend was observed between the Koc values of

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the neutral contaminants (carbamazepine, 17β-estradiol and phenanthrene) and aliphaticity of the

soil OM (Fig. 3.2a). For example, the Charred soil contained only an intermediate amount of

alkyl carbon but gave the highest Koc values. For carbamazepine, 17β-estradiol and

phenanthrene, the Grassland soil also had a higher Koc value than the aliphatic rich soil samples

(Agricultural and Peat). For sulfamethoxazole, a trend of increasing sorption with increasing

alkyl carbon content was observed for the Pine Forest, Grassland, Agricultural and Peat soils. As

the sulfamethoxazole is negatively charged at the experimental pH, electrostatic interactions may

have been important which altered the sorption trends compared to the neutral contaminants

(Tolls, 2001; Tülp et al., 2009; Vasudevan et al., 2009). For example, Ca2+ is present in the

solution so cation bridging between the pharmaceutical and deprotonated hydroxyl or

carboxylate groups of the OM may be an important sorption mechanism for sulfamethoxazole

(Tolls, 2001; Tülp et al., 2009; Vasudevan et al., 2009). However, the Charred soil also resulted

in highest Koc value for sulfamethoxazole. Furthermore, for all of the contaminants, the Charred

soil and the Grassland soil contained the same relative amount of alkyl carbon but had very

different Koc values. These observations suggest that aliphaticity alone is not an accurate

predictor of sorption.

The trend between Koc and aromaticity is also unclear (Fig. 3.2b). For carbamazepine, 17β-

estradiol and phenanthrene there is an increase in sorption with higher aromaticity based on

sorption to the Pine Forest, Agricultural, Grassland and Charred soils. However, the Peat resulted

in high Koc values but contains the lowest amount of aromatic carbon. For sulfamethoxazole, the

Grassland soil, which is the second most aromatic of the samples, had one of the lowest Koc

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values whereas the Peat had the second highest Koc value. Therefore, OM aromaticity may not be

a suitable predictor of pharmaceutical and personal care product sorption.

To further assess the role of OM composition in sorption, the Koc values of the contaminants

were also plotted against the O-alkyl content of the five soils (Fig. 3.2c). The Koc values of all

contaminants generally decreased with increasing O-alkyl content of the soils. Linear regression

parameters for the relationship between Koc and O-alkyl carbon content are reported in Table 3.5.

Based on these values, there is a strong and significant (α = 0.1) negative correlation between Koc

and O-alkyl carbon content for carbamazepine, 17β-estradiol and phenanthrene. The relationship

for sulfamethoxazole is relatively weak and is not significant (α = 0.1) which further exemplifies

the difference in reactivity of this negatively charged contaminant compared to the neutral

compounds studied in this experiment. The negative relationship between Koc and the O-alkyl

content of the soils is consistent with reports which have observed a negative correlation between

sorbent polarity and Koc values of polycyclic aromatic hydrocarbons (Xing, 1997; Chiou et al.,

1998; Chen et al., 2005). Studies have also shown that carbohydrates act as poor sorption sites

for contaminants (Salloum et al., 2002; Simpson et al., 2003; Chen et al., 2005). For example,

Salloum et al. (2002) studied the sorption of phenanthrene to natural OM samples including

algae, cellulose, cuticle and lignin and observed the lowest Koc for sorption to cellulose.

Similarly, OM samples treated to remove polysaccharides displayed higher phenanthrene

sorption than untreated samples (Simpson et al., 2003; Chen et al., 2005). One explanation for

this observed negative relationship may be that O-alkyl carbon blocks important contaminant

sorption sites. For soils with low O-alkyl carbon content, high affinity sorption sites may be

revealed resulting in increased contaminant sorption. The Koc value for the Charred soil may be

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higher than the other soils not only due to the presence of highly condensed and aromatic black

carbon but also due to the lack of O-alkyl species ensuring that these black carbon components

are fully accessible for interaction. While aromatic or aliphatic carbon may interact favourably

with contaminants, their accessibility rather than their presence in soil may be the factor

controlling sorption. Niederer et al. (2007) similarly suggested that sorption of contaminants was

most influenced by the availability of sorption sites rather than specific chemical characteristics

of the soil OM. To explain this phenomenon, the authors proposed that OM components such as

carboxylic and phenolic groups could undergo internal hydrogen bonding forming highly cross-

linked regions into which contaminants would generally be unable to enter for sorption (Niederer

et al., 2007). As O-alkyl components are polar, these could be involved in this cross-linking and

therefore the above reasoning could provide an alternative explanation for the observed decrease

in sorption with increasing O-alkyl content.

