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This is to certify that the
dissertation entitled

INTEGRATING MOLECULAR SIMULATIONS WITH
EXPERIMENTS TO STUDY ORGANIC POLLUTANT
INTERACTIONS WITH CLAY MINERALS

presented by

VANEET AGGARWAL

has been accepted towards fulfillment
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Ph.D. degree in CROP AND SOIL SCIENCE

 

 

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-i_____

INTEGRATING MOLECULAR SIMULATIONS WITH EXPERIMENTS TO STUDY
ORGANIC POLLUTANT INTERACTIONS WITH CLAY MINERALS

By

VANEET AGGARWAL

A DISSERTATION

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

DOCTOR OF PHILOSOPHY
Department of Crop & Soil Science

2005

ABSTRACT
INTEGRATING MOLECULAR SIMULATIONS WITH EXPERIMENTS TO STUDY
ORGANIC POLLUTANT INTERACTIONS WITH CLAY MINERALS
By

VANEET AGGARWAL

Triazine herbicides such as atrazine and simazine have been applied for more than
40 years in the United States. Atrazine is one of the most common herbicide residues
found in groundwaters of Midwest. Sorption to soil components can control the
bioavailability, persistence, and transport of xenobiotics. We investigated the sorption of
three triazine herbicides (atrazine, simazine, and metribuzin) by saponite and beidellite
clay minerals saturated with K+, Cs+, Na+, and Ca“. Saponite clay sorbed a larger
fraction of each pesticide from aqueous solution than did beidellite clay. The lower cation
exchange capacity in saponite (vs. beidellite) presumably results in a less crowded
interlayer with more siloxane surface available for adsorption. Generally, Cs-saturated
clays sorbed more triazines than clays saturated by K+, Na+ or Ca2+. We attribute this to
the smaller hydrated radius of Cs+, which again increases the siloxane surface available
for sorption. Furthermore, the relatively weak hydration of Cs+ reduces the swelling of
clay interlayers, thus making sorption domains less hydrated and more receptive to
hydrophobic molecules. Cs-saponite manifested a sorption of more than 1% atrazine by
weight above equilibrium concentrations of 6 mg/L. This is the largest sorption of
neutral atrazine from water yet reported for an inorganic sorbent. Molecular dynamics

simulations indicate that atrazine interacts with both clay basal planes and with multiple

cations in the clay interlayer forming both inner- and outer-sphere complexes. Similar
effects of exchangeable cation on trichloroethene sorption by saponite clay were also
observed, and our general concept of the factors controlling sorption of nonpolar organics
by clay minerals is becoming more clear.

Molecular simillations were also used to estimate thermodynamics of polar
organic sorption to montmorillonites. Our method shows promise in that simulated
adsorption enthalpies for nitroaromatic compounds have the correct Sign and are
significantly different from that of p-xylene. Further work is still needed so that enthalpy

predictions can be more precise.

DEDICATION

To my daughter, Muskaan, wife Deepa, and parents,
for everything

iv

ACKNOWLEDGEMENTS

I would especially like to thank my advisor, Dr. Brian Teppen, for his guidance,
support, and encouragement throughout my program. I have enjoyed our interactions
immensely. I thank Dr. Teppen for providing me with countless opportunities to grow
both personally and professionally. This dissertation could not have been completed
without his help and support. I thank my guidance committee, Dr. Stephen Boyd, Dr.
Donald Penner, and Dr Thomas Voice for their insight and helpful comments.

I would like to offer my gratitude to numerous people, especially, Dr. Hui Li for
helping me with the laboratory work, Michael Roberts for being there whenever I would
need him, Simone, Jacqui, Jim, Chunhui, Elica, Dr. Yao-Ying, and several others, who
have encouraged and supported me.

I would also like to acknowledge the Dissertation Completion Fellowships
provided by the Environmental Science and Policy Program and by College of
Agriculture and Natural Resources, Kirk and Marjorie Lawton for the graduate student
support, chair of Crop & Soil Sciences department for the Retention Fellowship, and
Punjab Agricultural University for the study leave to pursue my dream of getting PhD. in
the US.

Finally, I have been truly blessed to have the love and support of my wonderful
and encouraging family, my wife Deepa, our precious daughter Muskaan, my parents, my
sister Dimple and her husband Ajay. Their encouragement, understanding, prayers, and
being here on my defense and graduation are much appreciated. This accomplishment is

as much theirs as it is mine, and to them I am forever thankful and grateful.

TABLE OF CONTENTS

LIST OF TABLES viii

CHAPTER I: INTEGRATING MOLECULAR SIMULATIONS WITH
EXPERIMENTS TO STUDY ORGANIC POLLUTANT INTERACTIONS

WITH CLAY MINERALS .................................................................... 1
INTRODUCTION ............................................................................. 2
SORPTION OF ATRAZINE TO SOIL COLLOIDS ......................................................... 6
PROPOSAL ............................................................................................................ 13
LITERATURE CITED ............................................................................................. 22

CHAPTER H: TRIAZINE ADSORPTION BY SAPONITE AND

BEIDELLITE CLAY MINERALS ............................................................................ 29
ABSTRACT ............................................................................................................ 30
INTRODUCTION .................................................................................................... 31
MATERIALS AND METHODS ................................................................................ 34
RESULTS AND DISCUSSION ................................................................................. 4O
LITERATURE CITED ............................................................................................. 55

CHAPTER III: ATRAZINE INTERACTIONS WITH SMECTITE CLAYS:

MOLECULAR SIMULATION APPROACH .......................................................... 59
INTRODUCTION .................................................................................................... 60
MATERIALS AND METHODS ................................................................................ 65
RESULTS AND DISCUSSION ................................................................................. 69
LITERATURE CITED ............................................................................................ 86

CHAPTER IV: ENHANCED SORPTION OF TRICHLOROETHENE BY

SMECTITE CLAY EXCHANGED WITH CS+ ...................................................... 91
ABSTRACT ........................................................................................................... 92
INTRODUCTION ................................................................................................... 93
MATERIALS AND METHODS ............................................................................... 98
RESULTS AND DISCUSSION ............................................................................... 103
LITERATURE CITED ........................................................................................... 1 13

vi

CHAPTER V: MOLECULAR SIMULATIONS TO ESTIMATE
THERMODYNAMICS OF POLAR ORGANIC SORPTION TO

MONTMORILLONITE ........................................................................................... 1 18
INTRODUCTION .................................................................................................. 1 19
MATERIALS AND METHODS .............................................................................. 124
RESULTS AND DISCUSSION ................................................................................ 134
LITERATURE CITED ........................................................................................... 149

vii

LIST OF TABLES

CHAPTER II

Table 1. Selected Physicochemical Properties of Pesticides Used in this Sorption
Study ..................................................................................... 35

CHAPTER V

Table 1. Estimated and measured enthalpies of reaction (kJ/mol) as defined by the
equation designated as cycle [2] and Figure 2. The same estimates for AH;
and AH5 were used for each compound ......................................................... 135

Table 2. The F reundlich adsorption coefficients for substituted nitrobenzenes and
the adsorption energies of the corresponding substituents .......................... . 137

Table 3. Application of algorithm 2 to estimate experimental adsorption
enthalpies ....................................................................................................... 1 47

viii

LIST OF FIGURES

CHAPTER I

Figure 1. Schematic for decomposing the energy changes for transfer of a
compound from aqueous solution to the clay layer (after G033 and

Schwarzenbach, 2001) ...................................................................................... 19
CHAPTER II
Figure 1. Charges on atrazine molecule used in simulation study ............................. 39

Figure 2. Sorption isotherms representing atrazine (a,b), metribuzin (c,d), and
simazine (e,t) uptake from water by reference saponite (SapCa-2) and
beidellite (SBId-l) clays saturated with Cs+, K+, Na+ and Ca2+ ...................... 42

Figure 3. Comparison of our sorption results with all earlier studies of atrazine
sorption to saponites, as well as two representative montmorillonites ............ 44

Figure 4. X-ray diffraction patterns of air-dried Cs-SapCa-2 clay films as a
function of atrazine sorption ............................................................................. 50

Figure 5. Snapshot of a molecular dynamic Simulation of atrazine in the
interlayer region of homoionic K-smectite (K-SapCa—Z). Clay layers were
parallel to the page and have been removed in order to expose the interlayer
complexes. The atom coloring scheme is K = green, O = red, H = white,
C =gray, N =b1ue, and C1 = green (associated with the atrazine molecule) ..... 52

CHAPTER III

Figure 1. A typical starting configuration before beginning the molecular
dynamics simulation. Here, atrazine has been placed into the interlayer
of Cs-SWy-Z with 55 water molecules. The equilibrium dam-Spacing will
eventually be 12.3 A ........................................................................................ 67

Figure 2. Partial atomic charges on the atoms of the atrazine molecule, as used
in all simulation studies. The total charge on the molecule sums to zero ........ 68

Figure 3. Snapshot showing the orientation of atrazine within the interlayer of
a K-saturated smectite like SWy-2 containing (A) one, (B) two, and (C)
three layers of hydration water (red = oxygen, white = hydrogen,

blue = nitrogen, gray = carbon, green = interlayer K+ cations). Images

on the left depict a side-view (x,z plane) of the atrazine-smectite interlayer
complex while the images on the right depict the same from a top—view

(x,y plane), having removed the silicate layers. The simulated door-spacings

for A, B, and C are 12.4, 15.5, and 17.3 A, respectively ............................. 71-72

Figure 4. Snapshots of molecular dynamic simulations of atrazine in the interlayer
regions of four different homoionic smectites, (A) K-SWy-Z, (B) Cs-SWy-2,
(C) K-SAz-l, and (D) K-SapCa-2. Clay layers were parallel to the page and
have been removed in order to expose the interlayer complexes. The atom
coloring scheme is K = green, 0 = red, H = white, C =gray, N =b1ue, and
C1 = green (associated with the atrazine molecule) ......................................... 73

Figure 5. Trajectory of the Cs -- Cl inner sphere nonbonded complex in the
Simulated atrazine-Cs-SWy-Z system represented by Figure 4B .................... 74

Figure 6. Representative snapshot of the relationships of atrazine and cations
to charged sites in model SWy-2 with a monolayer of water for the K- (left)
and Cs-saturated (right) cases. Blue octahedra are the locations of Mg-for-
Al octahedral substitution sites and pink tetrahedral are the locations of Al-
for-Si tetrahedral substitution sites in the montmorillonite layer. The upper
clay layer and the rest of the lower clay layer have been removed for
clarity ................................................................................................................ 78

Figure 7. Snapshots of molecular dynamic Simulations of atrazine in the interlayer
regions of four different homoionic smectites, (A) K-SAz-l, (B) Ca-SAz-l,
(C) K-SWy-2, and (D) Cs-SWy-Z. Clay layers were parallel to the page
and have been removed in order to expose the interlayer complexes. The
atom coloring scheme is K = green, 0 = red, H = white, C =gray, N =b1ue,
and C1 = green (associated with the atrazine molecule) ................................... 79

Figure 8. Trajectory of the K -- N inner sphere nonbonded complex in the
simulated atrazine-K-SWy-Z system represented by Figure 3B and 7C .......... 80

Figure 9. Representative snapshots of the relationships of atrazine and cations
to charged sites in model SWy-Z with a bilayer of water for the K- (left)
and Cs-saturated (right) cases. Blue octahedra are the locations of Mg-for-
Al octahedral substitution sites and pink tetrahedral are the locations of Al-
for-Si tetrahedral substitution sites in the montmorillonite layer. The lower
clay layer and the rest of the upper clay layer have been removed for
clarity ................................................................................................................ 81

Figure 10. Snapshots of molecular dynamic simulations of atrazine in the
interlayer regions of two different homoionic smectites, (A) C a-SAz-l ,
and (B) K-SWy-2. Clay layers were parallel to the page and have been

removed in order to expose the interlayer complexes. The atom coloring
scheme is K = green, 0 = red, H = white, C =gray, N =b1ue, and C1 = green
(associated with the atrazine molecule) ........................................................... 83

CHAPTER IV
Figure l. Sorption isotherms representing trichloroethene uptake from water by
reference saponite (SapCa-2) clay saturated with (a) calcium, and potassium,

and (b) cesium ............................................................................................... 105

Figure 2. X-ray diffraction pattern of reference saponite (SapCa-2) clay in oriented
clay films as a function of different cation saturation at 100% humidity ...... 109

CHAPTER V
Figure 1. Charges on para-dinitrobenzene molecule used in simulation study ......... 126

Figure 2. Schematic for decomposing the energy changes for transfer of a
compound from aqueous solution to the clay layer (afier G053 and

Schwarzenbach, 2001) ................................................................................... 128
Figure 3. Validity of algorithm with experimental data ............................................ 144
Figure 4. Sensitivity Analysis .................................................................................... 146

“Images in this thesis/dissertation are presented in color.”

xi

CHAPTER I

INTEGRATING MOLECULAR SIMULATIONS WITH EXPERIMENTS TO

STUDY ORGANIC POLLUTANT INTERACTIONS WITH CLAY MINERALS

Introduction

Atrazine is one of the most commonly used agricultural herbicides in the world
and has been applied for more than 30 years by com and sorghum growers in the United
States (Ribaudo and Bouzaher, 1994). In 1994, over 47 million pounds of atrazine were
applied nation-wide covering about 77 % of the area cropped under com-soybean rotation
(Pike and Knake, 1996). About 2.4 million pounds of atrazine were applied on 80 0/0 of
the cropped area in Michigan, making it the most widely used herbicide in Michigan
(Pike and Knake, 1996). Atrazine is also one of the most common herbicide residues
found in groundwaters of Midwest (Blanchard and Donald, 1997; Burkart et al., 1999;
Kolpin et al., 1998). Atrazine and two of its metabolites were detected in 26 % of 303
sampled Midwest wells in the United States (Kolpin et al., 1996) and is detected 10 to 20
times more frequently than the next most commonly detected pesticide (Belluck et al.,
1991). Based on 41 land-use studies across the United Sates during 1993-1995, (Kolpin
et al., 1998) indicated that of the 46 pesticide compounds examined, 39 were detected in
shallow groundwater. Atrazine was the most frequently detected (38 0/0) followed by its
decomposition product, deethylatrazine (34 %). The presence of atrazine in groundwater
is a concern because of potential carcinogenic effects ofs-triazines, which include
atrazine, simazine, and propazine (Biradar and Rayburn, 1995). Other toxic effects of
atrazine include skin irritation, respiratory effects, central nervous system disorders, and
blood cell disorders (US. Congress, Office of Technology Assessment, 1984).

Adsorption and desorption processes on soil components are important
phenomena that control the environmental fate and behavior of pesticides including

bioavailabilty, persistence and transport of pesticides in soils and subsoils. Clay minerals

and soil organic matter (SOM) are the two chemically most active components of the
soils. Generally, cultivated soils contain a larger fraction of clay than organic matter, yet
most of the research conducted in the last 20 years has considered SOM as the primary
sorptive phase of organic contaminants interacting with soils (Chiou et al., 1979; Chiou et
al., 1983; Chiou et al., 2000; Huang et al., 1997; Kleineidam et al., 1999; Weber and
Huang, 1996; Xia and Ball, 2000; Xing and Pignatello, 1997). Chiou et a1. (1979) in a
study with seven neutral organic molecules did not observe any indication of adsorption
isotherm curvature even at concentrations approaching saturation. Therefore, they
concluded that adsorption of neutral organic molecules is controlled by soil organic
matter and occurs essentially by partition processes. Chiou et a1. (1983) even suggested
that sorption of non-ionic organic compounds on the soil mineral fraction is relatively
unimportant in moist soils. Xing and Pignatello (1997) in a study with chlorinated
benzenes, observed that the value of Freundlich coefficient, N, was unaffected by
complete removal of mineral phase, again indicating that soil organic matter controls the
sorption of such compounds in soils. It has also been found that in addition to the organic
carbon content, the chemical and structural properties of soil organic matter also
significantly affect the sorption of Slightly polar and non-polar hydrophobic organic
contaminants (Grathwohl, 1990; McGinley et al., 1993; Weber et al., 1992; Young and
Weber, 1995). Geologically mature organic matter had much higher sorption capacities
than relatively young humic organic matter in soils (Grathwohl, 1990; Young and Weber.
1995). Kleineidam et a1. (1999) in a study with low organic matter soils observed that the

reduction of organic carbon significantly decreased sorption. They therefore concluded

that organic matter dominates sorption of hydrophobic organic compounds, even in
samples with very low organic carbon contents.

Clay minerals are layered aluminosilicates, most of which have structural
negative charges due to cation substitution and surface areas exceeding 105 in2 kg71
(Borchardt, 1989; Brown et al., 1978). The large surface area per unit weight associated
with clay minerals is important for sorption of pesticides (Karickhoff, 1984). Among
commonly found clay minerals, expandable 2:1 layer silicate clays are especially
important due to their high surface areas, high cation exchange capacities (C EC 3) and
surface reactivities (McBride, 1994). Smectites are important members of the
expandable 2:1 layer silicate clays and contribute much of the inorganic surface areas of
many soils. Therefore, smectites have a high potential for influencing the fate of
pesticides in the soil environment. The sorption capacity of smectites for atrazine varies
widely depending on the surface charge density of the smectite (Laird et al., 1992), the
nature of the adsorbed cation (Sheng et al., 2001), and the pH of the soil (Vaz et al.,
1997)

Hayes and Mingelgrin (1991) reported that in low organic matter soils and
sediments, clays might control the sorption/desorption interactions. Khan et al. (1979)
observed that inorganic constituents could have a dominant role in pesticide retention in
low organic matter soils. Means et al. (1982) reported that pesticide sorption is
proportional to the percent organic carbon and where percent organic carbon is low, the
amount of clay becomes important in determining pesticide sorption. Karickhoff (1984)
concluded that at clay to organic matter ratios greater than 30, the soil mineral phase

becomes important for the organic contaminant sorption. Rebhun et al. (1992) reported

that sorption of organic contaminants by the organic fraction was 8 to 20 times higher
than by pure clay minerals. However, in soils with the commonly encountered range of
organic matter (0.5 — 5 %) and high clay content, the contribution of the mineral fraction
exceeded that of the organic component. In a study on the adsorption of some
nitroaromatics by clay minerals, Haderlein and Schwarzenbach (1993) and Haderlein et
a]. (1996) found that clays with high surface areas (e. g., smectites) and exchangeable
cations with low hydration energies (e.g., Kr and CS+) promoted adsorption. Boyd et al.
(2001) demonstrated that the strong sorption of nitroaromatiCs occurs because the organic
compound forms interlayer complexes with exchangeable cations of the clay. Recently,
Sheng et a1. (2001) reported that on a unit mass basis, some smectite clays were more
effective sorbents than SOM for four out of the seven pesticides studied. Hance (1965)
suggested that under field conditions in the absence of excess water, the soil mineral
fraction may play an important role in adsorption. Similar results were observed for
adsorption of parathion (Yaron and Saltzman, 1972), and parathion and lindane (C hiou et
al., 1985) on soils. In a dry soil—hexane—parathion system the Slightly polar parathion
molecules efficiently compete with the apolar hexane molecules for adsorption Sites
(Yaron and Saltzman, 1972). Generally as the soil water content increases, parathion
adsorption decreases because of the decreasing contribution of the soil mineral fraction.
When soils are fully hydrated, adsorption of neutral organic solutes by soil minerals often
becomes relatively insignificant compared to the uptake by partitioning into soil organic
matter, presumably because water is preferentially adsorbed by minerals (Chiou et al..