Other studies of the sorption of organic contaminants have also suggested the importance of OM

conformation and accessibility (Murphy et al., 1994; Salloum et al., 2001; Feng et al., 2006;

Pignatello et al., 2006; Bonin and Simpson, 2007b). For example, Murphy et al. (1994) proposed

that the conformation of OM sorbed to a mineral surface is dependent on its configuration in

solution where it could adopt a coiled or elongated structure based on pH, ionic strength and

cation valence. The sorption affinity of hydrophobic organic contaminants has been observed to

vary for OM-mineral complexes formed under different solution conditions, which suggests that

different conformations of OM on mineral surfaces result in the exposure or concealment of

contaminant sorption sites (Murphy et al., 1994; Feng et al., 2006). Previous studies have found

that reconstituted Koc values (calculated from the Koc values for contaminant sorption to humin

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and humic acid fractions) were greater than measured Koc values for sorption to whole soil

(Salloum et al., 2001; Bonin and Simpson, 2007b). This implies that with fractionation additional

or more favourable sorption sites are exposed which are not accessible in whole soils (Salloum et

al., 2001; Bonin and Simpson, 2007b). As well, Feng et al. (2006) observed phenanthrene had a

higher sorption affinity for OM-kaolinite complexes than OM-montmorillonite complexes. Both

mineral surfaces preferentially sorbed polymethylene compounds however, montmorillonite also

sorbed peptides which the authors believed could limit the accessibility of the hydrophobic

domains for contaminants (Feng et al., 2006). Furthermore, Pignatello et al. (2006) observed

suppressed sorption of organic compounds to wood char coated with humic and fulvic acids

which was attributed to the blocking of high affinity sorption sites.

A previous study of a whole soil by 1H High Resolution-Magic Angle Spinning (HR-MAS)

NMR spectroscopy determined that carbohydrates are one of the main types of compounds

detected at the soil-water interface (Simpson et al., 2001). 1H HR-MAS NMR is a technique

which allows the application of solution-state NMR experiments to samples that are not fully

soluble, such as soils (Simpson et al., 2001). The soil is swollen by an NMR solvent and the only

compounds detected are those which are in contact with the solvent (Simpson et al., 2011). The

use of D2O provides detail about which components would be available for interaction at the

solid-water interface (Simpson et al., 2001). The presence of O-alkyl species at the soil-water

interface supports that these compounds could indeed be blocking other soil OM components

which would act as more favourable sorption sites for contaminants. Conversely, aromatic

compounds were only detected after soils were swollen by DMSO-d6, a more penetrating solvent

than D2O which can break hydrogen bonds (Simpson et al., 2001). Aromatic carbon is likely

79

buried in hydrophobic domains and therefore not available for direct interaction with the

contaminants which may be the reason that a strong relationship between Koc and aromaticity is

not observed (Simpson et al., 2001). Combined with the weak relationships between Koc and

aliphaticity and aromaticity, the negative relationship between Koc and soil O-alkyl content

suggests that OM composition alone cannot be used to predict sorption of pharmaceuticals and

personal care products but that OM conformation and accessibility must also be considered.

3.4. Conclusions

Contaminants studied here sorbed more to the soil OM than kaolinite and montmorillonite

suggesting that OM is the primary sorbent property that governs the fate of these contaminants in

soil. The neutral pharmaceuticals (carbamazepine and 17β-estradiol) displayed trends in Koc

which were comparable to those of phenanthrene. At the pH of this study, sulfamethoxazole was

negatively charged and displayed different trends in Koc values from the neutral contaminants

which may have been the result of significant electrostatic interactions such as cation bridging

between sulfamethoxazole and negative functional groups of the soil OM (Tolls, 2001; Tülp et

al., 2009; Vasudevan et al., 2009). The contaminant properties such as log Kow had a more

notable impact on the amount of chemical which was sorbed to soil and therefore would

determine the relative contaminant mobility. Carbamazepine and sulfamethoxazole Koc values

were considerably lower than those for 17β-estradiol and phenanthrene and therefore these

pharmaceuticals may have a higher mobility in soil. This is consistent with reports that observed

carbamazepine and sulfamethoxazole in groundwater (Ternes et al., 2007; Barnes et al., 2008).

Weak relationships were observed between Koc and soil aliphaticity or aromaticity. Not only

should the types of OC present in soils be considered but also the accessibility of those

80

components as this will influence their ability for interaction. A negative relationship between

Koc and the O-alkyl content may be due to carbohydrates blocking high affinity sorption sites or

due to O-alkyl and carboxylic components forming highly cross-linked regions for which

contaminants have restricted access. Consequently, application of reclaimed wastewater to soils

with relatively fresh OM enriched in O-alkyl components, such as the Pine Forest and

Agricultural soil used in this study, may result in higher mobility of pharmaceuticals and

personal care products. Additional studies involving sorption of various pharmaceuticals to soils

over a range of environmentally relevant pH will assist in verifying these observations.

Specifically, the sorption of pharmaceuticals to a soil which has been coated with O-alkyl

compounds should be compared with sorption to the unaltered soil to determine whether addition

of these components causes a decrease in sorption affinity.

81

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3.6. Tables

Table 3.1: Selected chemical and physical properties of the contaminants. Chemical Structure Molecular

Formula Molecular

Weight (g/mol)

Water Solubility

(mg/L)

Log Kow

pKa

Carbamazepine

C15H12N2O 236.27 17.7a 2.45a 14a

Sulfamethoxazole C10H11N3O3S 253.28 356b 0.9b 1.85, 5.29b

17β-estradiol

C18H24O2 272.38 3.85c 4.01c 10.23c

Phenanthrene C14H10 178.23 1.12d 4.57d -

a Monteiro and Boxall (2009). b Hou et al. (2010). c Yamamoto et al. (2003). d Yu and Huang (2005). Table 3.2: Solid-state 13C NMR integration results for soil samples used in sorption studies.