1985).

Sorption of atrazine to soil colloids

a. Organic matter and clays

Talbert and F letchall (1965) showed that atrazine adsorption to whole soils was
strongly correlated with both clay content and organic matter content. However, Roy and
Krapac (1994) did not find any correlation between atrazine sorption and clay content.
Terce and Calvet (1977) reported little atrazine sorption by Ca-smectites. Borggaard and
Streibi g (1988) reported that smectites and iron oxides adsorbed little or no atrazine in the
pH range of 4 — 8 and suggested that organic matter is possibly the only atrazine-
adsorbing constituent in many natural soils. Brouwer et al. (1990) observed a logarithmic
increase of atrazine distribution coefficient with organic matter content. Wang et a1.
(1990) repOrted the preferential binding of atrazine to the higher molecular weight
fraction of fulvic acids and proposed weak mechanisms of interaction such as hydrogen
bonding and hydrophobic bonding to interpret their findings. On the basis of NMR
spectroscopic investigation, Welhouse and Bleam (1992) concluded that due to the
separation of charge within the atrazine molecule, atrazine is able to both donate and
accept hydrogen bonds. On the basis Of intermediate values of atrazine’s donating (am).
and accepting (13m) parameters, they suggested that particularly strong complexes could
be expected between atrazine and acid functional groups of humic substances (Welhouse
and Bleam, 1993a) and concluded that hydrogen bonding is an important mechanism for
atrazine adsorption to soil organic matter (Welhouse and Bleam, 1993b). Grundl and
Small (1993) evaluated the role of mineral-phase sorption of atrazine and alachlor by

different sediments and concluded that they were sorbed by both natural organic carbon

(OC) and clays. The critical clay/OC ratios at which mineral-phase sorption accounted
for 50 % of the overall sorption was 62 and 84 for atrazine and alachlor, respectively. In
a study with a Webster soil, Laird et al. (1994) found that organic and inorganic
components were 11 and 89 % of the mass and contributed 68 and 32 "/0 respectively, of
the affinity for atrazine. Laird et al. (1992) worked in the low-concentration range up to
0.04 mg atrazine/ g clay and reported that atrazine adsorption by 14 different Ca-smectites
ranged from 0 to 100 %. Recently, Sheng et al. (2001) showed that a K-Saturated

smectite and a muck soil sorbed equal amounts of atrazine from aqueous solution.

b. Exchangeable cations

The type of exchangeable cation on the clay also affects pesticide sorption. Sheng
et al. (2001) reported that homoionic K-montmorillonite adsorbed up to 240 mg
atrazine/kg clay and is seven times more effective sorbent for atrazine than homoionic
Ca-montmorillonite. They reported that the lower effectiveness of Ca-montmorillonite
might be due in part to the higher hydration energy and hence larger hydrated radius of
Ca2+ than that of K+. Thus the waters of hydration associated with Ca2+ obscure a greater
portion of the clay surface than those associated with K+, which in turn will shrink the
size of the adsorptive domains between exchangeable cations. They reported that 12.5 A
basal spacing associated with many K-smectites, as opposed to >15 A spacing for C a-
smectites, may be optimal for adsorption of semiplanar aromatic pesticides like atrazine.
Sawhney and Singh (1997) also reported higher atrazine adsorption by Al-smectite (3.82
mg atrazine/ g clay) than by Ca-smectite (1.90 mg atrazine/ g clay). They attributed higher

atrazine sorption by Al-Smeetite to the more polarized water molecules associated with

Al3+ than with Ca2+; those water molecules are postulated to electrostatically attract N-
groups of atrazine. Macnamara and Toth (1970) observed that K-clay systems Showed a
greater adsorption capacity for linuron and malathion than Ca-, Mg-, and H/Al- clay
systems. Using on-line microfiltration and HPLC technique, Gilchrist et al. (1993)
observed that most of the added atrazine was rapidly and reversibly sorbed on the
montmorillonite and that Na+ saturated montmorillonite sorbed more atrazine than Ca2+
saturated montmorillonite. The Ca2+ saturated montmorillonite had a higher surface

charge density than Na+.

c. Surface charge density (SC D)

Laird et al. (1992) worked in the low-concentration range up to 0.04 mg
atrazine/ g clay and reported that atrazine adsorption by 14 different Ca-smectites ranged
from 0 to 100 %. They reported that Panther Creek beidellite adsorbed virtually all of the
added atrazine while Otay, CA montmorillonite did not adsorb any atrazine. The amount
of atrazine adsorbed by the remaining 12 smectites varied between these two extremes.
Panther Creek beidellite had the lowest CBC (88 cmol kg‘l) while Otay, CA
montmorillonite had the highest CEC (134 cmol kg'l), therefore they concluded that
atrazine adsorption is inversely related with the CEC of the smectite clays. They found
that 83 “/0 of the variation in the values of the Freundlich constant (log K) was related to
the surface charge density (SCD) and surface area (SA) of the clays. Panther Creek
beidellite had lower values of SCD (1.42 umol m'z) and SA (619 m2 g") while Otay, CA
montmorillonite had higher SCD (1.85 umol m'z) and SA (724 m2 g") values. In a study

with six different species of K-smectites, Sheng et al. (2001), reported that the amount of

atrazine adsorbed by K-saponite was 0.25 mg/ g clay and by K-montmorillonite was 0.10
mg/ g clay. On the basis of adsorption coefficient values (Kr), they concluded that K-
saponite (CEC 94.9 cmol/kg) is about 15 times more effective in the removal of atrazine
from the solution than K-montmorillonite (CEC 130 cmol/kg). These observations are in
agreement with the results of Laird et a1. (1992) who concluded that atrazine sorption by
clays is inversely dependent on clay properties such as SCD and CEC. Based on this
evidence, they concluded that under neutral pH conditions, atrazine is dominantly sorbed
on smectites as molecular species. Hamaker and Thompson (1972) suggested hydrogen
bonding as a mechanism for the sorption of s-triazine herbicides to clay surfaces. Laird
(1996) reported that under neutral conditions, molecular atrazine is initially sorbed on
smectites by a combination of water-bridging and hydrophobic interactions. He proposed
that the lone pair of electrons on the ring N atoms of atrazine molecules interact with
water molecules solvating metal cations adsorbed on smectite surfaces. At the same
time, the alkyl-side chains of atrazine molecules out-compete water molecules for
retention on hydrophobic microsites on smectite surfaces. Hydrophobic bonding is also
consistent with the observed decrease in sorption with increasing surface charge density
of smectites because the amount of hydrophobic surface on smectites decreases with
increasing surface charge density. One reason for this decrease in sorption with increase
in SCD is that greater SCD causes the water molecules at the surface of the clays to be
more strongly bound and have a more ordered structure (Gilchrist et al., 1993). Since
atrazine is only sparingly water soluble (water solubility is only 30 mg/L), it is less likely
to approach and adsorb to surfaces having rigid, ordered water layers than to surfaces

without such layers. Barriuso et a1. (1994) reported that in spite of the lower adsorption

capacity of high SCD clays, atrazine was held by stronger binding mechanisms on these

clays.

d. pH

The extent of pesticide adsorption by clays depends strongly on the molecular
structure including the basicity of the pesticides and thus on the pH of the solution, with
maximum adsorption typically occurring at a pH near the pKa values of the pesticides
(Weber, 1970). Atrazine has pKa of 1.68, so maximum adsorption of atrazine is
expected in very acidic soils. Sorption of atrazine on acidic humic acid (pH 2.5) was ten
times greater (Kd 627 cm3 g'l) than on neutral (pH 7.0) humic acid (Kd 62.2 cm3 g")
(McGlamery and Slife, 1966). This pH effect was attributed to increased ionic bonding
caused by protonation of the amino groups on the atrazine molecule at low pH
(McGlamery and Slife, 1966). Weber (1970) concluded that the adsorption of s-triazines
was due to complexation of the triazine molecules with functional groups on the organic
colloids and/or adsorption of s-triazine cations by ion-exchange forces. Gilmour and
Coleman (1971) showed that cation exchange causes a large increase in atrazine sorption
to humic acid at low pH. Charge transfer complexes have also been postulated as a
possible mechanism involved in the adsorption of s-triazine herbicide onto SOM
(Hamaker and Thompson, 1972). Borggaard and Streibig (1988) reported that smectites
and iron oxides adsorbed no or insignificant atrazine in the pH range of 4 — 8. Madsen et
a1. (2000) reported that sorption of atrazine depends primarily on pH, the sorption being
high at low pH (<6.7) compared with sorption at higher pH (>7.4). At lower pH, more

hydrogen ions associate with the atrazine molecules giving more cationic characteristics,

which leads to increased adsorption (Frissel and Bolt, 1962). Baskaran et al. (1996)
reported that as the pH increases, the proportion of atrazine cations in the solution
decreases and hence the decrease in sorption with increase in pH. Clay et al. (1988) also
reported that with the application of ammonium sulfate and elemental S as fertilizers to
the soils, the adsorption of atrazine increased as the fertilizers decreased soil pH from 6.0
to 4.5. Atrazine is an extremely weak base (pKa of 1.68), hence little protonation of
adsorptive phases can be expected at pH values 2 or more units above the pKa of the
adsorbate (Hayes and Mingelgrin, 1991). Therefore, atrazine should be sorbed primarily
as a neutral molecule in cultivated soils, which have a pH range of 4 — 8. Laird et al.
(1992) also reported that atrazine is dominantly sorbed on smectites as a neutral species

for the pH range 4.75 to 6.45.

e. Temperature

Harris and Warren (1964) reported that adsorption of atrazine, monuron, and
simazine by bentonite was greater at 0°C than at 50°C. Adsorption by muck was similar
at the two temperatures for all compounds. They concluded that bentonite had a higher
bonding enthalpy than muck. Talbert and Fletchall (1965) also reported that the amount
of atrazine adsorbed decreased as the temperature was raised from 0 to 50 °C. Huang et
al. (1984) did not observe any effect of temperature on the adsorption of atrazine by
whole soils over a temperature range of 5 to 25°C, while Lavy (1970) reported that
diffusion coefficients of many triazines decreased when temperature was lowered from
25 to 5°C, coincident with decreased adsorption. The effect of temperature on the

sorption of atrazine on humic acid was quite marked; sorption was nearly twice as great

11

at 40 °C than at 0.5 °C (McGlamery and Slife, 1966). This endothermicity would imply
that sorption of atrazine to soil organic matter is governed by favorable entropy changes.
Dao and Lavy (1978) reported that greater amounts of atrazine were adsorbed at 30°C
than at 5 °C on four soils at 0.1 bar moisture content. They also reported that a decrease
in water: soil ratio and in soil moisture content led to an increased adsorption of atrazine.
Thus, the influence of temperature on the adsorption by soils seems to be quite variable
and it seems to vary with the nature of the soil components (Harris and Warren, 1964).
The available data imply that sorption of atrazine to clay minerals is enthalpically

favored, while sorption to organic matter is entropically favored.

f. Fe and Al oxides

Huang et al. (1984) showed that oxides of iron and aluminum play an important
role in the adsorption of atrazine by soils and that their removal drastically decreased the
adsorption of atrazine by the soils. Terce and Calvet (1977) also reported similar results,
while Seta and Karathanasis (1997), and Laird et al. (1994) reported that high contents of
iron- and aluminum- oxyhydroxides suppressed atrazine adsorption in soils. This
suppression could be due to the inhibition of smectite swelling by these oxyhydroxides.
Laird et al. (1994) also observed that these free iron- and aluminum- oxyhydroxides coat
silicate minerals, rendering their surfaces more hydrophilic and thus suppressing sorption
of hydrophobic compounds like atrazine.

Atrazine adsorption-desorption is the major process affecting atrazine behavior in
soils and is affected by various factors such as organic matter, clay content, pH,

temperature, and iron- and aluminum- oxyhydroxides. We can see that there are many

12

discrepancies in the literature on the effect of inorganic fractions of soils, so atrazine
adsorption to minerals needs to be further studied with new methodologies. Molecular
Simulations are ideally suited to contribute to our understanding of pesticide behavior in
soils because they simultaneously provide mechanistic insight (that can be compared with
the available spectroscopic information) and kinetic and thermodynamic estimates (that
can be correlated with the many bulk measurements of pesticide diffusion, adsorption
isotherms, and distribution coefficients). Thus, we propose to study the atrazine

interactions with smectites using molecular simulation methods.

Proposal

A. Simulating atrazine-clay interactions

To simulate the adsorption of atrazine by clays, several types of computational
studies will be performed. The minimum-energy structure for atrazine will be computed
using density functional theory (Kohn et al., 1996) as implemented in the ab initio
programs PQS and Gaussian98 (Frisch et al., 1998). The B3LYP functionals (Becke,
1993) were used with the 6-311G** basis set (Krishnan et al., 1980). For the optimized
structure, partial charges for each atom will be estimated using the CHELPG method
(Breneman and Wiberg, 1990). Models for K-SWy-2 and K-SAz-l will be created and
molecular dynamics simulations of atrazine-clay complexes will be performed in order to
explore the interactions between atrazine, cations and water in the interlayer region. The
model for K-SWy-2 will have composition of K7(A142Mg6)(Si%Al1)Oz40(OH)43 with 3

CBC of 79 cmol kg'l, while the model for K-SAz-l will have composition of

K12(Al3ng.2)Si%0240(OH)43 with a CEC of 132 cmol kg". Atrazine will be added to
each unit cell, with the atrazine molecular plane perpendicular to the basal planes of the
clay. The interactions between atrazine, a variety of interlayer cations, and water in the
interlayer region will be observed at different loading rates, that is, different numbers of
atrazine molecules will be added to the unit cell. Water molecules will also be added (in
random interlayer positions) so that constant-pressure simulations result in clay layer
spacings corresponding to 1, 2, or 3 discrete layers of interlayer water. Target dom—
spacings for these simulations will be about 12.5, 15.5, and 18.5 A, respectively. The
energies will be computed using a force field deveIOped especially for clays (Teppen et
al., 1997) combined with the pcff force field (Maple et al., 1994; MSI, 2000) for the
organics and water. The molecular dynamics simulation will be run in the NPT ensemble
for 0.1 ns, much longer than the time required for the system volume and energy to

equilibrate.

8. Experimental studies of atrazine-clay interactions

Adsorption of atrazine on smectites has been studied by a number of workers, but
most of these studies have been limited to montmorillonite, beidellite, and nontronite, all
of which are 2:1 dioctahedral clays. Only a limited number of studies have been
conducted with 2:1 trioctahedral clays. Also, there have been very few studies that
systematically investigated atrazine sorption to smectites as a function of the saturating
cation, as has been shown to be so important in the sorption of other neutral organics
(Sheng et al., 2002). These limited studies indicate that 2:1 trioctahedral clays Often have

higher adsorption capacities for atrazine than 2:1 dioctahedral clays. For example, Laird

l4

et al. (1992) worked with 14 different clays and found that beidellite, hectorite, and
saponite sorbed almost all of the added atrazine. We propose to conduct a laboratory
experiment that will study the adsorption of atrazine by saponite, a 2:1 trioctahedral clay,
saturated with Na+, K+, Cs+, and Ca2+ on the exchange complex. Depending on the
results, we may extend this study to include other triazines: Simazine and metribuzin are
two other pesticides belonging to the class of triazines, but with very different water
solubilities than atrazine. Simazine is only slightly soluble in water (5 mg/L), while
metribuzin has a solubility of 1200 mg/L. Due to differences in water solubility, these
triazines are expected to sorb differently in soils. If the saponite results are interesting,
we may also use another clay mineral such as hectorite or beidellite to study the
adsorption of these triazines: Both hectorite (a trioctahedral smectite like saponite, but
differing in that hectorite has dominantly octahedral substitution) and beidellite
(dominated by tetrahedral substitution, like saponite, but differing in that beidellite is
dioctahedral) showed strong affinity for atrazine in Ca-saturated systems (Laird et al.,

1992).

C. Prediction of organic-clay interactions

Many of the organic pollutants in the enviromnent were introduced years or
sometimes decades ago at a time when industry and the public were not adequately aware
of the scope, magnitude, and importance of environmental pollution (Alexander, 2000).
Thousands of chemicals are manufactured world-over and many new chemicals are
added to this ever expanding list. Many of these chemicals end up in the environment

(e. g. soil, groundwater, and surface water etc.), where they are a possible source of

15

concern. Adsorption and transport of many of these chemicals, especially pesticides,
have been studied for more than 40 years. However, it would require unfeasible financial
and scientific resources to study the partition constants of such an immense range of
chemicals with an equally high number of different natural phases. Thus there is interest
in predictive models that can estimate the potential for a given organic compound to
partition into a given sorptive phase (G035 and Schwarzenbach, 2001). We propose to
use molecular dynamic simulations to estimate the adsorption energies of some chemicals
that have been heavily studied from an experimental standpoint, and then correlate these
adsorption energies with the adsorption coefficients obtained from the experiments. If
the computed adsorption energies follow the experimental trends in adsorption
coefficients, then we will have demonstrated the potential for molecular simulations to
predict the sorption of neutral organic compounds to clay minerals.

To estimate the adsorption energies, we will simulate a series of nitroaromatic
compounds interacting with smectite clay minerals as studied in a recent experimental
dataset (Boyd et al., 2001). This experimental study is very useful for validation
purposes because it reports sorption of 14 nitrobenzene compounds, differing only in the
identity or location of a second functional group. A model for the K-saturated SAz-l
smectite clay mineral used in the experimental study will be created, with composition

K12(Al36Mg12)Si960240(OH)48 and a resultant cation-exchange capacity (C EC ) of

132 cmol/kg. In order to prepare for molecular dynamics Simulations of nitrobenzene-
clay complexes, the interlayer region of the clay will be expanded to 20 A and one
molecule from a series of substituted nitrobenzenes will be added to the unit cell. with the

nitrobenzene molecular plane perpendicular to the basal planes of the clay. Water

16

molecules will then be added (in random interlayer positions) and molecular dynamics
performed on each clay-water-organic system. The energies and forces will be computed
using a force field developed especially for clays (Teppen et al., 1997) and combined
with the pcff force field (Maple et al., 1994; M81, 2000) for the organics and water. liach
molecular dynamics Simulations will be run in the NPT ensemble for at least 50 ps. which
is about twice as long as required for the system volume and energy to equilibrate. If the
initial Simulation does not result in a clay layer Spacing of 12.3 A. then water molecules
will either be removed or added (after expanding the clay) and the simulation re-run so
that the equilibrium clay layer Spacing is 12.3 i 0.1 A, in accord with the experimental
data for substituted nitrobenzenes in air-dried films of K-SAz-l (Sheng et al.. 2001;
Sheng et al., 2002). A trajectory of each simulation, containing equally spaced snapshots
of geometries, energies, and velocities, will be recorded for later analysis.

We will use the following algorithm to calculate the clay-organic interaction
energy.

E=E

m1

,0, — El — E2; where
°E,O, = Total clay—organic system energy
0E] = Energy of system after organic has been deleted

°E2 = Energy of the system after clay and water have been deleted

°Compute Em, for at least 150 snapshots and average

Electrostatic and van der Waal components Of Eml will be calculated the same way, in
order to estimate the contributions of each component to the overall sorption energy. This

clay-organic interaction energy (Em) is only a portion of the total adsorption energy, but

in the following we also propose a method for using molecular Simulations to estimate

the overall adsorption energy.

The typical experimental scheme for measuring total adsorption energy is

Claytaq) + orgm) t::> clay-orgtaq) + HZOtaql
Eexpt

where Eexp, is the experimentally determined adsorption energy (for example, the overall
free energy, enthalpy, or entropy of the transfer process). G055 and Schwarzenbach
(2001) illustrated that the overall free energy change of transfer AG, , could be first
decomposed into individual terms for a more mechanistic understanding of the partition

process as Shown in Figure 1.

l8

 

Partitioning of small organics between a clay phase and an aqueous phase
Clay phase Aqueous phase

   
   

AG.- = E. + Bitty mystic?“ ’3‘”

Figure 1. Schematic for decomposing the energy changes for transfer

of a compound from aqueous solution to the clay layer
(after Goss and Schwarzenbach, 2001).

 

 

 

Since the experimental reaction is too large and too slow to directly model with
molecular simulation methods, an alternative path is proposed that is more
computationally feasible and follows the spirit of Figure 1. We propose to employ the

following cycle for computing the total adsorption energy:

Claym) + orgtaq) => clay-orguq) + H200”,
Eexpt

Clay-cavity(,q) + orgm) + H20 => clay-organ) + cavitym) + H20
E4

This approach is similar to many schemes (i.e., see review in (Halperin et al., 2002)) that
have been used to compare molecular simulations with experimental data for the binding
of drugs to polypeptides. Since the free energy, enthalpy, and entropy are state functions

(Atkins, 1998), their magnitudes should be independent of path, so that

Eexpt : E3 + E4 + E5
WhCI‘C E4 = Ei-clay—Ei-aqueous = E(clay-cav —> clay-org) + E(orglaq) ——->0rgca\'(aq))

E4 = Em: + E(org(aq> —> ca\'ily(aq))

E3 is the energy for creating a cavity in the clay, by removing enough water
molecules to make a cavity that can later receive the organic. E3 will be calculated in the
same way as the algorithm above for Em, using the smectite with a 12.3 A d-spacing and

only cations and water in the interlayer.

E4 is similar to the sum of two components as shown in Figure 1. While we are
not using the gas-phase configuration that Goss and Schwarzenbach (2001) used, our
term E4 can be estimated the same way. The first term in E4 is Em, the energy for filling
the cavity in the clay with the organic molecule. The second term in E4 is the energy for
moving the organic solute out of the aqueous phase, thereby leaving an empty cavity in

water.

E5 is the energy for filling this aqueous cavity with water, and E5 will again be

calculated using the algorithm for Eim that was presented above.

20

We will use all of the nitroaromatic compounds from the experimental database
(Boyd et al., 2001) and calculate the clay-organic interaction energies for each compound
using the algorithm above for Eim. Then we will fit these interaction energy values into
the larger context, using the thermodynamic cycle given above to estimate the overall
adsorption enthalpies for each compound. The enthalpy estimates will be checked against
experimental data in two ways. In relative terms, we hypothesize that the enthalpy
estimates will follow the same trends as the sorption coefficients for the nitroaromatic
compounds (Boyd et al., 2001). In absolute terms, we hypothesize that our enthalpy
estimates will compare favorably with experimental enthalpies recently estimated from
the temperature dependence of adsorption isotherms (Li et al., 2004). If predictive
capability can be demonstrated by using these nitroaromatic compounds as a training set,
then we will use the methods to estimate the enthalpies of adsorption for atrazine in

smectite clay minerals.

21

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Sawhney, BL, and SS. Singh. 1997. Sorption of atrazine by Al- and C a-saturated
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Seta, AK, and AD. Karathanasis. 1997. Atrazine adsorption by soil colloids and C 0-
transport through subsurface environments. Soil Science Society of America
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Sheng, G.Y., C.T. Johnston, B.J. Teppen, and SA. Boyd. 2001. Potential contributions of
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Sheng, G.Y., C.T. Johnston, B.J. Teppen, and SA. Boyd. 2002. Adsorption of
dinitrophenol herbicides from water by montmorillonites. Clays and Clay
Minerals 50:25—34.

Talbert, R.E., and OH. Fletchall. 1965. The adsorption of some s-triazines in soils.
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Vaz, C.M.P., S. Crestana, S.A.S. Machado, L.H. Mazo, and LA. Avaca. 1997.
Adsorption isotherms for atrazine on soils measured by differential pulse
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188.

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27

Weber, W.J., and W.L. Huang. 1996. A distributed reactivity model for sorption by soils
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888.

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Environmental Science & Technology 2621955-1962.

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Environmental Science & Technology 29:92-97.

28

CHAPTER II

TRIAZINE ADSORPTION BY SAPONITE AND BEIDELLITE CLAY

MINERALS

29

Triazine adsorption by saponite and beidellite clay minerals

Abstract

We investigated the sorption of three triazine herbicides (atrazine, simazine, and
metribuzin) by saponite and beidellite clay minerals saturated with K+, Cs+, Na+, and
Ca2+. Saponite clay sorbed a larger fraction of each pesticide from aqueous solution than
did beidellite clay. The lower cation exchange capacity in saponite (vs. beidellite)
presumably results in a less crowded interlayer with more siloxane surface available for
adsorption. Generally, Cs-saturated clays sorbed more triazines than clays saturated by
K+, Na+ or Ca2+. We attribute this to the smaller hydrated radius of C 8+, which again
increases the siloxane surface available for sorption. Furthermore, the relatively weak
hydration of Cs+ reduces the swelling of clay interlayers, thus making sorption domains
less hydrated and more receptive to hydrophobic molecules. C S-saponite manifested a
sorption of more than 1% atrazine by weight above equilibrium concentrations of 6 mg/L.
This is the largest sorption of neutral atrazine from water yet reported for an inorganic
sorbent. Molecular dynamics Simulations indicate that atrazine interacts with both clay

basal planes and with multiple cations in the clay interlayer.

30

Introduction

Concern about the contamination of ground and surface waters by pesticides has
been an area of concern for over thirty years, mainly because of the continuous use of
such chemicals in agricultural production. There is a need to more accurately predict the
fate and movement of pesticides and assess their risks in soil and subsurface
environments. Sorption by soils/sediments is a major process that controls the transport,
bioavailability, degradation and ultimately the fate of organic contaminants in the
environment (Koskinen and Harper, 1990). Sorption of organic chemicals diminishes
their bioavailability to target organisms (Albanis et al., 1988), affects both chemical and
microbiological transformations (Armstrong and Chesters, 1968; Gamble and Khan,
1985), and reduces the leaching to groundwater (Sonon and Schwab, 1995).

Clay minerals and soil organic matter (SOM) are typically the two most
chemically active components interacting with organic chemicals that enter soils and
sediments. SOM has been accepted as the primary sorptive domain for the uptake of
hydrophobic organic compounds (Chiou et al., 1979; Chiou et al., 2000; Huang et al.,
1997; Kleineidam et al., 1999; Weber and Huang, 1996; Xia and Ball, 2000). The
contaminant transport models often use soil-organic carbon-normalized sorption
coefficients (Koc) to predict the distribution of organic chemicals between soil and water.
This approach ignores the contribution of soil minerals to the retention of organic
contaminants in soils and subsoils, based on the assumption that organic compounds
cannot effectively compete with water molecules for the sorptive sites on mineral

surfaces.

31

Clay minerals are commonly found in the soil environment, and smectites are the
most important group for interacting with pesticides due to their high surface areas,
moderate cation exchange capacities (CECS), and surface reactivities (McBride, 1994).
Earlier studies have shown that smectite clays can adsorb many major types of pesticides
such as carbamates, ureas, nitrophenols and triazines to an appreciable extent (Li et al.,
2004; Li et al., 2003; Sheng et al., 2001). In a study of atrazine sorption by a series of
Ca-smectite clays with CECs ranging from 79 to 134 cmolc/kg, Laird et al. (1992)
reported that zero to nearly 100% of added atrazine was removed from water, with the
extent of adsorption correlated inversely with the clay CEC values. Sheng et al. (2001)
observed a similar relationship for sorption of atrazine by K-smectites. Sawhney and
Singh (1997) reported that atrazine adsorption by Al-smectite (>1 g kg!) was larger than
that by Ca-smectite. The authors discounted protonation of atrazine and attributed the
greater adsorption on Al-clay to the stronger H-bonding between atrazine and the more
polarized water molecules associated with trivalent Al3+ versus divalent Ca2+. 1n the
present study, we have avoided Al-Smectites due to the difficulty in Simultaneously
maintaining a homoionic Al-smectite while preventing side-reactions in which atrazine
protonates and consequently adsorbs strongly by cation exchange.

Clays present in whole soils also play a major role in adsorbing triazine pesticides
(Huang et al., 1984; Talbert and Fletchall, 1965). Talbert and F letchall (1965) Showed
that atrazine sorption by soils was positively correlated with soil clay content as well as
SOM content. Huang et al. (1984) reported that the inorganic fractions of soils with

particle size <20 um have a significant capacity to retain atrazine. Several previous

studies have attempted to identify the critical ratio of clay minerals/SOM at which

32

sorption by the mineral phase plays an important role (Grundl and Small, 1993;
Karickhoff, 1984; Means et al., 1982). Karickhoff (1984) concluded that at the ratio of
swelling clays to soil organic carbon >30 the mineral phase in whole soils contributed
significantly to sorption of simazine. Grundl and Small (1993) estimated that at a mass
ratio near 62 for clay to organic carbon, the mineral phase accounted for half of the
overall atrazine sorption.

Most studies of atrazine and simazine adsorption by smectites were conducted
using smectites with negative charges originating from octahedral substitution (Gilchrist
et al., 1993; Sannino et al., 1999; Sheng et al., 2001); few studies have used smectites
with negative charges arising from tetrahedral sites. Also, investigations of the effect of
the location of permanent negative charges on organic sorption by smectites are scarce.
Compared with octahedrally substituted smectites, the negative charges in tetrahedrally
substituted smectites are more localized to fewer oxygens on siloxane Sheets (Farmer,
1978; McBride, 1989), resulting in differences in clay interlayer environments that may
influence pesticide sorption. Beidellite and saponite clays are smectites with negative
charges originating largely from tetrahedral substitution. This study was, therefore,
initiated to study the adsorption of three s-triazine herbicides (atrazine, simazine, and
metribuzin) by saponite and beidellite clays saturated with different cations (KI, Na+, C 5+
and Ca2+), and to compare the results with previous studies on sorption by octahedrally
substituted smectites. K+, Na+ and Ca2+ are common exchangeable cations in the soil
environment, while Cs+ has smaller hydrated radius relative to other cations studied here

hence is expected to create a more favorable (i.e., less hydrated) clay interlayer

environment for pesticide adsorption.

33

Materials and Methods

Clays and Herbicides

Two reference clays, saponite (SapCa-Z, CEC 95 cmol/kg (Post, 1984)) and
beidellite (SBId-l, CEC 129 cmol/kg (Hetzel and Doner, 1993)), were obtained from the
Clay Minerals Society Source Clay Repository (Department of Agronomy, Purdue
University, West Lafayette, Indiana, USA). The <2 urn clay-sized particles were
separated by wet sedimentation, and then exchanged with K+, Na+, CS+, and C a2+ to
prepare homoionic clays. To do so, 25 g of clay was dispersed in a 0.1 M chloride
solution (500 ml) with the respective cation. The clay suspensions were shaken for 24 h,
and fresh chloride salt solutions were used to replace the original solutions after
centrifugation. This process was repeated four times to ensure complete cation
saturation. The excess chlorides were removed by repeatedly washing with Milli-Q
water until Cl’ was determined negative by reacting with AgNO; solution. The clay
suspensions were then quick-frozen, freeze-dried and stored in closed containers for later
use. The pH values of the smectite suspensions were 8.1 and 8.0 for Ca-saponite and
-beidellite, respectively, 8.6 and 8.8 for the Na-smectites, 8.1 and 8.0 for the C s-
smectites, and 8.2 and 8.0 for the K-smectites.

The three pesticides (atrazine, simazine, and metribuzin) used in the adsorption
experiment were purchased from ChemService (West Chestnut, Pennsylvania) with
purities of >99%, and used as received. The selected physical and chemical properties of

the three pesticides along with their chemical structures are summarized in Table 1.

34

 

Z

 

Axum luv ,2
_ __
. . . z/ \oluxfuv
owm co _ cm? 0 _ om Em xw ENBEDE
O
fmozz iwszxwt zzfmc
z z \ z 5
Womm omm w mo." 8. SN JO MEMBER
5
x z /
m:Nozz AK wizzfmu
2,, \z
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gsaem

 

Siam zeta—com £5 5 com: moEocmom we wear—2.8m _ao_Eo:ooooo_m.Em tote—om A 03:.

Given the pKa values of the herbicides and the high pH values of our clay
suspensions,there was no possibility for triazine sorption as any cationic species in the

present study. Rather, the triazines were all neutral, moderately hydrophobic solutes.

Sorption Isotherm Measurements

Pesticide sorption isotherms were determined for each clay using a batch
equilibration method. A series of initial pesticide concentrations was prepared by
dissolving each pesticide in 0.01 M KCl, NaCl, and CSCl, and 0.005 M CaClz. The initial
solutions were then added into glass centrifuge tubes containing clay (0.01 to 0.1 g) that
was saturated with the same cation in the respective pesticide initial solution, and the
tubes were closed with Teflon-lined screw caps. Duplicate samples were prepared to
measure sorption. The ratio of solution volume to clay amount was optimized to make
sure that 20-80 % of pesticide in solution phase was sorbed by clay during the
equilibration. The tubes were then Shaken reciprocally for 12 hours at 40 rpm at room
temperature (23 i 2 °C); preliminary studies showed sorption was complete within this
time. The tubes were then centrifuged at 2110 g for 20 min to separate the liquid and
solid phases. The concentration of herbicide in supernatant was analyzed using a
reversed-phase high-performance liquid chromatography (HPLC, Perkin-Elmer,
Norwalk, CT) set at the appropriate wavelength for maximum absorption by each
pesticide (Table 1). A platinum extended polar selectivity (EPS) C18 column was used.
The mobile phase was a mixture of either 65% methanol (for atrazine and metribuzin) or
65% acetonitrile (for simazine) and water with a flow rate of 1.0 mL/min. Controls

consisted of the initial pesticide solutions in the supporting electrolyte. Abiotic

36

degradation of atrazine and Simazine in water at pH 7 to 9 is negligible (Comber, 1999;
Noblet et al., 1996) in the absence of strong UV light. However, due to the relatively
high pH values in the clay slurries and the lack of information about possible smectite-
catalyzed hydrolysis, we monitored the HPLC chromatograms for peaks indicative of
triazine degradation products. None were observed, and there were no significant
differences between peaks or shapes of chromatograms for triazine controls versus
triazine in clay suspensions. Also, no changes in solute concentrations were detected in
the tubes devoid of clay within the experimental period; therefore solute mass lost in the
supernatant from clay slurries was assumed to be sorbed by the clay. The amount of
pesticide sorbed on the clays was calculated from the difference between the amount
added and that remaining in the final equilibrated solution. Data presented are the means

of duplicate samples.

X-ray Diffraction

After supematants were sampled for HPLC analysis, the remaining pesticide-clay
suspensions were used to prepare clay films for X-ray diffraction analysis. The clay
solids were resuspended with the supernatant (~ 1 to 2 mL) remaining in the vial, and
then dropped onto glass slides and air-dried overnight at ambient conditions to Obtain
oriented clay films. X-ray diffraction patterns were recorded using Cu-KOL radiation and
a Philips APD 3720 automated X-ray diffractiometer, fitted with an APD 3521
goniometer and a 0-compensating slit, a 0.2 mm receiving slit and a diffracted beam

graphite monochromator. Diffraction patterns were recorded from 5.5 to 16.0 °20 at a

step of 0.005 °20 with 4 s/step. The X-ray diffraction patterns were re-recorded after the

37

clay films were allowed to equilibrate with 100% relative humidity for 2 d in a closed

container.

Molecular Simulations

Molecular dynamics simulations of atrazine-clay complexation were performed in
order to explore possible interaction mechanism(s) between atrazine, cations, and water
in the clay interlayer regions. In this approach, electrons are not treated explicitly, but
their effects are parameterized into classical (that is, non-quantum) mechanical energy
functions comprising a force field. The model for K-saponite had composition
K9(Sig7A19)Mg720240(OH)48 with a CEC of 95 cmol/kg. The smectite surfaces were
constructed using methods previously detailed (Teppen et al., 1997), in which a
muscovite mica layer was the starting point and Al-for-Si substitutions were random
except that no two A1 tetrahedral were allowed to be immediate neighbors. One molecule
of atrazine was added to the unit cell, with the atrazine molecular plane perpendicular to
the basal planes of the clay. The simulated loading rate was equivalent to 23 mg atrazine
per g clay. Water molecules were also added (in random interlayer positions) so that
constant-pressure simulations resulted in ~12.5 A clay layer spacings, in accord with our
experimental data for atrazine in films of K-saponite.

The energies were computed using a force field developed for clays (Teppen et
al., 1997) combined with the polymer consistent force field (PCFF) (Maple et al., 1994;
M81, 2000) for the organics and water. AS a further modification, charges on the organic
solute were assigned using the COMPASS force field (Sun, 1998), which optimized

organic atomic charges for condensed phase systems. In previous work (Boyd et al.,

38

2001; Sheng et al., 2002), we have found COMPASS charges to more faithfully
reproduce quantum mechanical charges for nitrogenous organic molecules than do PC F F
charges. In aqueous interfacial systems, the charges on atoms of organic solutes play a
critical role in determining organic solute interactions with the clay and interlayer water
and cations. In classical molecular dynamics, the nuclear charge and the local electron
density are subsumed into one partial atomic charge and these charges for atrazine are

illustrated in Figure l.