Relative Percentage of Total 13C NMR Signal (0-200 ppm) Sample Alkyl

(0-50 ppm) O-Alkyl

(50-110 ppm) Aromatic

(110-160 ppm) Carbonyl

(160-200 ppm) Peat 40 37 16 7

Charred 27 20 52 1 Pine Forest 26 50 17 7 Grassland 27 32 32 9

Agricultural 29 44 19 8

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Table 3.3: Freundlich and linear sorption isotherm parameters for contaminant sorption to various soils. Contaminant abbreviations are as follows: carbamazepine (CBZ), sulfamethoxazole (SMX), 17β-estradiol (E2) and phenanthrene (PHN).

Sorbent-Sorbate System KF ± SE (mg/g)/(mg/L)n

n ± SE

Freundlich r2

Kd ± SE (L/g)

Linear r2

KOC (L/g)

Peat – CBZ 2.1×10-1 ± 1×10-2 0.66 ± 0.04 0.991 1.17×10-1 ± 8×10-3 0.969 2.4×10-1 ± 2×10-2 Charred – CBZ 3.5×10-1 ± 3×10-2 0.51 ± 0.06 0.981 1.9×10-1 ± 2×10-2 0.968 1.3 ± 2×10-1

Pine Forest – CBZ 3.8×10-2 ± 2×10-3 0.76 ± 0.03 0.997 2.6×10-2 ± 1×10-3 0.994 1.14×10-1 ± 5×10-3 Grassland – CBZ 2.2×10-2 ± 1×10-3 0.71 ± 0.03 0.996 1.43×10-2 ± 8×10-4 0.994 2.7×10-1 ± 1×10-2

Agricultural – CBZ 3.76×10-3 ± 7×10-5 0.85 ± 0.01 0.9997 2.88×10-3 ± 8×10-5 0.998 1.73×10-1 ± 5×10-3

Montmorillonite - CBZ 1.0×10-4 ± 5×10-5 4.0 ± 0.3 0.994 not calculated a - n/a bKaolinite - CBZ not detected - - not detected - n/a b

Peat – SMZ 1.6×10-1 ± 1×10-2 0.72 ± 0.04 0.993 9.9×10-2 ± 6×10-3 0.980 2.0×10-1 ± 1×10-2 Charred – SMZ 1.9×10-1 ± 1×10-2 0.64 ± 0.03 0.994 1.04×10-1 ± 8×10-3 0.973 7.2×10-1 ± 5×10-2

Pine Forest – SMZ 4.7×10-3 ± 4×10-4 0.88 ± 0.04 0.996 3.81×10-3 ± 9×10-5 0.998 1.65×10-2 ± 4×10-4 Grassland – SMZ 5×10-3 ± 1×10-3 0.7 ± 0.2 0.893 2.8×10-3 ± 2×10-4 0.886 5.2×10-2 ± 4×10-3

Agricultural – SMZ 3.23×10-3 ± 9×10-5 0.78 ± 0.017 0.999 2.19×10-3 ± 9×10-5 0.998 1.32×10-1 ± 5×10-3 Montmorillonite - SMX not detected - - not detected - n/a b

Kaolinite - SMX not detected - - not detected - n/a b

Peat – E2 9×10-1 ± 1×10-1 0.60 ± 0.08 0.971 1.5 ± 1×10-1 0.960 3.1 ± 3×10-1 Charred – E2 1.8 ± 3×10-1 0.5 ± 0.1 0.932 3.2 ± 4×10-1 0.907 22 ± 3

Pine Forest – E2 1.8×10-1 ± 1×10-2 0.59 ± 0.04 0.990 3.0×10-1 ± 3×10-2 0.982 1.3 ± 1×10-1 Grassland – E2 1.99×10-1 ± 7×10-3 0.60 ± 0.02 0.997 3.2×10-1 ± 3×10-2 0.992 6.2 ± 5×10-1

Agricultural – E2 2.7×10-2 ± 1×10-3 0.61 ± 0.06 0.981 3.6×10-2 ± 3×10-3 0.997 2.2 ± 2×10-1 Montmorillonite – E2 6.2×10-3 ± 4×10-4 1.0 ± 0.1 0.984 6.3×10-3 ± 2×10-4 0.984 n/a b

Kaolinite – E2 not detected - - not detected - n/a b

Peat – PHN 5.2 ± 5×10-1 0.71 ± 0.07 0.983 7.6 ± 4×10-1 0.993 15.7 ± 9×10-1 Charred – PHN 3.1 ± 1×10-1 0.41 ± 0.02 0.994 6.1 ± 9×10-1 0.928 42 ± 6

Pine Forest – PHN 0.9 ± 1×10-1 0.89 ± 0.09 0.984 1.04 ± 3×10-2 0.991 4.5 ± 1×10-1 Grassland – PHN 4×10-1 ± 2×10-2 0.56 ± 0.03 0.994 6.4×10-1 ± 6×10-2 0.994 12 ± 1

Agricultural – PHN 1.12×10-1 ± 8×10-3 0.76 ± 0.07 0.986 1.43×10-1 ± 7×10-3 0.989 8.6 ± 4×10-1 Montmorillonite – PHN not detected c - - not detected c - n/a b

Kaolinite - PHN not detected c - - not detected c - n/a ba Kd not calculated due to the non-linearity of the sorption isotherm (see Fig. A2). b n/a = not applicable, montmorillonite and kaolinite contain no organic carbon so Koc values could not be calculated. c Feng et al. (2006), Bonin and Simpson (2007b).