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

('31
C1
N1/ \\N1
CII (I: Charge
H / 2\ ¢ 2‘.‘ H1 C1 +0.5830
‘\N’,.-" N1 ‘\.\ / C2 +0.5637
7— H N2 C3 +0.1048
2 C4 -0.1590
H
2\c3/ H2 C5 +0.1578
H -
2\C /C4—H2 Nl 0.4810
5 \ N2 -0.5422
C4 H2 H1 +0.2487
2 C4—H2 C1 -01020
Hz /\
H2

Figure 1. Charges on atrazine molecule used in simulation study.

39

Commercial software (M81, 2000) was used to compute the classical molecular
dynamics. No constraints were imposed on our systems during their molecular dynamics
except for the contents of the simulation cell: All atoms were completely free to move,
and unit cell shape and volume were unconstrained. The molecular dynamics simulations
were run in the NPT (constant pressure/constant temperature) ensemble for 75 ps, which

is much longer than the time required for the system volume and energy to equilibrate.

Results and Discussion

Sorption Isotherm

Pesticide sorption isotherms are shown in Figure 2. The greatest sorption of
atrazine was observed for Cs-saturated saponite, followed by K-, Na-, and least by C a-
saturated saponite (Figure 2a). This sorption of more than 1% atrazine by weight of clay
at an aqueous atrazine concentration of 6 mg/L is by far the largest sorption of neutral
atrazine from water yet reported for any inorganic material. The higher effectiveness of
Cs-SapCa-2 in sorption may be due to the lower hydration energy of Cs+ and hence
smaller hydrated radius of Cs+ compared to the other cations. The water molecules
associated with exchangeable Ca2+, for example, obscure a larger fraction of the clay
surface than those associated with Cs+, resulting in the reduction in the size of adsorptive
domains for Ca2+ saturated clays. Typical estimates of hydrated radii are 9.6 A for C a”
and 3.6 A for CS+ (Evangelou, 1998). Thus the cross-sectional area of hydrated C a2+ is
about 290 A2, more than seven times that of Cs+(40 A2); therefore, one hydrated Ca2+

takes up 3.5 times as much room as do the two Cs+ ions in clay interlayer. We

40

hypothesize that the CS-smectites may sorb triazines more strongly because the triazines
only have to compete with weakly bound water between hydrated cations, while on C a-,
K-, and Na-smectites the triazines must compete with more strongly bound water within
the hydrated radii of cations. Sheng et al. (2001, 2002) also observed similar effects of
exchangeable cations on the Size of sorptive domains of smectite clays in pesticide

sorption.

41

 

 

 

 

 

 

 

 

 

 

 

 

 

16 - - _-. _,
I (a) Atrazine Saponite I
14 p ' .
12 1 on-
I V
10 1; 0 .
' V 0‘ .
8 ‘1 I 9 .
I v 01 o ‘ l
6 I - ‘
'V o . a
4 I 02 ‘
IV 0‘.
2 I 0009’ - .
o 2 4 O I 10 12 u.
0 In—WMJ‘FEL ,_ ' _,
0 2 4 6 8 10 12 14
a 1.4
El (c) Mctribuzin Saponite
E 1.2 1
V
B y " V
'E 1.0 .
0°: v
g 0.8 1
.2 '
E} 0.6 « .
G- . .
h—
o 0.4 1 ‘ ‘ A AA
‘15: v
V
O . ‘A ~
E 0.2 'v ‘ . I I .
< A, .
0.0 05-1.. .4 . . .
0 10 20 3O 40 50 ()0
2.5
' (e) Simazine Saponite
2.0 1
V
V . K
v
1.5 1 , A Ca
' I Na
v Cs
1.0 1
v'
0.51). . .
I I
A
-. «T
0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8

 

0.8 1

0.6 1

0.4 1

 

0.2

0.0

a.

V

I

(b) Atrazine Betdcllttc I

 

0.7

0.6 1

0.5 1

0.4 1

0.3 1

0.2 1

0.11

 

 

 

0.0

80 l 00

 

0.4

0.3 1

0.2 1

0.11

 

(t) Simazine Beidellite

V
V O
O

 

 

0.0
O

Equilibrium Concentration of Pesticide (mg/L)

Figure 2. Sorption isotherms representing atrazine (a,b), metribuzin (c,d), and
simazine (e,i) uptake from water by reference saponite (SapCa-Z) and beidellite
(SBId-l) clays saturated with CS+, 1C, Na“ and Ca“.

42

Differences in the swelling behavior between Cs- and Ca-SapCa-2 also affect the
degree of pesticide adsorption. The relatively weak hydration of Cs+ causes Cs-saponite
to swell less in water, which again provides a more favorable sorption domain and allows
the organic compound to better compete with waters for interlayer sites. C s-smectites
maintain ~12.3 A basal spacings (monolayer of hydration) at 100% humidity (data not
shown) and even in aqueous suspension (Sato et al., 1992; Suquet and Pezerat, 1987).
Ca-smectites, on the other hand, swell to more than 15 A basal spacings. The thickness of
the ring of an atrazine molecule is ~3.0 A (Sheng et al., 2002). This would imply the
lowest limit of the basal spacing for atrazine intercalation to be about 12.2 A (the
thickness of a clay layer alone for pyrophyllite is 9.2 A). In this scenario, an atrazine
molecule thus contacts both clay siloxane surfaces simultaneously (Sheng et al., 2001 ).
This orientation favors the intercalation of organic contaminants and pesticides, because
it minimizes the contact with water molecules. The efficacy of such a hydrophobic
adsorption mechanism even in inorganic smectites is implied by observations (Li et al.,
2004; Sheng et al., 2002) that intercalated organic solutes inhibit swelling when the
smectite is exposed to water.

Our results for atrazine sorption by saponite clay compare favorably with the
results reported by Sheng et al. (2001) for K-SapCa-Z and Laird et al. (1992) for another
Ca-saponite (CEC = 106 cmol/kg) (Figure 3). For further comparison, isothemTS
measured by Sawhney and Singh (1997) and Celis et al. (1997) for atrazine sorption by
other Ca-saturated reference smectites (CEC = 85 and 76 cmol/kg, respectively) are also

included (Figure 3). These six datasets (except that of Celis et al. (1997)) follow the

43

general trend that, for a given exchangeable cation, atrazine sorption tends to increase as

layer charge decreases (Laird et al., 1992).

 

(I8

Oti-

(14 ~ 0

(12 - v

Amount of atrazine sorbed, mg/g clay

 

 

I I I

O 5 10 15 20

01)

Equilibrium cone. of atrazine, mg/L

 

K-SapCa-l (this study)

K-SapCa-l (Sheng et al., 2001)

Ca-SapCa-l (this study)

Ca-saponite (Laird et al., 1992)
Ca-montmorillonite (Sawhney and Singh, 1997)
Ca-montmorillonite (Celis et al., 1997)

04.0.0

 

 

 

Figure 3. Comparison of our sorption results with all earlier studies of atrazine
sorption to saponites, as well as two representative montmorillonites.

44

The type of smectite clay does affect triazine pesticide adsorption, as illustrated
by the greater atrazine adsorption by saponite than that by beidellite. (Figure 2 a and b).
Sorption of atrazine by Cs-SapCa-2 was 9.2 mg/g at the aqueous concentration of 2
mg/L, and is 37 times the sorption by Cs-SBId-l at the same aqueous concentration.
Similarly, sorption of atrazine by Na- and Ca-SapCa-2 was greater than the sorption by
Na-, and Ca-SBId-l clay, respectively. These observations are again probably due to the
availability of sorptive area on clay surfaces. SBId-l clay has a CEC of 129 cmol/kg
(Hetzel and Doner, 1993) while SapCa-2 clay has a relatively lower CEC (95 cmol/kg,
(Post, 1984)). Generally, clay with a higher CEC retains more hydrated exchangeable
cations that cover a larger portion of clay surfaces, hence results in less siloxane surface
available for pesticide adsorption, as demonstrated by previous studies of an inverse
relationship between clay CBC and pesticide adsorption (Laird et al., 1992; Lee et al.,
1990; Sheng et al., 2001; Sheng et al., 2002). In studies of atrazine adsorption by both
Ca- and K-saturated smectites, the clay with the highest CEC was consistently the least
effective adsorbent for atrazine (Laird et al., 1992; Sheng et al., 2001). Celis et al. (1997)
reported the similar observation that Ca-SAz-l (C EC = 120 cmol/kg) sorbed less atrazine
than Ca-SWy-l (CEC = 76.4 cmol/kg), which is also consistent with the observation that
adsorption of atrazine decreases as clay CEC increases. The current study along with
previous studies supports the hypothesis that exchangeable cations (e.g., Ca2+) with larger
hydration spheres obscure more of the neutral clay siloxane surface thereby reducing
hydrophobic interactions between clay surfaces and pesticides. The lower charge-density
clay (SapCa-Z) manifested greater adsorption compared to the higher charge-density

SBId-l clay. Lower charged clay apparently has larger exposed areas between

45

exchangeable cations on the neutral siloxane surfaces, which contribute more favorable
pesticide adsorptive domains. Crowding interlayer cations upon pesticide adsorption may
increase the electrostatic repulsion among these cations, thereby contributing as well to
the reduction of pesticide adsorption by the higher charge clay.

The site of negative charges on the clay also affects sorption. Permanent negative
charge on clay arises from isomorphic substitution in either tetrahedral or octahedral
Sites. The CEC of SBId-l is 129 cmol/kg with the negative charges arising from mostly
tetrahedral substitution, and SAz-l has a similar CEC (130 cmol/kg) but with the
negative charges primarily from octahedral substitution. Data for atrazine sorption by K-
SAz-l were reported by Sheng et al. (2001), and the atrazine sorbed by these two KI-
saturated clays varied substantially with the much greater sorption for K-SBId-l clay. For
example, the amounts of atrazine sorbed by K-SBld-l and K-SAz-l were 0.79 and 0.04
mg/g, respectively, at an aqueous atrazine concentration of 8 mg/L. Since CEC values
and surfacie areas of the two clays are similar, differences in sorption result primarily
from the sites of negative charges. Compared with the charges from octahedral
substitution (SAz-l), the charges from tetrahedral substitution (SBId-l) are localized to
fewer surface oxygens (Farmer, 1978; McBride, 1989) resulting in more ordered
distribution of exchangeable cations on mineral surfaces (Odom, 1984). These
exchangeable cations (at a given hydration state) can always approach the tetrahedrally-
substituted sites more closely hence demonstrating a stronger attraction compared with
octahedrally-substituted sites (Bleam, 1990). AS a result, SBId-l clay has a smaller clay
layer spacing, which favors atrazine adsorption in clay interlayer environments (C happell

et al., 2005). In addition, the charge localization of tetrahedral substitution results in a

46

larger portion of neutral siloxane surfaces associated with mineral surfaces, which may
provide large neutral surface domains for the adsorption of organic compounds.

The K-smectites studied here exhibit interesting sorption behavior. In contrast to
the trend discussed above, in which sorption generally decreases as clay layer-charge
increases, the higher-charged K-SBId-l sorbs slightly more atrazine than does K-SapCa-
2. Indeed, the sorption by K-SBId-l is equivalent to that of Cs-SBId-l, while sorption by
K-SapCa-Z is much below that of Cs-SapCa-Z. One hypothesis is that some C s-SBId-l
interlayers collapse and completely dehydrate even in suspension, thus rendering them
unavailable for sorption of atrazine. A second hypothesis might be that K-SBId-l sorbs
like Cs-SBId-l because the two are Similarly swollen with ~12.5-A dom-spacings: It is
well-known (McLean and Watson, 1985) that removal of K+ from an illite clay results in
a tetrahedrally substituted clay mineral resembling a high-charged beidellite, while
addition of K+ to these weathered illites results in K+ “fixation” to form a mica-like
mineral with dehydrated interlayers. The tendency for clay minerals to fix K+ correlates
with their layer charge (Sparks and Huang, 1985). While true smectites like K-beidellite
should not completely dehydrate and fix K+ (Sparks and Huang, 1985) because their layer
charges are lower than those of illites, it is plausible that a hi gher-charged beidellite like
K-SBId-l would swell less than a lower-charged smectite like K-SapCa-Z. Thus, despite
its high charge, the interlayers of K-SBId-2 should be more likely to retain duo, -spacings
near 12.5-A when fully hydrated, and thus be relatively favorable for hydrophobic
interactions. Testing of these hypotheses would involve measuring door—spacings of these
smectites in aqueous suspension (Chappell et al., 2005), which we hope to do in the

future. In the present study we measured the dooj-spacings of air-dried films of K-SBld-2

47

and K-SapCa-2 and found them to be 10.2 and 12.0 A, respectively, so interlayers of the
K-beidellite almost completely dehydrate while K-saponite retains one layer of
interlamellar water when air-dried. Thus, it is plausible that the beidellite would swell
less than the saponite in suspension, as well.

The sorption of metribuzin by saponite and beidellite followed the same order as
for atrazine, i.e., Cs-saturated > K-saturated z Ca-saturated > Na-saturated clay, but the
magnitude of metribuzin sorption was smaller than atrazine sorption by both clays
(Figure 2 c and d). This could be due to the higher solubility of metribuzin (1050 mg/L)
compared to atrazine (33 mg/L), Since we hypothesize (Boyd et al., 2001) that
hydrophobic effects contribute to sorption of neutral organic solutes by smectite clays.
Sorption of metribuzin by saponite clay was two to five times (Figure 2c) that by
beidellite (Figure 2 d) at a given metribuzin concentration, which may again be due to the
comparably lower CEC of saponite as discussed above.

The sorption of simazine by these two clays manifested similar results. C s-
SapCa-2 adsorbed appreciably higher simazine than K-, Ca- and Na-SapCa-2 (Figure 2
e), and beidellite sorbed considerably less simazine than saponite clay. This again shows
that a smectite clay with a lower CEC and/or saturated with a less hydrated exchangeable
cation (i.e., CS+) usually demonstrates a greater sorption capacity. Simazine sorption by
K-SBId-l behaved like atrazine sorption in that it was almost identical to the sorption by
Cs-SBId-l.

If hydrophobic effects do contribute to triazine adsorption by smectites, then at a
given aqueous triazine concentration, the sorption to a given clay system should have

followed the order simazine a: atrazine > metribuzin. Simazine and atrazine are difficult

48

to rank, because atrazine has a slightly higher aqueous solubility (presumably because it
is less symmetric) yet also has a larger octanol-water partition coefficient because it
contains one more methylene group (Table 1). For the Cs-saponite at 0.1 ppm aqueous
triazine, the adsorbed amounts of triazine were 2.0, 2.6, and 0.02 mg/g, for simazine,
atrazine, and metribuzin, respectively (Figure 2). For K-saponite at 1.3 ppm aqueous
triazine, the adsorbed amounts of triazine were 0.3, 0.1, and 0.03 mg/g, respectively.
Similarly, for K-beidellite at 0.5 ppm aqueous triazine, the adsorbed amounts were 0.2,
0.2, and 0.01 mg/ g, respectively. Thus, our limited data are consistent with a
hydrophobic component to triazine adsorption by smectites, which implies that one role
of the clay interlayer is to partially remove the hydrophobic solute from water (Boyd et
al., 2001). Interlayers that perform this role most effectively, such as C s-saponite, should

then be the strongest clay mineral sorbent phases.

X -ray Diffraction

As the atrazine loading increased, the door-spacing (corresponding to the centroid
of the diffraction peak) for the air-dried Cs-smectite increased to 12.5 A for the 13.7 g/kg
atrazine loading, from 12.1 A for the atrazine-free clay (Figure 4). The increase in basal
spacing suggests that atrazine is intercalated in the interlamellar region of the clay, with
an increasing fraction of the clay domains at larger d-spacings (~12.5 A) as the atrazine
loading increases. This is consistent with our previous observations of intercalated
aromatic compounds (Li et al., 2004), in which the dom-spaeings of air-dried K-smectite
clay films increase slowly toward ~12.5 A as the aromatic compound loading increases

from zero to 200 umol/g clay. Here, the maximum atrazine loading rate is about 65

49

umol/g clay (Figure 4), so sorption is rather small to have a major impact on the d001-
spacing (Li et al., 2004), but the systematically increasing dam-spacings are consistent
with our hypothesis of interlayer sorption of atrazine. The sorption of metribuzin and
simazine did not result in a change in clay basal spacing compared with that in the
absence of organics (data not presented). The reason is that the amount of sorbed
metribuzin and Simazine in clay interlayer was not sufficient to cause the swelling of clay
basal spacing. Li et al. (2004) also reported that at a low amount of organic solute

loadings the clay interlayer did not Show the expansion.

 

 

 

 

 

12.5 A
12.4 A
12.2 A
12.1 A
.1? r 13-7 mg/g
8 .
8 1
.5
8.8 mg/g
0.4 mg/g
"“w -- er:- :va
Blank
4 6 8 10 12
2 Theta

Figure 4. X-ray diffraction patterns of air-dried Cs-SapCa-2 clay films as a function
of atrazine sorption.

50

Molecular Simulations

The molecular dynamic simulation results offer structural hypotheses for atrazine
interactions in the interlayer region of K-SapCa-2. Figure 5 is a representative snapshot
taken from a molecular dynamics simulation. As such, it is not a minimum energy
structure, but is representative of the dynamic equilibrium. The types of interactions
shown in Figure 5 are comparable to others obtained during some 20 similar simulations
we have done for atrazine in smectite interlayers. Within a monolayer of interlayer water
between the mineral surfaces, atrazine is oriented parallel to the clay basal plane with the
aromatic ring interacting with both of the opposing clay siloxane sheets. In this
simulated configuration, the negatively charged ring-N atoms of atrazine also form inner-
and outer-sphere complexes with multiple K+ ions. The inner-sphere complex shown is a
chemically reasonable structure because a K+ ion is able to interact with two (ring N and
C 1 substituents) negatively charged (Figure l) and closely spaced sites simultaneously.

Our molecular dynamic simulations also support the hypothetical interactions
between atrazine and smectite siloxane surfaces that were proposed by Laird (1996): He
proposed that the lone pair electrons on the ring N atoms of atrazine interact with water
molecules solvating exchangeable cations associated with smectite surfaces; such
complexes were indeed observed in our simulations (Figure 5). At the same time, the
alkyl-side chains of atrazine molecules were hypothesized (Laird, 1996) to out-compete
water molecules for retention on hydrophobic microsites on smectite surfaces. Again,

this behavior is observed in the simulations (Figure 5). Such hydrophobic bonding is

51

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52

consistent with the observed decrease in sorption with increasing clay CEC because the
amount of neutral siloxane surface on smectites decreases with increasing surface charge
density (SCD). On surfaces with low SCD the water molecules at the clay

surface are less strongly bound and considered to have a less ordered structure (Gilchrist
et al., 1993). Sparingly water-soluble organic compounds (e. g., atrazine, water solubility
of 30 mg/L) are more likely to approach and adsorb to surfaces that sorb water less
strongly. A lower SCD also means that there are less exchangeable cations competing
for adsorption domains in the interlayer region (Sheng et al., 2002).