89

Table 3.4: Sorbate descriptors for the studied contaminants. Abbreviations are as follows: excess molar refractivity (E), molar volume (V), H-bond acidity (A) and basicity (B) and dipolarity/ polarizability (S).

Chemical E V A B S Carbamazepine 2.15a 1.1811a 0.42 ± 0.07a 1.11 ± 0.05a 1.79 ± 0.16a

Sulfamethoxazole 1.99b 1.7244b 0.59b 1.21b 2.43b

17β-estradiol 1.800c 2.1988c 0.88c 0.95c 3.30c

Phenanthrene 2.055d 1.454d 0.000d 0.26d 1.290d

a Tülp et al. (2008). b Abraham et al. (2009). c Lázaro et al. (2009). d Acree and Abraham (2001). Table 3.5: Linear regression parameters for the relationship between organic carbon normalized sorption coefficients (Koc) and O-alkyl carbon content.

Chemical Equation r p-valuea

Carbamazepine Koc = (-0.0101 ± 0.0007)(% O-alkyl) + (0.62 ± 0.03)

-0.922 0.03

Sulfamethoxazole Koc = (-0.0045 ± 0.0002)(% O-alkyl) + (0.24 ± 0.01)

-0.603 0.3

17β-estradiol Koc = (-0.20 ± 0.02)(% O-alkyl) + (11.3 ± 0.8)

-0.849 0.07

Phenanthrene Koc = (-0.69 ± 0.04)(% O-alkyl) + (39 ± 2)

-0.950 0.01

a p-value for the t-test of the null hypothesis that slope = 0

90

3.7. Figures

Figure 3.1: Solid-state 13C cross polarization-magic angle spinning (CP-MAS) NMR spectra and organic carbon content of soils.

180 160 140 120 100 80 60 40 20 ppm

Charred14.3% OC

Peat48.35% OC

Grassland5.26% OC

Pine Forest23.1% OC

Agricultural1.66% OC

Chemical Shift (ppm)

alkyl

anomeric

aromaticcarboxyl + carbonyl O‐alkyl

OCH3

CH2

CH3

91

Figure 3.2: Relationships between organic carbon normalized sorption coefficients (Koc) and soil a) alkyl carbon content, b) aromatic carbon content and c) O-alkyl carbon content.

15 20 25 30 35 40 45 50

0

5

10

15

20

25

30

35

40

45

50

17β-estradiol Phenanthrene

K oc (L

/g)

% O-alkyl carbon

15 20 25 30 35 40 45 50

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6 Carbamazepine Sulfamethoxazole

Koc

(L/g

)

% O-alkyl carbon15 20 25 30 35 40 45 50 55

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6 carbamazepine sulfamethoxazole

Koc

(L/g

)

% aromatic carbon26 28 30 32 34 36 38 40

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6 carbamazepine sulfamethoxazole

K oc (L

/g)

% alkyl carbon

26 28 30 32 34 36 38 400

5

10

15

20

25

30

35

40

45

50

17β-estradiol phenanthrene

Koc

(L/g

)

% alkyl carbon15 20 25 30 35 40 45 50 55

0

5

10

15

20

25

30

35

40

45

50

17β-estradiol phenanthrene

K oc (L

/g)

% aromatic carbon

a) Koc vs. % alkyl carbon b) Koc vs. % aromatic carbon c) Koc vs. % O-alkyl carbon

92

Chapter 4: Summary and Synthesis The research presented within this thesis demonstrates that similar components of dissolved

organic matter (OM) isolated from Leonardite humic acid, Peat humic acid and a forest soil were

sorbed to clay surfaces even though the initial structure of the dissolved OM isolates varied.

These results suggest that initial dissolved OM composition may not influence preferential

sorption. Therefore, observations made about preferential sorption of a single OM sample should

be applicable to systems with differing OM inputs. Conversely, preferential sorption was

influenced by the type of clay with which the OM interacted. Mainly long chain aliphatic

components which were likely cutin-derived were observed on the kaolinite surface using 1H

High Resolution-Magic Angle Spinning Nuclear Magnetic Resonance (HR-MAS NMR)

spectroscopy. The compounds observed on the montmorillonite surface included both peptides

and aliphatic compounds of various chain lengths. The variation in the components sorbed to

each clay surface was consistent with previous publications which determined that preferential

sorption is influenced by mineral type (Chorover and Amistadi, 2001; Feng et al., 2005; Wang

and Xing, 2005; Ghosh et al., 2009).

1H HR-MAS NMR analysis of two Grassland soils established that the OM components

prevalent at the soil-water interface were aliphatic compounds, carbohydrates and amino acids.

Aromatic components were not observed at the soil-water interface, but instead appeared to exist

within more hydrophobic domains. Similar results were obtained with a Peat soil which had a

low mineral content; this therefore demonstrates the importance of OM-OM interactions.

Consequently, OM-OM interactions may be responsible for the observed differences between the

93

OM sorbed to the clay surfaces and the OM components present at the soil-water interface. For

example, carbohydrates were prevalent at the soil-water interface while these compounds were

only detected in low amounts on the clay surfaces. Therefore, a clearer understanding of the

factors governing OM-OM interactions and not just OM-mineral interactions is required to fully

understand the sorption and preservation of OM in soils. Understanding the interactions of OM

in soils and which OM components are present at the soil-water interface is important as this can

govern the sorption of contaminants.

Sorption of carbamazepine, sulfamethoxazole, 17β-estradiol and phenanthrene to soils was

positively related to the log Kow values of the contaminants, and sorption generally increased

with soil organic carbon content. These results suggest that these contaminants were interacting

with components of the soil OM and not the soil minerals. This hypothesis was supported by

experiments in which no contaminant sorption was detected on kaolinite and only low amounts

of carbamazepine and 17β-estradiol were sorbed to montmorillonite whereas sorption of

sulfamethoxazole and phenanthrene to montmorillonite was not detected. Analysis of the soils by

13C Cross Polarization-Magic Angle Spinning (CP-MAS) NMR showed differences in the soil

OM compositions. No clear relationships between the contaminant organic carbon normalized

sorption coefficients (Koc) and the aliphatic or aromatic content of the soils were observed. A

strong and significant (α = 0.1) negative correlation between the contaminant Koc values and the

O-alkyl carbon content of the soils was observed for carbamazepine, 17β-estradiol and

phenanthrene. One explanation for this may be that O-alkyl components such as carbohydrates

(which were shown to be available at the soil-water interface) are blocking contaminant access to

high affinity sorption sites. Alternatively, O-alkyl components could form hydrogen bonds with

94

functional groups such as carboxylic acids creating highly cross-linked regions within the OM

into which contaminants would be unable to enter for sorption (Niederer et al., 2007). These

results are consistent with reports that have observed a negative relationship between sorption

and sorbent polarity as well as those that have observed low sorption to isolated polysaccharides

(Xing, 1997; Chiou et al., 1998; Salloum et al., 2002; Simpson et al., 2003; Chen et al., 2005).

This also agrees with studies that have suggested the importance of soil OM conformation and

accessibility for contaminant sorption (Murphy et al., 1994; Salloum et al., 2001; Feng et al.,

2006; Pignatello et al., 2006; Bonin and Simpson, 2007). Similar relationships were observed for

the neutral contaminants studied (carbamazepine, 17β-estradiol and phenanthrene). However,

sulfamethoxazole, which was negatively charged at the solution conditions used in this

experiment, showed different sorption trends from the other contaminants. Likely, an

electrostatic interaction such as cation bridging between sulfamethoxazole and negative

functional groups of the OM was an important sorption mechanism for this contaminant (Tolls,

2001; Tülp et al., 2009; Vasudevan et al., 2009). Carbamazepine and sulfamethoxazole would

likely be more mobile in the soil environment as the Koc values for these compounds were much

lower than those of 17β-estradiol and phenanthrene. The results of this study also suggest that the

application of reclaimed wastewater to soils which are enriched in O-alkyl carbon will result in

higher contaminant mobility.

Since the pharmaceuticals and personal care products interacted mainly with the soil OM, factors

which influence the preferential sorption of OM to mineral surfaces and soils also indirectly

control the sorption of these contaminants. In soils, if the conditions do not favour preferential

sorption of O-alkyl components, there should be less blocking of high affinity sites by these

95

species and a larger amount of contaminant sorption. Future experiments which involve the

sequential loading of OM to mineral surfaces and the characterization of the OM present at the

solid-water interface after each loading may provide insight into OM-OM interactions. These

interactions may be as important as OM-mineral interactions for sorption and preservation of

OM and could influence which components are present at the soil-water interface and available

for interaction with contaminants. 1H HR-MAS NMR spectroscopy is valuable for these types of

analyses as it allows direct, molecular-level characterization of the OM components sorbed to the

clays. Furthermore, the sorption of pharmaceuticals and personal care products to organo-clay

complexes or soils with various loadings of O-alkyl compounds should be compared to

determine whether addition of these components causes a decrease in sorption affinity. The

results of this thesis and these future studies will improve the fundamental understanding of

wastewater constituents and their chemical reactions in soil and highlight potential risks of using

wastewater for irrigation.

96

4.1. References

Bonin, J.L., Simpson, M.J., 2007. Variation in phenanthrene sorption coefficients with soil organic matter fractionation: The result of structure or conformation? Environmental Science & Technology 41, 153-159.

Chen, B., Johnson, E., Chefetz, B., Zhu, L., Xing, B., 2005. Sorption of polar and nonpolar aromatic organic contaminants by plant cuticular materials: Role of polarity and accessibility. Environmental Science & Technology 39, 6138-6146.