In summary, we observed very strong atrazine sorption by Cs-saponite, and a
lesser sorption by K-, Ca-, and Na-saponites. These results are another indication of
interplay between the polar functional groups of neutral organic molecules and the
hydration of interlayer cations. As the organic functional group becomes more polar, the
ability to displace water from clay interlayers becomes stronger, resulting in a substantial
adsorption. Thus, pesticides with multiple strongly polar functional groups like —N03 —
C=O or -CEN demonstrate a strong affiliation with K- and Cs-smectites (Boyd et al.,
2001; Johnston et al., 2002; Sheng et al., 2002) due to pesticide-cation complexation.
However, our results indicate that even less polar compounds (e. g., atrazine) can be
strongly adsorbed by Cs-saponite, due to the optimal clay interlayer structures, i.e.,
maintaining ~12.5 A basal spacing and non-overlapping radii of hydration for C s+.

Thus, Cs-saponite interlayers contain the largest proportion of relatively hydrophobic
sites, in which we hypothesize that triazines can be at least partially dehydrated, and our
data are consistent with a hydrophobic component to triazine adsorption by smectites.

We also found that the site of negative charge on the clay significantly influenced the

53

retention of pesticide. The smectite clays with the negative charge arising from
tetrahedral substitution generally displayed a higher adsorption compared to smectites
with octahedral substitution, presumably because tetrahedrally substituted smectites swell
less in water and thereby contain narrower slit-pores that favor hydrophobic sorption.
However, if the clay layer charge is large, adsorption of triazine herbicides by even
tetrahedrally substituted smectites can be sharply reduced, because hydration radii for the
greater number of exchangeable cations begin to overlap. Apparently, an optimal
inorganic sorbent for triazine herbicides should be a Cs+-saturated smectite with a low
layer charge resulting from tetrahedral substitution. These criteria maximize adsorption
domains parallel to the clay surfaces while optimizing (near 12.5 A) the adsorption
domains perpendicular to the clay surfaces. If true, these hypotheses should apply

broadly to hydrophobic organic solutes.

54

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benzene, toluene, and xylene by two tetrarnethylarnmonium-smectites having
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Li, H., G.Y. Sheng, B.J. Teppen, C.T. Johnston, and SA. Boyd. 2003. Sorption and
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57

Sheng, G.Y., C.T. Johnston, B.J. Teppen, and SA. Boyd. 2001. Potential contributions of
smectite clays and organic matter to pesticide retention in soils. Journal of
Agricultural and Food Chemistry 49:2899-2907.

Sheng, G.Y., C.T. Johnston, B.J. Teppen, and SA. Boyd. 2002. Adsorption of
dinitrophenol herbicides from water by montmorillonites. Clays and Clay
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58

CHAPTER III

ATRAZINE INTERACTIONS WITH SMECTITE CLAYS: MOLECULAR

SIMULATION APPROACH

59

Atrazine interactions with smectite clays: Molecular Simulation approach

Introduction

Atrazine is one of the most commonly used agricultural herbicides in the world
and has been applied for more than 40 years by com and sorghum growers in the United
States (Ribaudo and Bouzaher, 1994). In 1994, over 47 million pounds of atrazine were
applied nation-wide covering about 77 % of the area cropped under com-soybean rotation
(Pike and Knake, 1996). Atrazine is also one of the most common herbicide residues
found in groundwaters of Midwest (Blanchard and Donald, 1997; Burkart et al., 1999;
Kolpin et al., 1998), and is detected 10 to 20 times more frequently than the next most
commonly detected pesticide (Belluck et al., 1991). The presence of atrazine in
groundwater is a concern because atrazine is a potential carcinogen (Biradar and
Rayburn, 1995), and a suspected endocrine disrupter. Toxicological concerns have
resulted in the EPA setting a maximum amount of 3 ug/L of atrazine in drinking water in
the US. Despite restrictions, atrazine exceeds the maximum contaminant level of 3ppb
at certain times in several states in the US.

It has been shown that adsorption of organics on smectite clays is an important
phenomena that controls the bioavailability, persistence and movement of organics in
soils and subsoils (Boyd et al., 2001; Laird et al., 1992; Li et al., 2003; Sheng et al.,
2001). The sorption capacity of smectites for organics varies widely depending on the

surface charge density of the smectite (Laird et al., 1992), the nature of the adsorbed

60

cation (Sheng et al., 2001), the pH of the soil (Vaz et al., 1997), and the location of
negative charge on the smectite.

In parallel with atrazine usage in agriculture, researchers have studied atrazine
adsorption by soils and soil components for more than 40 years. Due to its long-term
agronomic importance and its leaching risk, atrazine sorption processes in soils have
probably been studied more than those of any other organic solute. Most studies, of
course, have been phenomenological in nature, striving to characterize the extent of
sorption and correlate sorption with known properties of atrazine and its related triazines,
with soil properties, and with solution conditions such as pH. However, there has been
continual and intense interest in mechanistically understanding the forces that drive
triazine sorption.

Already by 1962, Frissel and Bolt (1962) described triazine sorption to
montmorillonite as a cation exchange reaction within a few pH units of each triazine pKa.
Furthermore, they speculated that >90% of triazine molecules would be adsorbed by
clay-rich soils, even at neutral pH, through the action of van der Waals forces between
triazines and clay siloxane surfaces (Frissel and Bolt, 1962). Weber et al. (1965) used X-
ray diffraction to Show that prometone (a methoxy— rather than chloro-triazine, with pKa
of 4.3) intercalated between clay layers. Sorption of 150 to 300 umol/g caused the dry
clay to swell to ~13 A, consistent with triazine rings oriented parallel to the Silicate sheet.
Sorption of 450 umol/g caused clay swelling to ~18 A, and the authors used molecular
models to postulate triazine molecules either perpendicular to the clay surfaces or
arranged in double stacks in the interlayer. Most sorption to the montmorillonite was by

cation exchange, since the experiments were done just 1.7 pH units above the pKa,

61

though the authors postulated neutral species sorption via H-bonding of triazines to
siloxane surface 0 atoms of the clay (Weber et al., 1965).

Bailey et al. (1968) studied sorption of six triazines by Na-montmorillonite and
found that sorption was generally larger for the more soluble triazines, the opposite of the
expected behavior if a nonspecific, hydrophobic adsorption mechanism were operant
(Hance, 1967). However, triazine cation formation can only be discounted for the three
low-pKa triazines studied (including atrazine), and for these compounds there was not a
consistent relationship between sorption and aqueous solubility (Bailey et al., 1968). For
neutral triazines interacting with clay minerals, they hypothesized that possible
adsorption mechanisms might include van der Waals interactions and H-bonding between
amino N-H groups and clay siloxane O atoms.

For many years, though, the dominant mechanism for atrazine adsorption by clay
minerals was thought to be cation exchange, because no sorption of uncharged
(“molecular”) atrazine by clays was conclusively demonstrated. However, Laird et al.
(1992) argued that, at neutral pH, molecular atrazine was the species sorbing to a variety
of smectites, and they showed that adsorption was inversely correlated to smectite layer
charge. In a later study, Barriuso et al. (1994) measured desorption from many of these
same smectites, showed that there was little hysteresis at work, and concluded that
adsorption of molecular atrazine to smectites is generally reversible. As a result, they
hypothesized that the atrazine sorption mechanisms were mainly van der Waals or
hydrogen bonding to smectite surfaces.

The finding that atrazine adsorption was inversely correlated with smectite layer

charge (Laird et al., 1992) implied a hydrophobic sorption mechanism, as had been

62

 

asserted by Jaynes and Boyd (1991) to explain their observations that benzene adsorption
by smectites increased with decreasing clay layer charge. Expandable clay minerals are
known to be strongly hydrophilic on a macroscopic scale (Mooney et al., 1952; Xu et al.,
2000). However, on a molecular scale there are uncharged regions between the charged
sites on smectite siloxane surfaces, so smectites with lower layer charges will have a
greater proportion of uncharged surface. J aynes and Boyd (1991) showed that, when
smectites are saturated by organic rather than inorganic cations, the uncharged regions of
the smectite surface will preferentially adsorb benzene versus water from aqueous
solution, so they described these uncharged regions as hydrophobic. In smectites, these
hydrophobic regions might comprise more than 40% of the total surface area (Laird,
1996).

In other contexts, the sorption of nonpolar, nonionic organic solutes is known to
be controlled by “hydrophobic effects,” a combination of relatively small van der Waals
bonding forces and a substantial entropic gradient that drives the organic molecules out
of aqueous solution (Israelachvili, 1992; Tanford, 1980; Voice and Weber, 1983).
Previous studies (Boyd et al., 2001; Curtiss et al., 1986; Laird and Fleming, 1999; Mader
et al., 1997; Piatt et al., 1996; Rebhun et al., 1992; Schwarzenbach and Westall, 1981;
Van Blade] and Moreale, 1974) have observed that the association of organic solutes with
an inorganic surface increased with decreasing aqueous solubility of the compounds.

Laird (1996) described the reversible sorption of molecular atrazine as a
hydrophobic process, in which the hydrophobic alkyl side-chains of atrazine interact with

the hydrophobic microsites located between smectite charge sites. At the same time, he

63

hypothesized (Laird, 1996), N atoms of the triazine ring may accept H-bonds from water
molecules that solvate nearby cations, further stabilizing the atrazine-clay complex.

Laird (1996) predicted that, if the hydrophobic sorption mechanism for atrazine
on smectites was true, then triazine sorption on montmorillonite should increase with the
length of the alkyl side-chains, which had been seen in earlier data (Bailey et al., 1968;
Weber, 1970) but erroneously ascribed to pKa differences. Laird and Fleming (1999)
tested this hypothesis by comparing the sorption of molecular pyridine and butylpyridine
at high pH: The sorption of butylpyridine was far larger than that of pyridine, thus
supporting the concept of hydrophobic contributions to organic adsorption on smectites.

Spectroscopic characterization of atrazine sorption mechanisms has been limited
to organic matter as a sorbent phase (Chien and Bleam, 1997; Welhouse and Bleam,
1992; Welhouse and Bleam, 1993), and we are aware of no spectroscopic studies of
molecular atrazine in clays. Protonation (Cruz et al., 1968) and hydrolysis (Russell et al.,
1968) of triazines on smectites have been studied by FTIR spectroscopy, but not the
bonding environment of the neutral triazine.

Since atrazine adsorbs to clays as the molecular species at neutral pH (Laird et al.,
1992), we wish to understand its mechanism of bonding. It is difficult to visualize the
possible interactions between clay surfaces, water, and organics at the molecular level.
Molecular dynamics simulations can help in understanding these interactions by
providing an objective, energy-based approach to modeling the structural arrangements
likely to occur and persist in smectite interlayers. Ideally, simulation results would be
coupled with other molecular-scale data such as vibrational, NMR, neutron scattering, or

EXAF S spectroscopies (Teppen et al., 1997) in order to employ multiple perspectives.

64

In the present case, we only have bulk adsorption data plus X-ray diffraction data to
constrain the smectite interlayer spacing. The objective of the present study, therefore,
was to couple molecular dynamics simulations with the available data in order to explore

the behavior of atrazine on hydrated smectite clay surfaces.

Materials and Methods

Molecular dynamics simulations of atrazine-clay complexes in the interlayer
region were performed to explore the interactions between atrazine, exchangeable
cations, and water. In this approach, electrons are not treated explicitly, but their effects
are parameterized into classical (that is, non-quantum) mechanical energy functions
comprising a force field.

Models for three different smectites were created, representing a Wyoming
montmorillonite (lower charge due to octahedral substitution), an Arizona
montmorillonite (higher charge due to octahedral substitution), and a California saponite
(moderate charge due to tetrahedral substitution). The three correspond to the reference
clay minerals SWy-2, SAz-l, and SapCa-2, respectively, for which we have atrazine
sorption data. Homoionic versions of each smectite with K+, CS+, or Ca2+ in the
interlayer were simulated. The composition of the model K-saturated Wyoming smectite
(K—SWy) was K7(A142Mg6)(Si95A11)Ozao(OH)48 , while that of Cs-smectite (Cs-SWy) was
Cs7(A142Mg6)(S195A11)0240(OH)43. The CEC of our model K-SWy was 79 cmol kg", and
the comparable experimental value is 83 cmol kg'l (Sposito et al., 1983). We also created
models for K- and Ca-saturated Arizona smectite (SAz-l). The composition of this

model K-smectite (K-SAz) was K12(Al3(,Mg12)(Si%)0240(OH)48 . while that of C a-

65

smectite (Ca-SAz) was Ca6(A136Mg12)(Si96)Ozao(OH)43. The K-SAz was a high-charged
smectite, with a modeled CEC of 132 cmol kg’l, which compares with the experimental
value of 130 cmol kg'1 (Laird et al., 1992). The model for K-saponite had composition
K9(S137A19)Mg720240(OH)43 with a CEC of 95 cmol/kg, in agreement with its
experimental CEC (Post, 1984).

The smectite surfaces were constructed using methods previously detailed
(Teppen et al., 1997), in which the muscovite mica structure (Brigatti and Guggenheim,
2002) was the starting point and Al-for-Si or Mg-for-Al substitutions were made to
charge the layers. One molecule of atrazine was added to each unit cell of the smectite,
with the atrazine ring perpendicular to the basal plane of the clay. The simulated loading
rate was therefore ~24 mg atrazine per gram clay for each atrazine-smectite system.

This is larger than our experimental loading rates, which are all less than 14 mg/ g (see
Chapter 2), but lower simulated loading rates would have required larger unit cells and
therefore much slower simulations. As will be shown below, the present unit cell is large
enough to allow simulation of relatively isolated atrazine molecules, while small enough
to allow many different systems to be sampled. Water molecules were then added (in
random interlayer positions) so that constant-pressure simulations resulted in 12.5, 15.5
or 18.5 A clay layer spacings. One example of an initial configuration is shown in

Figure 1.

66

 

Figure l. A typical starting configuration before beginning the molecular dynamics
simulation. Here, atrazine has been placed into the interlayer of Cs-SWy-2 with 55
water molecules. The equilibrium dam-spacing will eventually be 12.3 A.

The energies and forces were computed using a force field developed for clays
(Teppen et al., 1997) combined with the polymer consistent force field or pcff (Maple et
al., 1994; M81, 2000) for the organics and water. As a further modification, charges on
the atrazine were assigned using the COMPASS force field (Sun, 1998), which optimized
atrazine atomic charges for the condensed phase systems. In previous work (Boyd et al.,
2001; Sheng et al., 2002) we have found COMPASS charges to more faithfully reproduce
quantum mechanical charges for nitrogenous organic molecules than do PCFF charges. In
aqueous interfacial systems, the charges on the atoms of organic solutes play a critical
role in determining organic solute interactions with the clay and interlayer water and

cations. In classical molecular dynamics, the nuclear charge and the local electron density

67

are subsumed into one partial atomic charge and these charges for atrazine are illustrated

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

inFigure2.
Tl
C1
N1/ \\N1
H ’I’r’ / ““ H arge
‘\N2/ N1 \ / 1 Cl +0.583O
H N2 C2 +0.5637
2 C +01048
a \ . .
2\ 3/ /H7- Ca 01590
H2\c,—--——C4—H2 C5 +0.1578
\ N. -O.4810
c4 / 112 N2 -05422
H2// \H2 H1 +0.24g7
H2 C4—H2 H2 +0.0530
H2 \H C1 01020
2

Figure 2. Partial atomic charges on the atoms of the atrazine molecule, as used in
all simulation studies. The total charge on the molecule sums to zero.

Commercial software (M81, 2000) was used to compute the classical molecular
dynamics, with a timestep of 0.2 fs. Ewald summation was used for both electrostatic
and van der Waals forces (Karasawa and Goddard, 1989). No constraints were imposed
on our systems during their molecular dynamics except for the contents of the simulation
cell: All atoms were completely free to move, and unit cell shape and volume were
unconstrained. The molecular dynamics simulations were run in the NPT (constant

number of atoms, pressure, and temperature) ensemble for at least 50 ps, much longer

68

 

than the time required for the system volume and energy to equilibrate. The unit cell
volume equilibrates primarily through changes in the dew-spacing, which is directly
comparable to X-ray diffraction data. Note that the unit cell has a constant composition.
We used periodic boundary conditions, which allow our small unit cell to more faithfully
represent a condensed phase. As such, atoms are free to diffuse out of the central unit
cell, but a copy of the atom will simultaneously diffuse in from the opposite side,

maintaining the constant composition.

Results and Discussion

Within approximately 5 ps of simulation time, the volume of the atrazine-clay-
water systems was typically equilibrated, while the energy was usually equilibrated
within the first 10 ps. Thus, the final 40 ps of each Simulation were used as a sampling of
the equilibrium dynamics of the model system. In many cases, systems of the same
composition (or varying by one to a few water molecules) were simulated multiple times
and the equilibrium structures were quite consistent. The molecular dynamic simulation
results offer structural hypotheses for atrazine interactions in the interlayer regions of
three smectites, saturated with three different cations, at three different water contents.
Representative snapshots of the various equilibrated atrazine-water-clay systems taken
from molecular dynamics simulations are shown in Figures 3-10. These snapshots are
not minimum energy structures, but are representative of the dynamic equilibrium.

Atrazine is thought to interact with smectite surfaces through a variety of bonding
mechanisms (Laird, 1996; Laird and Sawhney, 2002). Our molecular simulations

revealed possible orientations of atrazine in the interlayers of our simulated smectites

69

 

with one, two, and three layers of interlayer water molecules. Figure 3 allows for
comparison of the effect of water content on the atrazine molecular environment in K-
SWy-2. The swelling of the smectite is a critical factor controlling atrazine adsorption:
Chappell et al. (2005) recently showed that a hydrated K-smectite consisting of
interstratified monolayers and bilayers (i.e., Figure 3A and 3B) adsorbed more than ten
times the atrazine as did the same K-smectite when swelled to a trilayer (i.e., Figure 3C).
The authors (Chappell et al., 2005) used drying-induced hysteresis to cause swelling
differences, and were able to measure the K-smectite dam-spacings in the suspension
phase, so these spacings represent the interlayer hydration at which atrazine sorption
actually occurs. These recent measurements (Chappell et al., 2005) of the effect of clay
swelling on atrazine adsorption help to explain previous discoveries that addition of NHa+
fertilizer salts to soils increases atrazine adsorption (Clay et al., 1998) and that decreasing
the soil water content leads to increased atrazine sorption (Dao and Lavy, 1978).
Structural trends for the monolayer, bilayer, and trilayer cases and their relation to

differences in sorption will be discussed in more detail below.

70

 

4.»

Ii. q. V. a- 1, I!
1’&' :L'i' t'ir' I'a'é' 'ir

   
  

 

 

71

 

Figure 3. Snapshot showing the orientation of atrazine within the interlayer of a K-
saturated smectite like SWy-2 containing (A) one, (B) two, and (C) three layers of
hydration water (red = oxygen, white = hydrogen, blue = nitrogen, gray = carbon,
green = interlayer K+ cations). Images on the left depict a side-view (x,z plane) of
the atrazine-smectite interlayer complex while the images on the right depict the
same from a top-view (x,y plane), having removed the silicate layers. The simulated
dam-spacings for A, B, and C are 12.4, 15.5, and 17.3 A, respectively.

Figure 4 illustrates atrazine structural relationships to cations and water within the
interlayer regions of four different simulated smectites, all with door-spacings near 12.5
A. In such smectites containing a monolayer of interlayer water, atrazine is necessarily
oriented parallel to the basal surfaces and interacts directly (no intervening water
molecules) and simultaneously with both opposing basal surfaces. In this simulated
configuration, with its narrow door-spacing and limited volume to accommodate atrazine
and several cations, the negatively charged ring-N and -C1 atoms of atrazine could hardly

escape contact with multiple K+ (or CS+) ions.

 

 

 

 

 

 

 

 

 

 

 

Figure 4. Snapshots of molecular dynamic simulations of atrazine in the interlayer
regions of four different homoionic smectites, (A) K-SWy-Z, (B) Cs-SWy-Z, (C) K-
SAz-l, and (D) K-SapCa-Z. Clay layers were parallel to the page and have been
removed in order to expose the interlayer complexes. The atom coloring scheme is
K = green, 0 = red, H = white, C =gray, N =b1ue, and CI = green (associated with
the atrazine molecule).

73

In the snapshots of Figure 4, which are typical of their respective simulations, the
negatively charged ring-N and -Cl atoms of atrazine thus tended to form both inner- and
outer-sphere complexes with multiple interlayer cations.

The —C1 substituent of the atrazine ring often coordinated directly to a single K+
(or Cs+) ion (Figure 4) with average K (or Cs-) -Cl distances of 3.2 A :t 0.3 A (for K-
SAz), 3.9 d: 0.2 A (for Cs-SWy), 4.2 :1: 0.2 A (for K-SWy), and 4.0 :h 0.1 A (for K-
SapCa-Z). For example, in the Cs-SWy—2 system, approximately 5 ps of simulation time
passed before Cs+ and -Cl approached each other to form their inner sphere complex.

Once formed, this complex remained stable over the subsequent 45 ps of simulated time

(Figure 5).

 

 

o I 7 I I I
0 1 0 20 30 40 50

 

 

 

 

Figure 5. Trajectory of the Cs — Cl inner sphere nonbonded complex in the
simulated atrazine-Cs-SWy-Z system represented by Figure 4B.

74

Secondly, it was also typical that each ring -N- atom adjacent to the —Cl
substituent of atrazine coordinated directly (inner-sphere) to a single KI (or Cs+) ion
(Figure 4), with an average K (or Cs-) -N distance of 2.7 A d: 0.2 A (for K-SAz), 3.1 i
0.2 A (for Cs-SWy), 3.7 a 0.2 A (for K-SWy), and 3.2 :1: 0.2 A (for K-SapCa-2). The
ring —N- atom in the 5-position seems to be hindered from engaging in such complexes
by the alkyl substituents in the 4- and 6-positions. Our molecular dynamic simulations
also support the hypothetical interactions between atrazine and smectite siloxane surfaces
that were proposed by Laird (1996): He proposed that the lone pair electrons on the ring
N atoms of atrazine interact with water molecules solvating exchangeable interlayer
cations; such complexes were indeed observed in our simulations. Transient states
existed in which ring —N atoms (in the 1- and 3-positions) were indirectly coordinated to
K+ (or CS+) ions through the intermediation of water. For example, the outer-sphere C s --
N complex in our model SWy-2 (Figure 4B) exhibited average distance of 5.4 i 0.3 A.

Figure 4 shows that, for three of the four Cs- and K—smectites that we simulated, it
was typical for two atrazine functional group sites (ring-N and -C1) to simultaneously
interact with one interlayer cation. These inner-sphere complexes seem to be chemically
reasonable structures because a cation is able to interact with two negatively charged (-
0.48 and -0.10 from Figure 2) and closely spaced sites simultaneously. We were
somewhat surprised to see such complexes form, because 0 atoms of water have charges
of -0.82 in our system and therefore ought to out-compete atrazine for coordination sites
around cations. However, we see cations “complexed” to -Cl, -N- atoms, and even

methyl groups of atrazine (Figure 4).

75

 

The steric conditions of crowded interlayers at 12.5-A door-spacing may play a
role in formation of such “complexes.” For example, we have seen that the K -- Cl
complexes in SAz-l and SWy-2 had lengths of 3.2 A i 0.3 A and 4.2 :t 0.2 A,
respectively. Also, K -- N distances were 2.7 A d: 0.2 A and 3.7 i 0.2 A, respectively.
Part of the force driving the K-atrazine complexes closer in SAz-l may be mutual
repulsions of all the cations, since it is a higher-charged interlayer environment than is
SWy-2. It is visually apparent (Figure 4C) that atrazine has disrupted the average
spacing of the 12 K+ ions per unit cell in SAz-l, so another reason for shorter K-atrazine
complexes may simply be that atrazine is interfering with optimal interactions between
interlayer cations and charged sites in the clay layers. In the simulations, we force
atrazine into the K-SAz-l interlayer, but in reality it does not need to adsorb. Thus
(Sheng et al., 2001), K-SAz-l (Figure 4C) adsorbs only about 1/6"1 the atrazine as does
K-SWy-2 (Figure 4A), presumably because the former has higher charge, so has a more
hydrophilic interlayer, so atrazine has a harder time competing with water for interlayer
sites (Laird, 1996). Indeed, Laird et al. (1992) showed that atarazine sorption by Ca-
smectites was inversely correlated with smectite layer charge.

Figures 4A and B compare the environment of atrazine in K- versus Cs-SWy-2
montmorillonite. Note that K+ cations near atrazine (Figure 4A) are more likely to have
water molecules competing for them, disrupting the K-atrazine complexes, while C s+
cations (Figure 4B) are able to form two inner-sphere complexes with atrazine. This is
consistent with the less favorable hydration energy of Cs+. As a result, the Cs-smectite
interlayer is less hydrophilic and atrazine competes better for interlayer sites. For

example, using typical estimates (Evangelou, 1998) for the hydrated radii of C 8+ (3.6 A)

76

 

and K+ (5.3 A), along with an assumed 700 mz/g basal surface area and monolayer
hydration status, we compute that only 56% of the Cs-SWy-2 interlayer surface is
occupied by hydrated cations while 122% of K-SWy-2 is occupied. Thus, more of the
relatively hydrophobic surface should be available for atrazine sorption in the C s-system.
We know (Chapter 2) that Cs-SapCa-2 sorbs about 50 times as much atrazine as K-
SapCa-Z, and the saponite’s layer charge is close to that of SWy-2, but we do not yet
have the sorption data for Cs-SWy-2.

Given Laird’s (Laird et al., 1992 and Laird, 1996) hypothesis that lower-charged

smectites adsorb more atrazine because the surface charge sites are more widely

 

separated, so a larger fraction of the smectite surface is hydrophobic, we looked to see if
indeed atrazine sorbs to hydrophobic nanosites in the smectite interlayers. Figure 6
shows that most of the atrazine molecule occupies a more-hydrophobic surface site
between charge sites in K-SWy-2. However, given the rather regular charge spacing in
our model SWy-2, atrazine cannot quite fit completely between charged sites. For the
Cs-SWy-2 (Figure 6), the atrazine ring resides almost directly over a negative site but the

two alkyl sidechains occupy more hydrophobic sites.

77

 

 

 

 

 

 

 

Figure 6. Representative snapshot of the relationships of atrazine and cations to
charged sites in model SWy-2 with a monolayer of water for the K— (left) and Cs-
saturated (right) cases. Blue octahedra are the locations of Mg-for-Al octahedral
substitution sites and pink tetrahedral are the locations of Al-for-Si tetrahedral
substitution sites in the montmorillonite layer. The upper clay layer and the rest of
the lower clay layer have been removed for clarity.

For the more expanded smectites with two or three layers of interlayer water
molecules, fewer atrazine-cation complexes formed, and the atrazine typically inclined at
an acute angle relative to the basal surfaces (Figure 3B and 3C). The two alkyl side
chains tended to interact directly with one basal surface each. The rest of the atrazine
molecule was forced to interact with interlayer water molecules.

Figure 7 shows snapshots from four simulations of atrazine in the interlayers of
hydrated smectites (both SWy and 8A2), each containing a bilayer of interlayer water
between the mineral surfaces. The equilibrium door—spacings of these systems were 15.5

i 0.1A, which is a realistic interlayer spacing for smectites having two layers of

78

A

:IW'IJEH *2 {“2
u "tii
QIp-j‘"~1 4. ti

 

 

 

 

E:

 

 

‘

 

VI

 

Figure 7. Snapshots of molecular dynamic simulations of atrazine in the interlayer
regions of four different homoionic smectites, (A) K-SAz-l, (B) Ca-SAz-l, (C) K-
SWy-2, and (D) Cs-SWy-Z. Clay layers were parallel to the page and have been
removed in order to expose the interlayer complexes. The atom coloring scheme is
K = green, 0 = red, H = white, C =gray, N =b1ue, and CI = green (associated with

the atrazine molecule).

79

interlayer water. In this environment atrazine did not orient its triazine ring parallel to the
clay surface during 50 ps of simulation time. In the Ca-saturated SAz-l (Figure 7B), the
negatively charged functional groups of atrazine formed only outer-sphere complexes
with interlayer Ca2+ ions.

In the K-SWy-2 bilayer system, the negatively charged ring-N atoms of atrazine
were again able to form both inner- and outer-sphere complexes with K+ ions. For
example, the distance between one of the interlayer K and ring-N atoms of atrazine
remained stable around 3.1 i 0.3A over almost all of the simulation time (Figure 8). For
the CS-SWy-Z bilayer system, just one interlayer Cs+ formed an inner-sphere complex
with a ring-N atom of the atrazine, compared with two inner-sphere and one outer-sphere

complex in the monolayer Cs-hydrate (Figure 4B).

 

 

 

0 1 0 20 30 40 50

 

 

 

Figure 8. Trajectory of the K - N inner sphere nonbonded complex in the
simulated atrazine-K-SWy-2 system represented by Figure 3B and 7C.

80

Again, it is of interest to observe the spatial relations between adsorbed atrazine,
interlayer cations, and charge substitution sites in the clay layers. Figure 9 shows that, in
the K-SWy-2 bilayer system, one alkyl chain of atrazine lies in one water layer (on one
basal surface) and the other alkyl chain lies in the second layer of interlayer water (on the
opposite basal surface). For the Cs-SWy bilayer system, too, the two alkyl side chains

interact directly (no intervening water molecules) with one basal surface each (Figure 9).

 

‘I I:“"" C‘Kf“ ‘5. Sta F
A 1r it A 1r 1t

.; ‘83.wa I gist-7 71:84
1:50;.(3 #313011

«A 'afit 1,...

\

   
 
    

 

 

 

 

 

Figure 9. Representative snapshots of the relationships of atrazine and cations to
charged sites in model SWy-2 with a bilayer of water for the K- (left) and Cs-
saturated (right) cases. Blue octahedra are the locations of Mg-for-AI octahedral
substitution sites and pink tetrahedral are the locations of AI-for-Si tetrahedral
substitution sites in the montmorillonite layer. The lower clay layer and the rest of
the upper clay layer have been removed for clarity.

Inspection of the molecular simulation results indicates that the two alkyl chains of
atrazine tend to interact with uncharged portions of the montmorillonite basal surface,
just as hypothesized by Laird (1996) and Laird and Fleming (1999). Perhaps if the

smectite charge distribution were more heterogeneous so that some larger uncharged

81

domains were created, atrazine would be stable when adsorbed flat on one surface. This
could be a subject for future simulations.

Finally, further simulations of atrazine in the interlayer space of SWy-2 or SAz-l
modeled even wetter conditions. Here, a trilayer of interlayer water formed between the
mineral surfaces. The equilibrium door-spacing was 18.5 i 0.1A for Ca-SAz-l and 17.3 :t
0.1A for K—SWy-2. These are also realistic interlayer spacings for smectites having three
layers of interlayer water. In this environment atrazine is again unable to orient parallel
to the clay surface during the 50-ps simulations. Rather, the situation is similar to the
bilayer case, with atrazine inclined at an even more acute angle relative to the basal
surfaces, and the two alkyl side chains interacting directly with one basal surface each
(Figure 10). This configuration in K-SWy-2 appears to be at least partially stabilized by
H-bonds between amino hydrogens and the smectite basal oxygen atoms; two such bonds
of 2.5 and 2.6 A formed between one amino group and one basal surface and were then
stable for the remainder of the simulation. The rest of the atrazine molecule is forced to
interact with interlayer water molecules. As the amount of interlayer water increases,
neither K+ ions nor interlayer Ca2+ ions are able to form inner-sphere complexes any
longer. Apparently, when atrazine and the cations are not forced together into a severely

limited space as in the monolayer case, water outcompetes atrazine for interlayer ions.

82

A B
5"“: ----- cf; N"’a%-7)'..

a
let.

‘—

   

.‘
P.J. , r .1}: e i“
it'- 1‘1 ’13:»
if! 2" 1.3, “‘ w .n r E
'_ “Mow-.. -----.|.v- ”Pkg! ----- J":- IA— 3":

Figure 10. Snapshots of molecular dynamic simulations of atrazine in the interlayer
regions of two different homoionic smectites, (A) Ca-SAz-l, and (B) K-SWy-Z. Clay
layers were parallel to the page and have been removed in order to expose the
interlayer complexes. The atom coloring scheme is K = green, 0 = red, H = white, C
=gray, N =b1ue, and C1 = green (associated with the atrazine molecule).

In summary, these simulations reinforce and illuminate our emerging
understanding of the interactions between smectite clays and relatively nonpolar organic
solutes. New X-ray diffraction evidence (Chappell et al., 2005) suggests that the
strongest sorption of atrazine to smectites occurs for interlayers containing a monolayer
of water. The molecular simulations of atrazine in these monolayer systems indicate that
interactions of atrazine with the clay mineral surface are by full molecular contact with
both basal surfaces, thus removing most of the atrazine molecule from water. It is

reasonable that the entropy increase on removing atrazine from a fully hydrated state in

83

the bulk solution to a lesser hydrated state in the smectite interlayer contributes a
favorable free energy change to the system (Li et al., 2004). This effect is similar to
capillary condensation of hydrophobic solutes (Corley et al., 1996), in which the free
energy of sorption becomes more favorable as pores in the adsorbent become smaller. In
this case, the smectite interlayer is a slit-pore and the optimal thickness for adsorption of
hydrophobic molecules is approximately the molecular thickness, which corresponds (for
aromatic molecules) to d-spacings near 12.5 A. Therefore, the difference in hydration
status of atrazine in the interlayers is believed to account for the dramatic effect of
crystalline swelling on the affinity of smectites for atrazine (Chappell et al., 2005).

Another way of stating this is that the monolayer is an interlayer phase (Laird and
Shang, 1997) that is much more conducive to hydrophobic organic solute partitioning
from water than are other, more hydrated and expanded, interlayer phases. In a
monolayer, water interactions with cations and two clay basal surfaces may reduce the
entropy of water so significantly (Sposito and Prost, 1982) that interlayer water becomes
almost a nonaqueous phase and the penalty for replacing it with a nonpolar organic solute
vanishes. This may explain why our simulations showed that cation-organic complexes
formed so much more readily in a monolayer of water than in the more hydrated
interlayers. Such interlayer cation-organic complexes have been demonstrated
spectroscopically for more polar solutes such as nitrobenzenes (Johnston et al., 2001;
Johnston et al., 2002) and carbamates (manuscript in preparation), but not yet for
atrazine.

As hypothesized (Laird, 1996), the simulated atrazine tended to gravitate to

uncharged and therefore relatively hydrophobic portions of the smectite basal surfaces.

84

This is in accord with observations that, all else being equal, lower-charged smectites will
adsorb more atrazine (Laird, 1996; Laird et al., 1992; Sheng et al., 2001; Chapter 2)
because they have more of the hydrophobic nanosites on their surfaces and also fewer
interlayer cations competing for adsorption domains in the interlayer region (Sheng et al.,
2002). Even in wetter systems swollen with two or three layers of water, our simulations
showed that the alkyl tails of the atrazine molecule tend to interact with the uncharged
nanosites on the smectite basal surfaces, i.e., valence-satisfied basal oxygens located

between charge sites.

85

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Belluck, D.A., S.L. Benjamin, and T. Dawson. 1991. Groundwater contamination by
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90

 

CHAPTER IV

ENHANCED SORPT ION OF TRICHLOROETHENE BY SMECTITE CLAY

EXCHANGED WITH CS“

91

Enhanced sorption of trichloroethene by smectite clay exchanged with Cs+

Abstract

Trichloroethene (TCE) is one of the most common pollutants in groundwater,
interestingly, Cs+ is a co-contaminant at many DOE sites contaminated with TCE.
Smectite clays have large surface areas, are common in soils, have high affinities for
some organic contaminants, hence the potential to influence the transport of organic
pollutants entering soils and sediments. The exchangeable cations present near smectite
clay surfaces can radically influence the sorption of organic pollutants by soil clays. This
research was undertaken to determine the effect of Cs+, and other common interlayer
cations, such as K+ and Ca2+, on the sorption of TCE by a reference smectite clay
saponite. Cs-saturated clay sorbed the most TCE, up to 3500 mg/kg, while Ca—saturated
smectite sorbed the least. We hypothesize that the stronger sorption of TCE by the Cs-
smectite can be attributed to the lower hydration energy and hence smaller hydrated
radius of Cs+, which expands the lateral clay surface domains available for sorption.
Also, Cs-smectite interlayers are only one or two water layers thick, which may drive
capillary condensation of TCE. Our results implicate enhanced retention of TC E in

aquifer materials containing smectites accompanied by Cs+ co-contamination.

92

 

Introduction

Chlorinated solvents such as trichloroethene (TCE, C2HC13) are common
environmental contaminants near thousands of DOE, DOD, NASA and private industry
facilities. TCE has been extensively used for rocket engine flushing, metal degreasing,
and organic extractions due to its effectiveness, noncorrosivity, and nonflammability. In
1994 alone, 42 million pounds of TCE were released into soil and water environments in
the USA, and TCE contamination was identified in at least 852 sites among the 1416 sites
proposed in the US. EPA National Priority list (Doherty, 2000). TCE is also one of the
pollutants detected most frequently and at high concentrations in water of industrial
regions (Patrick et al., 1987), and drinking water sources tested in the US in 1997 showed
a wide range of TCE contamination (Griffin, 2003). The US EPA Drinking Water
Standard for TCE is 0.005 mg L'1 (Spitz and Moreno, 1996), which is just 0.0004% of its
water solubility (water solubility of TCE is 1100 mg L'1 at 25°C, (Verschueren, 2001)),
therefore, even small amounts of TCE leaking into a drinking water supply can pollute a
large water volume. TCE is toxic, resistant to degradation and therefore persistent in
subsurface environments (Li, 2001). There is a need to accurately predict the long-terrn
fate and movement of TCE in soils and subsurface environments.