Chiou, C.T., McGroddy, S.E., Kile, D.E., 1998. Partition characteristics of polycyclic aromatic hydrocarbons on soils and sediments. Environmental Science & Technology 32, 264-269.

Chorover, J., Amistadi, M.K., 2001. Reaction of forest floor organic matter at goethite, birnessite and smectite surfaces. Geochimica Et Cosmochimica Acta 65, 95-109.

Feng, X.J., Simpson, A.J., Simpson, M.J., 2005. Chemical and mineralogical controls on humic acid sorption to clay mineral surfaces. Organic Geochemistry 36, 1553-1566.

Feng, X.J., Simpson, A.J., Simpson, M.J., 2006. Investigating the role of mineral-bound humic acid in phenanthrene sorption. Environmental Science & Technology 40, 3260-3266.

Ghosh, S., Wang, Z.Y., Kang, S., Bhowmik, P.C., Xing, B.S., 2009. Sorption and fractionation of a peat derived humic acid by kaolinite, montmorillonite, and goethite. Pedosphere 19, 21-30.

Murphy, E.M., Zachara, J.M., Smith, S.C., Phillips, J.L., Wietsma, T.W., 1994. Interaction of hydrophobic organic compounds with mineral-bound humic substances. Environmental Science & Technology 28, 1291-1299.

Niederer, C., Schwarzenbach, R.P., Goss, K., 2007. Elucidating differences in the sorption properties of 10 humic and fulvic acids for polar and nonpolar organic chemicals. Environmental Science & Technology 41, 6711-6717.

Pignatello, J.J., Kwon, S., Lu, Y., 2006. Effect of natural organic substances on the surface and adsorptive properties of environmental black carbon (char): Attenuation of surface activity by humic and fulvic acids. Environmental Science & Technology 40, 7757-7763.

Salloum, M.J., Dudas, M.J., McGill, W.B., 2001. Variation of 1-naphthol sorption with organic matter fractionation: The role of physical conformation. Organic Geochemistry 32, 709-719.

Salloum, M.J., Chefetz, B., Hatcher, P.G., 2002. Phenanthrene sorption by aliphatic-rich natural organic matter. Environmental Science & Technology 36, 1953-1958.

Simpson, M., Chefetz, B., Hatcher, P., 2003. Phenanthrene sorption to structurally modified humic acids. Journal of Environmental Quality 32, 1750-1758.

Tolls, J., 2001. Sorption of veterinary pharmaceuticals in soils: A review. Environmental Science & Technology 35, 3397-3406.

Tülp, H.C., Fenner, K., Schwarzenbach, R.P., Goss, K., 2009. pH-dependent sorption of acidic organic chemicals to soil organic matter. Environmental Science & Technology 43, 9189-9195.

Vasudevan, D., Bruland, G.L., Torrance, B.S., Upchurch, V.G., MacKay, A.A., 2009. pH-dependent ciprofloxacin sorption to soils: Interaction mechanisms and soil factors influencing sorption. Geoderma 151, 68-76.

Wang, K.J., Xing, B.S., 2005. Structural and sorption characteristics of adsorbed humic acid on clay minerals. Journal of Environmental Quality 34, 342-349.

97

Xing, B., 1997. The effect of the quality of soil organic matter on sorption of naphthalene. Chemosphere 35, 633-642.

98

Appendix A: Sorption Isotherms

Figure A.1: Carbamazepine sorption isotherms for Charred and Peat soils.

0 1 2 3 4 5 6 7

0.2

0.3

0.4

0.5

0.6

0.7

0.8

x/m

(mg/

g)

Ce (mg/L)

Charred Peat

Figure A.2: Carbamazepine sorption isotherms for Pine Forest, Grassland and Agricultural soils and Montmorillonite.

0 1 2 3 4 5 6 7 8

0.00

0.02

0.04

0.06

0.08

0.10

0.12

0.14

0.16

x/m

(mg/

g)

Ce (mg/L)

Pine Forest Grassland Agricultural Montmorillonite

99

Figure A.3: Sulfamethoxazole sorption isotherms for Charred and Peat soils.

0 1 2 3 4 5 6 7 80.1

0.2

0.3

0.4

0.5

0.6

0.7

x/m

(mg/

g)

Ce (mg/L)

Charred Peat

Figure A.4: Sulfamethoxazole sorption isotherms for Pine Forest, Grassland and Agricultural soils.

0 1 2 3 4 5 6 7 8

0.005

0.010

0.015

0.020

0.025

0.030

x/m

(mg/

g)

Ce (mg/L)

Pine Forest Grassland Agricultural

100

Figure A.5: 17β-estradiol sorption isotherms for Charred and Peat soils.

0.0 0.1 0.2 0.3 0.4 0.50.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0.9

1.0

1.1

1.2

x/m

(mg/

g)

Ce (mg/L)

Charred Peat

Figure A.6: 17β-estradiol sorption isotherms for Pine Forest, Grassland and Agricultural soils and Montmorillonite.