Sorption by soils and sediments is a major process that controls the transport,
bioavailability, degradation, and ultimately the fate of organic contaminants in the
environment (Koskinen and Harper, 1990). Sorption of organic chemicals diminishes
their bioavailability to target organisms (Alexander, 2000), affects chemical and
microbiological transformations (Gamble and Khan, 1985), and reduces risk of

groundwater contamination (Sonon and Schwab, 1995). Clay minerals and soil organic

93

 

 

matter (SOM) are typically the two most important components in controlling the fate of
organic chemicals that enter soils and sediments. Most research conducted in the last two
decades has indicated that SOM is the primary sorptive phase for the uptake of nonpolar
or slightly polar organic contaminants (Chiou et al., 1979; Miltner et al., 1989; Ong and
Lion, 1991; Wilmanski and Vanbreemen, 1990; Xia and Ball, 2000; Xing et al., 1996;
Zimmer et al., 1989). Transport models often use soil-organic carbon-normalized
sorption coefficients (K...) to predict the distribution of organic chemicals between soil
and water, but this approach ignores the potential contribution of soil minerals to the
retention of organic contaminants in soils and subsoils. Ofien this approach is justified on
the basis that water molecules can effectively compete with organic compounds for the
sorption sites on mineral surfaces, though this has proven to be an oversimplification.
Clay minerals are abundant in soils and aquifer materials, have large surface
areas, so have a great potential for sorption of organic compounds. Of the various clay
minerals commonly found in the soil environment, smectites are most important for
interacting with organics due to their high surface areas, moderate cation exchange
capacities (CECS), and surface reactivities (McBride, 1994). Several previous studies
have attempted to identify the critical ratio of clay minerals/SOM at which sorption by
the mineral phase plays an important role (Farrell and Reinhard, 1994; Grundl and Small,
1993; Karickhoff, 1984). Karickhoff (1984) concluded that at critical ratio of swelling
clays to soil organic C >30, the mineral phase in whole soil becomes important for
sorption of simazine. Grundl and Small (1993) estimated the critical ratio of 62 for
atrazine and 84 for alachlor at which mineral phase sorption accounted for half of the

overall sorption. Farrell and Reinhard (1994) concluded that for high clay to natural

94

organic matter ratios, mineral adsorption dominates organic matter for sorption of
organics.

Adsorption of TCE has been studied by a number of workers, but most such
studies are limited to organo-clays (Sheng et al., 1998; Zhao and Vance, 1998), zeolites
(Alvarez-Cohen et al., 1993; Farrell et al., 2003; Kleineidam et al., 2002; Li, 2001), and
carbon materials (Kleineidam et al., 2002; Langer et al., 1999). Only a few studies on the
interactions of clays with volatile, hydrophobic, low molecular weight halocarbons such
as TCE have been conducted to date (Estes et al., 1988; Farrell and Reinhard, 1994;
Rogers and McFarlane, 1981; Singhal and Singh, 1976; Werth and Reinhard, 1997).
Earlier studies have shown that smectite clays can adsorb organic solvents such as TC E
(1.3 mg g'1 clay, (Farrell and Reinhard, 1994)), and tetrachloroethene (PCE) (5.4 mg g"
clay, (Estes et al., 1988)), to an appreciable extent. Ong and Lion (1991) observed that
under oven-dried conditions the mineral fraction of whole soil plays a significant role in
the sorption of TCE, but competition from water reduces TCE sorption on mineral
surfaces. Based on the higher sorption by a low organic carbon containing silt fraction,
Pignatello (1990) concluded that the mineral fi‘action of whole soil plays an important
role in TCE sorption. He also concluded that mineral fraction of whole soil has a
significant role in the slow release of residual fraction as organic matter in interstitial
pores of particle aggregates are protected from equilibrium with bulk solution by such
mineral fractions (Pignatello, 1990). Results from the few previous studies on smectites
are inconsistent. For example, Farrell and Reinhard (1994) reported that montmorillonite
sorbed 1300 mg kg'1 TCE from the gas phase at 100% relative humidity, while Estes et

a1. (1988) observed little or no TCE sorption by a smectite rich soil. Rogers and

95

 

McFarlane (1981) reported that adsorption of TC E, carbon tetrachloride, and ethylene
dibromide by montmorillonite was variable, ranging from 0 tol 7 % of the chemical
added to the clay, depending on the type of the clay or interlayer cation. They reported
that while A13+ saturated montmorillonite sorbed 17% of the added TCE, Ca2+ saturated
montmorillonite did not sorb any TCE at all.

Thus, there are some indications that the sorption of chlorinated solvents by
smectites is cation-dependent. One effect of changing the interlayer cations in smectites
is to change the layer-spacing of the clay. Farrell et al. (2002) indicated that TCE
adsorption is energetically most favorable in pores that are just large enough to
accommodate a TCE molecule. The theory of capillary phase separation says that
sorption of a hydrophobic molecule into a slit-pore should become more favorable in
materials with small mesopores or micropores (Corley et al., 1996). We therefore
hypothesize that TCE should most effectively compete with water for sorption sites in
smectite clay interlayers when the smectite layer spacing is ~12.5 A, as these clay
interlayers are just large enough to accommodate TCE molecules. Cs+ saturated smectites
generally equilibrate with ~ 12.5 A layer spacings at 100% humidity and even in aqueous
suspensions (Sato et al., 1992), indicating that there is just one layer of water between
smectite layers. Thus, a priori, Cs+- smectites appear to possess optimal properties for
TCE sorption. By the same reasoning, TCE should not be able to compete as effectively
for sorption sites on Ca2+ saturated smectites because the strongly hydrated Ca2+ ions
(AGhyd Ca2+: -1593 kJ/mol while AGhyd Cs+: -284 kJ/mol, (Evangelou, 1998)) cause
greater swelling of the clay interlayers to d-spacings 218 A. Sorption sites in these larger

slit-pores should not be as energetically favorable for TCE condensation from water

96

(Corley et al., 1996), so Ca2+ saturated clays are expected to sorb less TCE than C 3+
saturated clays.

This hypothesis is of environmental interest, since TC E and C 5+ co-contamination
has been reported in groundwater plumes at a number of DOE sites (Brigmon et al.,
1998; McKinley et al., 2001; Riley and Zachara, 1992; Zachara et al., 2002). As one
example, mixed liquid waste containing both Cs+ and TCE was injected through a single
well into the Snake River Plain aquifer at Idaho National Engineering and Environmental
Laboratory from 1953 to 1972 (National Research Council, 2000).

Based on these hypotheses and observations, this study was initiated to study the
adsorption of TCE by saponite, a 2:1 trioctahedral clay, saturated with different
exchangeable cations (K+, Cs+ and Ca2+). K+ and Ca2+ were selected, as they are common
exchangeable cations in the soil environment. Cs+ was selected primarily because TC E
and Cs+ co-contamination has been reported in groundwater plumes at a number of DOE
sites (Brigmon et al., 1998; McKinley et al., 2001; Riley and Zachara, 1992; Zachara et
al., 2002). Secondly, Cs+ has a smaller hydrated radius compared with other cations
studied here, hence is expected to create a more favorable interlayer environment (i.e.
less hydrated) for TCE sorption. The negatively charged sites on saponite clay originate

largely from tetrahedral substitution.

97

 

Materials and Methods

Clays

The reference clay, saponite (SapCa-2, CEC 95 cmol kg") used in the adsorption
study, was purchased from the Clay Minerals Society Source Clay Repository
(Department of Agronomy, Purdue University, West Lafayette, Indiana, USA). The < 2
pm clay-sized particles were separated by wet sedimentation, and then exchanged with
K+, Cs+, and Ca2+ to prepare homoionic clays. To do so, 25 g of clay were dispersed in a
0.1 M chloride solution (500 ml) of the respective cation. The clay suspensions were
shaken for 24 h, and fresh chloride salt solutions were used to replace the original
solutions afier centrifugation. This process was repeated four times to ensure complete
cation saturation. The excess chlorides were removed by repeatedly washing with Milli-Q
water until CT was determined negative by reacting with AgNO3 solution. The clay
suspensions were then quick-frozen, freeze-dried and stored in a closed container for later

USO.

Solute

Analytical grade TCE used in the adsorption experiment was purchased from
Aldrich Chemical Co. (Milwaukee, Wisconsin) with a purity of >99%, and used as
received. The selected physicochemical properties of TCE are as follows: molecular
weight 131.4 g mol", vapor pressure at 30°C is 12.5 kPa, boiling point 87 °C, density
1.45 g cm'3, log K0w at 25°C is 2.53, dipole moment 0.9 Debye, molar volume 89.87 cm3

mol'1 at 25°C, and average molecular diameter is 6.8 A (Farrell and Reinhard, 1994).

98

Determination of Henry ’3 Law Constant

The equilibrium headspace technique was used to determine Henry’s Law
constant of TCE in different electrolytes, which is similar to the method reported by
Garbarini and Lion (1985) and Lincoff and Gossett (1984). Determinations of Henry’s
Law constant as well as sorption isotherms were conducted using 20 mL capacity glass
headspace GC-vials. Bottle volume was determined by weight by filling to capacity with
water at room temperature (23 i 2 °C). Three different TCE concentrations were prepared
in 0.1 M KCl and CsCl, and 0.05 M CaClz. Same volume (25 pL) of stock TCE solution
(prepared immediately prior to their use) (500 mg L'l) was injected into the vials
containing different volumes of each electrolyte solution to create different
concentrations. Vials were capped immediately with Teflon-lined silicone septa and
aluminum crimp caps to minimize volatile loss of TCE. The vials were then shaken
reciprocally overnight at 40 rpm at room temperature (23 i 2 °C); preliminary studies
showed that equilibrium was achieved within this time-period. TCE concentration in the
headspace of each vial was analyzed using a Varian 3700 gas chromatograph with flame
ionization detector (GC-FID) connected with a HP 19395A Headspace Autosampler. The

dimensionless Henry’s Law constant (Hg) was calculated (Lincoff and Gossett, 1984) as:

CgZVIZ —CglVIl (1)

‘=C V —Cg2V

gl :1 82

 

where C g. and C 32 are the TCE concentrations in gas phase of vials 1 and 2, Vgl and V£2

are the gas volumes in vials l and 2, and V,l and V,2 are the solution volumes in vials l

99

and 2. The experiment was performed in triplicate, and at least duplicate results are
presented. We used these values of Henry’s Law constant to calculate the distribution of
TCE in the sorption experiments, as Henry’s Law was dependent on the cation in

solution.

Sorption Isotherm Measurement

TCE sorption isotherms were determined using a batch equilibrium headspace
technique (Garbarini and Lion, 1985). Freeze-dried clay (0.1 to 0.6 g) was suspended in
5.0 mL of electrolyte solution (0.1 M for CsCl and KC], and 0.05 M for CaClz). A series
of initial TCE concentrations (0.5, 1.0, 2.0, 5.0, 10, 20, 40, 70 and 100 mg L") were
prepared by injecting different volumes of stock TCE solution (500 mg L") (prepared
immediately prior to their use in respective electrolyte solution) into vials using a syringe.
Vials were capped immediately with Teflon-lined silicone septa and aluminum crimp
caps to minimize volatile losses of TCE. The vials were then shaken reciprocally
overnight at 40 rpm at room temperature (23 i 2 °C); preliminary studies showed that
equilibrium was achieved within this time. TCE concentration in headspace of the vials
was analyzed using a Varian 3700 GC-FID as above. All isotherm experiments were
performed in triplicate, and at least duplicate results are presented.

After reaching partitioning equilibria, TCE distributions in air, solution and clay

were determined by comparing two systems, viz. systems containing clay and systems

without clay. The same solution volume (V, ), gas volume (Vg ), total mass of TC E ( M ,.)

was applied to both systems. Under such circumstances, sorption equilibria were

evaluated using the following relationships:

100

a. System without clay (at equilibrium):
MT =C,,V,+Cg,Vg (2)
b. System with clay (at equilibrium):

MT = CHI/I +Cg2Vg + C Molar (3)

day

where C ,, and C ,2 are the equilibrium TCE concentrations in solution phase of systems
without and with clay, respectively, Cg1 and C g 2 are equilibrium TCE concentrations in

the gas phase of systems without and with clay, respectively, C is the TCE

clay

concentration in the clay phase, and M is the mass of clay (g) added to the system.

clay
Since the same amounts of TCE were added into both systems, equations (2) and (3) are

equal, thus:

Cll V! + Cgl Vg = C12 V! + CgZVg + CclayMclay (4)
From Henry’s Law, we know that:
C8
C1 = H— (5)

and we determined Hc for each electrolyte using equation (1) as described above.

Substituting C, from equation (5) into equation (4), we get:

 

M clay (6)

day

Cg'V CV—CgZV C V C
F14» glaz-H 1+ ng+

(' t'.

101

From equation (6) the amount of TC E sorbed on the clay (Cclay) can be determined as:

C —C.. * V/H V
C.,.,..=( g‘ ”MK’ ‘)+ g] (7)

 

(lay

C 3, and C g 2 are determined using a Varian 3700 GC-FID. The TCE concentrations in

aqueous solution and clay phases were calculated using equation 5 and 7, respectively.

X -ray Diffraction

Clays were mixed with water to form suspensions, which were used to prepare
clay films for X-ray diffraction analysis. The suspensions were dropped onto glass slides
and air-dried overnight at ambient conditions to obtain oriented clay films. The clay films
were then allowed to equilibrate with water vapor for 2 d in a closed container containing
distilled water to expose the clay films at 100% relative humidity. X-ray diffraction
patterns were recorded using Cu-Ka radiation and a Philips APD 3720 automated X-ray
diffractionmeter, fitted with an APD 3521 goniometer and a 0-compensating slit, a 0.2
mm receiving slit and a diffracted beam graphite monochromator. Diffraction patterns

were recorded from 5.5 to 16.0 °20 at a step of 0.005 °20 with 4 s/step.

102

 

Results and Discussion

Henry ’5 Law Constant

Henry’s Law constant of TCE was found to be different in the three different
electrolyte solutions used for the sorption study. The values of Henry’s Law constant
were determined to be 0.61 for 0.1 M KCl, 0.74 for 0.05 M CaClz, and 0.92 for 0.1 M
CsCl, which are all different from the value (0.40) in the pure water-air system
(Verschueren, 2001). The presence of salt in the water phase makes the aqueous solution
more unfavorable for the dissolution of TCE; thereby enhancing the Henry’s Law
constant. Since these differences are large and since the Henry’s Law constant plays a
significant role in the calculation of sorption (eqs 5 and 7), we measured these values of
Henry’s Law constant for each salt solution and used these values of Henry’s Law
constant to calculate the TCE aqueous concentrations in the relevant electrolyte solutions

in our sorption experiments.

Sorption Isotherm

TCE sorption isotherms are shown in Figure 1. The data points are the average
values of equilibrium concentration (X-axis) and amount of TCE sorbed (Y-axis), and the
error bars show the standard deviations. Ca-saturated saponite sorbed the least TC E
(Figure 1a) and Cs-saturated saponite sorbed the most TCE (Figure 1b). K-saturated
saponite sorbed roughly twice that amount of the Ca-saponite (Figure 1a). The sorption of
0.34% (wt/wt.) TCE by the Cs-smectite at an equilibrium concentration of 90 mg L'I is
the largest TCE sorption reported for any clay mineral that we know of, though Alvarez-

Cohen et a1. (1993) observed 20% TCE sorption for a low-charge zeolite, and Farrell et

103

 

al. (2003) observed up to 50% TCE sorption for another low-charge zeolite under both
dry air and 100% relative humidity, while Farrell and Reinhard (1994) reported 0.13%
(wt/wt.) TCE sorption by montmorillonite and Rogers and McFarlane (1981) did not
observe any TCE sorption by Ca2+ saturated montmorillonite. While solute transport
models generally assume no TC E sorption to clays in aquifer materials, Figure 1 shows
that there might be significant sorption of TCE to aquifer smectites especially if Cs+ is a

co-contaminant in the plume.

104

 

800

 

 

 

 

 

 

 

 

 

Amount of TCE sorbed, mg/kg

 

 

 

 

100
4000
3000
2000
1000
0 , I 1 r
O 20 40 60 80 100

Equilibrium Concentration of TCE, mg/L

Figure 1. Sorption isotherms representing trichloroethene uptake from water by
reference saponite (SapCa-Z) clay saturated with (a) calcium, and potassium, and
(b) cesium.

105

 

The lower effectiveness of Ca-SapCa-Z may be due in part to the larger enthalpy
of hydration of Ca2+. The larger hydration sphere around Ca2+ causes Ca—saponite to swell
in water, thus expanding the slit-pore and decreasing the energy gained by removing TC E
from aqueous solution (Corley et al., 1996). Also, the larger hydration sphere of C a2+
obscures a larger fraction of the lateral siloxane surface between exchangeable cations
compared to those associated with either Cs+ or K”, and thereby effectively shrinks the
size of the adsorptive domains for Ca2+ saturated clays. Conversely, the higher
effectiveness of Cs-SapCa-2 in sorption may be due in part to the lower hydration energy
of Cs+ and hence smaller hydrated radius of Cs+ compared to K+ or Ca2+. The relatively
weak hydration of Cs)r also causes Cs-saponite to swell less in water, which potentially
provides a more favorable sorption domain (Corley et al., 1996), and allows the organic
to compete better with water for interlayer sites. The fewer water molecules associated
with exchangeable Cs’, for example, obscure a smaller fraction of the clay surface than
those associated with Ca2+, resulting in the increase in the adsorptive domains for Cs+
saturated clays. Zhu et al. (2004) also reported high sorption of polycyclic aromatic
hydrocarbons (PAHs) by Cs-saturated clays. They hypothesized that relatively weak
hydration of Cs+ resulted in stronger cation-1r interaction ensuring higher sorption of
PAHs, while bivalent cations like Ca2+ having high hydration energies induced weak
cation-1t interactions, resulting in low sorption. While ‘hydrated radii’ are poorly defined,
typical estimates of hydrated radii are 9.6 A for Ca”, 5.3 A for K+, and 3.6 A for C 5+
(Evangelou, 1998). Thus the cross-sectional area of hydrated Ca2+ is about 290 A2, more
than seven times that of Cs+ (40 A2); therefore one hydrated Ca2+ takes up 3.5 times as

much room in the clay interlayer as do the two Cs+ ions that Ca2+ replaces. Using these

106

 

hydrated radii along with an idealized smectite basal surface area of 750 m2 g"l and the 95
cmol kg'I CEC of saponite SapCa-2, we compute that the hydrated radii of Ca2+ or K+
must overlap in the interlayer regions of homoionic SapCa-2, meaning that TCE would
have to compete with strongly bound water for interlayer sorption sites, which it
apparently does not do very well. On the other hand, Cs" and its hydration shells are
projected to occupy only about 67% of the interlayer space, implying that TCE has to
compete only with less strongly bound water in the Cs-smectite interlayers. This may
explain why the Cs-smectite is a much more effective sorbent for TCE. Sheng et al.
(2001; 2002) also observed similar effects of cation exchange on domain size for the
sorption of atrazine and DNOC by SWy-2 clay.