0.0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.80.00

0.02

0.04

0.06

0.08

0.10

0.12

x/m

(mg/

g)

Ce (mg/L)

Pine Forest Grassland Agricultural Montmorillonite

101

Figure A.7: Phenanthrene sorption isotherms for Charred and Peat soils.

0.0 0.1 0.2 0.3 0.4 0.50.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

2.2

2.4

2.6

x/m

(mg/

g)

Ce (mg/L)

Charred Peat

Figure A.8: Phenanthrene sorption isotherms for Pine Forest, Grassland and Agricultural soils.

0.00 0.05 0.10 0.15 0.20 0.25 0.30 0.35 0.40 0.45 0.500.00

0.05

0.10

0.15

0.20

0.25

0.30

0.35

x/m

(mg/

g)

Ce (mg/L)

Pine Forest Grassland Agricultural

102

Appendix B: Aqueous-Phase Concentrations and Equilibrium Solid-Phase Concentrations for Contaminant Sorption to Soil

Table B.1: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Peat soil.

Ce (mg/L) x/m (mg/g) 1.092 ± 0.004 0.198 ± 0.002 2.34 ± 0.02 0.354 ± 0.004 3.75 ± 0.05 0.51 ± 0.01 5.09 ± 0.08 0.61 ± 0.01 6.8 ± 0.2 0.70 ± 0.05

Table B.2: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Charred soil.

Ce (mg/L) x/m (mg/g) 0.70 ± 0.07 0.28 ± 0.01 1.73 ± 0.03 0.487 ± 0.006 3.3 ± 0.1 0.61 ± 0.03 4.4 ± 0.2 0.76 ± 0.08

Table B.3: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Pine Forest soil.

Ce (mg/L) x/m (mg/g) 0.98 ± 0.02 0.037 ± 0.001 2.12 ± 0.03 0.0671 ± 0.0009 3.38 ± 0.03 0.098 ± 0.004 4.7 ± 0.1 0.119 ± 0.004

5.93 ± 0.09 0.149 ± 0.004 Table B.4: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Grassland soil.

Ce (mg/L) x/m (mg/g) 0.981 ± 0.003 0.0223 ± 0.0002 2.160 ± 0.004 0.0394 ± 0.0002 3.54 ± 0.06 0.055 ± 0.001 4.86 ± 0.03 0.0666 ± 0.0004 6.18 ± 0.06 0.084 ± 0.002

103

Table B.5: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Agricultural soil.

Ce (mg/L) x/m (mg/g) 1.340 ± 0.005 0.00481 ± 0.00007 2.74 ± 0.01 0.0089 ± 0.0001 4.32 ± 0.07 0.0128 ± 0.0003 5.7 ± 0.1 0.0164 ± 0.0009

7.26 ± 0.06 0.0200 ± 0.0005 Table B.6: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for carbamazepine sorption to Montmorillonite.

Ce (mg/L) x/m (mg/g) 1.77 ± 0.06 0.006 ± 0.002 3.3 ± 0.1 0.013 ± 0.002

4.39 ± 0.08 0.0330 ± 0.0008 5.05 ± 0.01 0.0606 ± 0.0006 5.56 ± 0.02 0.089 ± 0.002

Table B.7: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for sulfamethoxazole sorption to Peat soil.

Ce (mg/L) x/m (mg/g) 1.21 ± 0.03 0.17 ± 0.01 2.56 ± 0.01 0.304 ± 0.006 3.97 ± 0.08 0.44± 0.01 5.39 ± 0.07 0.552 ± 0.006 7.1 ± 0.2 0.64 ± 0.04

Table B.8: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for sulfamethoxazole sorption to Charred soil.

Ce (mg/L) x/m (mg/g) 1.13 ± 0.08 0.19 ± 0.02 2.42 ± 0.04 0.336 ± 0.008 3.9 ± 0.3 0.47 ± 0.07 5.4 ± 0.2 0.56 ± 0.04 7.1 ± 0.2 0.65 ± 0.05

Table B.9: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for sulfamethoxazole sorption to Pine Forest soil.

Ce (mg/L) x/m (mg/g) 1.38 ± 0.01 0.0067 ± 0.0002 2.86 ± 0.05 0.0119 ± 0.0003 4.43 ± 0.06 0.0172 ± 0.0004 5.9 ± 0.1 0.0218 ± 0.0006

7.43 ± 0.04 0.0281 ± 0.0009

104

Table B.10: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for sulfamethoxazole sorption to Grassland soil.

Ce (mg/L) x/m (mg/g) 1.31 ± 0.03 0.00526 ± 0.00008 2.74 ± 0.07 0.0104 ± 0.0002

4.196 ± 0.006 0.01376 ± 0.00009 6.3 ± 0.3 0.014 ± 0.002 7.3 ± 0.1 0.0206 ± 0.0004

Table B.11: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for sulfamethoxazole sorption to Agricultural soil.

Ce (mg/L) x/m (mg/g) 1.27 ± 0.02 0.0039 ± 0.0001 2.68 ± 0.05 0.0069 ± 0.0002 4.21 ± 0.08 0.0098 ± 0.0002 5.7 ± 0.1 0.0122 ± 0.0005 7.2 ± 0.1 0.0151 ± 0.0004

Table B.12: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Peat soil.

Ce (mg/L) x/m (mg/g) 0.050 ± 0.002 0.152 ± 0.002 0.134 ± 0.002 0.262 ± 0.008 0.205 ± 0.007 0.38 ± 0.01 0.32 ± 0.01 0.43 ± 0.02 0.39 ± 0.02 0.541 ± 0.004

Table B.13: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Charred soil.

Ce (mg/L) x/m (mg/g) 0.044 ± 0.009 0.332 ± 0.008 0.10 ± 0.02 0.56 ± 0.01 0.23 ± 0.04 0.71 ± 0.06 0.27 ± 0.04 1.0 ± 0.1 0.40 ± 0.05 1.1 ± 0.1

Table B.14: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Pine Forest soil.

Ce (mg/L) x/m (mg/g) 0.049 ± 0.004 0.032 ± 0.001 0.128 ± 0.006 0.053 ± 0.002 0.20 ± 0.01 0.075 ± 0.003 0.32 ± 0.01 0.089 ± 0.002 0.41 ± 0.02 0.110 ± 0.003

105

Table B.15: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Grassland soil.

Ce (mg/L) x/m (mg/g) 0.046 ± 0.002 0.0333 ± 0.0008 0.120 ± 0.009 0.055 ±0.002 0.207 ± 0.009 0.075 ± 0.002 0.289 ± 0.005 0.096 ± 0.002 0.39 ± 0.01 0.113 ± 0.003

Table B.16: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Agricultural soil.

Ce (mg/L) x/m (mg/g) 0.094 ± 0.002 0.0075 ± 0.0002 0.228 ± 0.008 0.0104 ± 0.0007 0.357 ± 0.002 0.0140 ± 0.0003 0.49 ± 0.02 0.018 ± 0.001 0.62 ± 0.01 0.0207 ± 0.0006

Table B.17: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for 17β-estradiol sorption to Montmorillonite.

Ce (mg/L) x/m (mg/g) 0.143 ± 0.004 0.0009 ± 0.0001 0.283 ± 0.002 0.0018 ± 0.0002 0.439 ± 0.005 0.0026 ± 0.0003 0.57 ± 0.02 0.0039 ± 0.0003

0.713 ± 0.005 0.0043 ± 0.0003 Table B.18: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for phenanthrene sorption to Peat soil.

Ce (mg/L) x/m (mg/g) 0.041 ± 0.003 0.64 ± 0.02 0.130 ± 0.009 1.11 ± 0.02 0.18 ± 0.02 1.56 ± 0.08 0.27 ± 0.01 1.98 ± 0.09 0.34 ± 0.03 2.48 ± 0.07

Table B.19: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for phenanthrene sorption to Charred soil.

Ce (mg/L) x/m (mg/g) 0.032 ± 0.007 0.71 ± 0.02 0.11 ± 0.03 1.3 ± 0.1 0.18 ± 0.04 1.6 ± 0.1 0.30 ± 0.05 1.9 ± 0.3 0.43 ± 0.04 2.2 ± 0.2

106

Table B.20: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for phenanthrene sorption to Pine Forest soil.

Ce (mg/L) x/m (mg/g) 0.050 ± 0.004 0.082 ± 0.002 0.132 ± 0.003 0.139 ± 0.002 0.20 ± 0.01 0.198 ± 0.004

0.259 ± 0.002 0.271 ± 0.003 0.32 ± 0.01 0.334 ± 0.003

Table B.21: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for phenanthrene sorption to Grassland soil.

Ce (mg/L) x/m (mg/g) 0.051 ± 0.002 0.0812 ± 0.0007 0.142 ± 0.007 0.135 ± 0.004 0.221 ± 0.001 0.17 ± 0.02 0.347 ± 0.005 0.216 ± 0.002 0.458 ± 0.009 0.263 ± 0.004

Table B.22: Aqueous-phase solute concentrations (Ce) and equilibrium solid-phase solute concentrations (x/m) for phenanthrene sorption to Agricultural soil.

Ce (mg/L) x/m (mg/g) 0.0785 ± 0.0007 0.0177 ± 0.0005 0.179 ± 0.003 0.0305 ± 0.0002

0.2786 ± 0.0003 0.0435 ± 0.0009 0.393 ± 0.006 0.0520 ± 0.0008 0.481 ± 0.009 0.067 ± 0.002

107

Appendix C: Relationships between Distribution Coefficients and Fraction of Organic Carbon

Figure C.1: Relationship between distribution coefficients (Kd) and the fraction of organic carbon (foc) in each soil for carbamazepine and sulfamethoxazole.

0.0 0.1 0.2 0.3 0.4 0.5

0.00

0.05

0.10

0.15

0.20

0.25 Carbamazepine Sulfamethoxazole

K d (L/g

)

foc

Figure C.2: Relationship between distribution coefficients (Kd) and the fraction of organic carbon (foc) in each soil for 17β-estradiol and phenanthrene.

0.0 0.1 0.2 0.3 0.4 0.5

0

1

2

3

4

5

6

7

8

17β-estradiol Phenanthrene

Kd (

L/g)

foc