Of the cations studied here, Cs-smectite is most likely to have a 12.5 A layer
spacing, even in aqueous suspension. If so, and assuming TCE sorption is in the
interlayer, then at most a monolayer of TCE can sorb and the system should exhibit
Langmuir-type behavior. The best-fitting Langmuir parameters are a monolayer coverage
of about 4500 mg kg'l, and Langmuir K of about 0.032. This Langmuir function does not
fit the data particularly well at low concentrations because of the upward curvature of the
isotherm data, but the Freundlich curve is even worse for the same reason. Similarly, a
Langmuir isotherm does not fit well for K-smectite at low concentrations, but can be fit at
high concentrations (Figure 1a). The Ca-system on the other hand will never have a 12.5
A layer spacing in suspension, so sorption of a simple monolayer of TCE is unlikely in
the Ca-smectite interlayer. For the Ca-system there was no evidence of an adsorption
plateau, so there is no way to fit the Langmuir equation. Therefore, we report F reundlich
l-N LN

parameters for our Ca-system: Kr (mg kg'l), the Freundlich sorption coefficient is

107

3.11, and N (unitless) is 1.16. The N value > 1.0 for Ca-smectite is consistent with
earlier results (Li et al., 2003; Sheng et al., 2001) for hydrophobic compound sorption to

Ca-smectites.

X -ray Diffraction

Differences in the swelling behavior between Cs- and Ca—SapCa-2 may also
contribute to their differential adsorption of organic compounds. Many Cs-smectites
equilibrate with ~12.5 A layer spacings (hydrate monolayer structures) at 100% humidity
and even in aqueous suspension (Sato et al., 1992). All Ca-smectites, on the other hand,
swell to more than 15 A in suspension. Our results show that Cs-saturated saponite swells
to 12.1 A at 100% humidity (K-saturated saponite swells to 12.2 A), while Ca-saturated
saponite swells to 14.9 A at 100% humidity (Figure 2). Of course, the d-spacings of these
clay films may be smaller than those of the smectites in aqueous suspension when the
adsorption experiments were carried out, although the clay films were exposed to 100%
RH. Teppen et al. (1998) concluded that, for idealized siloxane surfaces of layer charge
zero, the most stable mode of TCE sorption occurs by full molecular contact, coplanar
with the clay surface. They further concluded that this kind of interaction is suppressed
by increasing water loads. The average molecular diameter of TCE is 6.8 A (Farrell and
Reinhard, 1994) and the thickness perpendicular to the molecular plane is only about half
that. We propose that in a system with a monolayer of water (K+, and Cs+-saturated clays
near 12-A d-spacings as in Figure 2), TCE intercalates into clay interlayers and, coplanar
with the clay basal surfaces, interacts with two opposing clay layers by fiill molecular

contact with both via van der Waals and electrostatic forces. That is, TCE like many

108

 

 

planar aromatic molecules is only about 3 to 3.5 A thick so fits very well between two
smectite layers at d-spacings of 12-12.5 A (Sheng et al., 2002). Thus Cs-smectite, which
can reasonably be expected to have a 12.5-A d-spacing even in aqueous suspension
(Berend et al., 1995; Kawano and Tomita, 1991; MacEwan and Wilson, 1980; Suquet
and Pezerat, 1987; Suquet et al., 1975; Suquet et al., 1977), may have the optimal slit-
pore to maximize the energy of TCE condensation from solution (Corley et al., 1996),
because if the interlayer region was any more narrow than 12 A, then TCE could no

longer intercalate at all because it simply could not fit.

 

Intensity

 

 

 

 

 

 

.-
d
.1

2 Theta

Figure 2. X-ray diffraction pattern of reference saponite (SapCa-Z) clay in oriented
clay films as a function of different cation saturation at 100% humidity.

 

109

Given a 3-A thick interlayer region, 33% of the basal surface not covered by Cs+
hydration shells as described above, and the TCE molar volume cited above, we predict
that the Cs-saponite might be able to sorb up to 10% TCE by weight if all interlayer
volume outside primary Cs+ hydration shells were filled with TCE. Since we observed
only 0.34% sorption (Figure 1), our assumptions may be simplistic in that a) The kinetics
of tortuous TCE diffusion to many interlayer sites may be slow, or b) TCE is not able to
displace other interlayer water outside primary Cs+ hydration shells. In this regard, a
tetrahedrally substituted clay may have been an inopportune choice in that tetrahedral
charge is much more localized on fewer surface oxygens (McBride, 1989) and
exchangeable cation ordering is only found in tetrahedrally substituted smectites (Odom,
1984). The exchangeable cations (at a given hydration state) can always approach a
tetrahedral site more closely than an octahedral site (Bleam, 1990), and thereby
strengthen the hydrogen bonding of interlayer water to the siloxane surface itself (Yariv,
1992)

Sorption isotherms for Cs-, K-, and Ca-saturated saponite are nonlinear (Figure l
a and b). This nonlinearity could be due to several possible causes: 1) The heterogeneous
distribution of sorption sites on the clays, 2) Equilibrium was not achieved during the
course of the study, or 3) At least some of the postulates of a Langmuir-type sorption are
satisfied, particularly the presence of a finite number of sorptive sites that, when fully
occupied, constitute a ‘monolayer’ of adsorbate coverage. Li and Werth (2001)
concluded that since sorbate molecules first occupy those sorption sites that have the
highest adsorption potential, therefore sorption isotherms of TCE on two different natural

solids were nonlinear. Similarly, Estes et al. (1988) and Hamaker and Thompson (1972)

110

did not observe equilibrium for sorption of organic on montmorillonite clay over a period
of 30 days and concluded that slow changes in clay surface that resulted in the continuous
exposure of new adsorption sites were the primary reason for the failure to achieve
equilibrium. If our hypotheses about TCE sorption to the Cs-saponite are correct, then the
smectite interlayer should behave very nearly like an ideal Langmuir system because the
12-A d-spacing allows only monolayer coverage in the interlayer and the lateral
adsorption domains between hydrated cations are only a fraction of a finite smectite basal
surface area.

The type of smectite clay also affects TCE sorption. For example, our results
show that Ca-saturated saponite sorbed ~ 460 mg TCE/kg at the highest equilibrium
concentration, but Rogers and McFarlane (1981) in a study with Ca-saturated
montmorillonite (dioctahedral, CEC 80 cmol/kg) did not observe any apparent sorption of
TCE. Two possible reasons for these differences are: (i) differences in the CEC of the
clays, and (ii) type of the clay. However, we can ignore the differences in CEC to be
responsible for the differences in sorption of TCE by these two clays. Generally, the
lower the CEC of clay, the smaller is the clay surface occupied by the exchangeable
cations, which makes more clay siloxane surface area available for sorption. Several
workers have reported an inverse relationship between the CEC of the clay and the
amount of the organic adsorbed (Laird et al., 1992; Lee et al., 1990; Sheng et al., 2001;
Sheng et al., 2002). Based on these results, Ca-saturated montmorillonite of Rogers and
McFarlane (1981) should have sorbed more TCE than that observed in the present
investigation. Therefore, it seems that differences in TCE sorption may be related to the

source of negative charge on the clays and not to differences in the CEC of the clays. The

111

 

negative charge on the saponite clay arises from isomorphous substitution in tetrahedral
structures. Generally, for a given layer charge, clays with negative charge from
tetrahedral substitution sorb more organics than do clays where negative charge develops
from octahedral substitution. Compared to smectites with negative charges from
octahedral isomorphous substitution, the 2:1 smectites with negative charges from
tetrahedral isomorphous substitution exhibit less swelling of the interlayer space in water.
The theory of capillary phase separation (Corley et al., 1996) thus implies that sorption of
hydrophobic molecules from water should be more favorable in tetrahedrally versus
octahedrally substituted smectites (for a given layer charge and interlayer cation), and
such behavior has been observed (Laird et al., 1992; Sheng et al., 2001) though not

explained.

112

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117

CHAPTER V

MOLECULAR SIMULATIONS TO ESTIMATE THERMODYNAMICS OF

POLAR ORGANIC SORPTION TO MONTMORILLONITE

118

 

Molecular simulations to estimate thermodynamics of polar organic sorption to

montmorillonite

Introduction

The list of anthropogenic organic chemicals that enter the environment expands
continually. Soils and sediments are important environmental sinks for these chemicals
and often contain the bulk of such chemicals after release. Hence, the potential for
sorption to solid phases is a key parameter in estimating the environmental risk and fate
of these chemicals, and has therefore been the subject of intensive investigation (Bailey
and White, 1970; Boyd et al., 2001; Chiou, 1998; Chiou et al., 1979; Haderlein and
Schwarzenbach, 1993; Karickhoff and Morris, 1985; Karickhoff et al., 1979; Sheng et al.,
2002; Weissmahr et al., 1999).

Soil adsorption coefficients can be determined by batch equilibrium methods, but
these can be difficult and expensive procedures. Despite 40 years of experimental work,
soil sorption coefficients are in many cases unavailable even for high-production-volume
chemicals (Sabljic et al., 1995). Even when sorption data are available for a given
compound, they pertain to only a very limited number of sorbents. It would require
unfeasible financial and scientific resources to measure adsorption coefficients for all
pertinent chemicals interacting with all relevant natural phases. Thus there is interest in
predictive models that can estimate the potential for organic compound sorption to a
given solid phase (Andersson et al., 2002; Goss and Schwarzenbach, 2001). Such models
could aid our prediction of contaminant transport and fate, while potentially even pre-

screening the probable fates of new compounds in order to minimize environmental risk.

119

For neutral organic compounds (NOCs), a breakthrough in predictive sorption
modeling came with the realization that their sorption from water to soil organic matter
(SOM) is analogous to partitioning into a bulk organic phase (Chiou et al., 1979;
Karickhoff et al., 1979). Numerous approaches to estimate K0c (the soil organic C-
normalized linear sorption coefficient) and therefore sorption to SOM have been
published during the last 25 years, such as estimation of K0C based on water solubility
(Briggs, 1981; Chiou et al., 1983; Gerstl, 1990; Karickhoff, 1981), n-octanol/water
partition coefficient (Briggs, 1981; Chiou et al., 1983; Gerstl, 1990; Gerstl and
Mingelgrin, 1984; Karickhoff, 1981), reversed-phase high-performance liquid
chromatographic (RP-HPLC) capacity factor (Guo et al., 2002; Kanazawa, 1989; Muller
and Kordel, 1996), topological indices (Gerstl, 1990; Sabljic et al., 1995), molecular
parameters and linear solvation energy relationships (Gramatica et al., 2000; Huuskonen,
2003; Kanazawa, 1989; Lohninger, 1994), fragment constants and structural correction
factors (Tao et al., 1999), and one-parameter linear free energy relationships (LFER)
(Goss and Schwarzenbach, 2001). However, there are drawbacks with most of these
approaches. For example, estimation of K0c based on water solubility and octanol/water
partition coefficients can be very difficult for very hydrophobic substances (Gawlik et al.,
1997). Furthermore, Goss and Schwarzenbach (2001) argued that the predictive power of
one-parameter LFERs (i.e., Kow) is limited, because no single parameter can completely
describe all the molecular interactions that determine environmental phase partitioning
processes.

For NOCs with polar functional groups, specific interactions between adsorbate

and adsorbent seem most likely to impact sorption. For example, the strong sorption of

120

nitroaromatic compounds to certain smectites (swelling clay minerals) seems to be driven
by complexation of interlayer cations by solute -NOz groups (Boyd et al., 2001; Johnston
et al., 2001; Johnston et al., 2002). In addition to such complexation, hydrogen bonding
(Laird, 1997) and hydrophobic interactions (Boyd et al., 2001; J aynes and Boyd, 1991)
appear to play a role in NOC sorption to clay minerals, and evidence has been marshaled
for other forces such as cation-pi bonding (Zhu et al., 2004) and electron donor-acceptor
complexes (Haderlein and Schwarzenbach, 1993; Weissmahr etal., 1997). Since
smectites have large surface areas and can contribute to and even dominate (Sheng et al.,
2001) the sorption of many NOCs, it would be useful to be able to predict the strength of
NOC sorption to smectites. To do so, however, it is apparent that many different types of
molecular interactions need to be considered.

Molecular modeling techniques, therefore, may be appropriate tools to estimate
smectite-NOC sorption affinities. Molecular simulations can help interpret experimental
observation, discriminate among competing models to explain macroscopic or
spectroscopic data, integrate bulk (P,V,T) and molecular properties, and test or generate
scientific hypotheses (Bleam, 1993; Cygan, 2001; Lasaga and Gibbs, 1990; Sauer and
Ahlrichs, 1990; Teppen et al., 1998). In particular, classical molecular simulations (often
termed molecular mechanics) that parameterize all electronic effects have been used
successfully to study heterogeneous clay mineral systems (Chang et al., 1997; Cygan,
1998; Delville, 1991; Greathouse et al., 2000; Kalinichev et al., 2000; Skipper, 1998;
Skipper et al., 1991; Smith, 1998; Sutton and Sposito, 2001; Teppen etal., 1997). Such
simulation methods are most commonly used to probe structural properties, because it is

much more difficult to accurately compute energy relations (Schafer et al., 1990).

121

Nonetheless, the methods are reaching a reasonable level of maturity where it may be
fruitful to test the ability of molecular mechanics methods to compute adsorption

energies.

Choice of compounds

Several experimental studies (Haderlein and Schwarzenbach, 1993; Haderlein et
al., 1996; Weissmahr et al., 1999; Weissmahr et al., 1997) indicated that nitroaromatic
compounds adsorb strongly to smectite clay minerals, especially when the smectite is
saturated with K+ or Cs+ counterions. In comparing molecules of very different organic
functionality, there were indications that planar compounds with polar functional groups
that might complex hydrated interlayer cations were most favorably adsorbed (Sheng et
al., 2001). To test and refine this concept, Boyd et a1. (2001) measured the adsorption
from water of a planar series of para- and meta-substituted nitrobenzenes (SNBs) by the
K-saturated Arizona smectite clay mineral SAz-l, a high-charge montmorillonite. Fourier
transform infrared (FTIR) spectroscopy revealed shifts in the —NOz vibrational modes of
1,3,5-trinitrobenzene (TNB) (Boyd et al., 2001) and other SNBs (Johnston et al., 2001;
Sheng et al., 2002) upon adsorption to K+-smectite, results that are consistent with the
complexation of K+ by —NOz groups. Such vibrational shifts were absent or less
pronounced for smectites saturated with more strongly hydrated cations (i.e., Na+, M g”,
Ca2+, Bar"). Molecular dynamic simulations of dinitrobenzene systems (Boyd et al.,
2001; Sheng et al., 2002) predicted that each -NOz group interacted with one or more

different interlayer K+ counterions. The experimental adsorption data (Boyd et al., 2001)

122

indicated that, among the nitrobenzenes tested, F reundlich complexation constants were
strongest for the carboxylic methyl ester functional group, and decreased in the order:
R-COOCH3 > R-NOz > R-COCH; > R-CN > R-CHO > R-OCOCH3 > R-OCH3 > R-NHg,
where R = nitrobenzene. Such an ordering is plausible from a complexation standpoint,
considering both the presence and steric availability of atoms that can interact favorably
with KL Two hypotheses that have been presented (Boyd et al., 2001) to explain this
selectivity sequence for nitroaromatic molecules are 1) Additive complexation
interactions between organic molecule functional groups and multiple interlayer cations,
and 2) Hydration energy differences between the various organic molecules. Recently,
enthalpies were measured (Li et al., 2004) for the adsorption of three nitroaromatic
solutes to smectites. Thus, the nitroaromatic compounds are a promising system for
developing molecular simulation methods to study adsorption energies, because
systematic bulk adsorption data (Boyd et al., 2001), spectroscopic (Johnston et al., 2001)
and X-ray diffraction (Li et al., 2004; Sheng et al., 2002) investigations of interlayer
environments, and adsorption enthalpy data (Li et al., 2004) are all available. A
simulation model is of potential importance since it might enable us to estimate the
adsorption energies of new chemicals even before they are synthesized. If we are
successful, it might be possible in the future to screen interesting chemicals (i.e.,
pesticides) for their soil sorption and leaching potentials even before chemical synthesis
and biological efficacy testing are done.

Our goal in the present study is to develop and test two methods that use
molecular dynamic simulations to estimate the enthalpies of adsorption for several NOCs

interacting with smectites, and then correlate these enthalpy predictions with

123

experimental enthalpies and adsorption coefficients. The objectives of this study are to
simulate a series of these nitroaromatic molecules in the interlayer region of
montmorillonite clay, to use the molecular dynamics results to compute the relative
adsorption energies for these nitroaromatics, and to determine the dominant components

of the adsorption energies.

Materials and Methods

A model for K-SAz-l was created and molecular dynamics simulations of para-
substituted nitrobenzene-clay complexes were performed in order to estimate the
adsorption energies for a series of nitroaromatic compounds. In this approach, electrons
are not treated explicitly, but their effects are parameterized into classical (that is, non-
quantum) mechanical energy functions comprising a force field. The clay model had the
composition Kl2(Al36Mg12)Si96024o(OH)4g with a CEC of 132 cmol/kg. The smectite
surfaces were constructed using methods previously detailed (Teppen et al., 1997), in
which a muscovite mica layer was the starting point and octahedral Mg-for-Al
substitutions were random except that no two Mg octahedra were allowed to be
immediate neighbors. In order to prepare for molecular dynamics simulations of
nitrobenzene-clay complexes, the interlayer region of the clay was expanded to 20 A and
one molecule from the series of substituted nitrobenzene was added to the unit cell, with
its molecular plane perpendicular to the basal planes of the clay in an attempt to avoid
biasing the final structure. The simulated loading rate was equivalent to 18 mg
m'troaromatic per g clay. Water molecules were also added (in random interlayer

positions) so that constant-pressure simulations resulted in 12.3 A clay layer spacings. If

124

the simulation did not result in a clay layer spacing of 12.3 A, then water molecules were
either removed or added (after expanding the clay) so that the equilibrium clay layer
spacing was 12.3 :1: 0.1 A, in accord with the experimental data for substituted
nitrobenzenes in air-dried films of K-SAz-l.

The energies were computed using a force field developed for clays (Teppen et
al., 1997) combined with the polymer consistent force field (PCFF) (Maple et al., 1994;
MSI, 2000) for the organics and water. As a further modification, charges on the organic
solute were assigned using the COMPASS force field (Sun, 1998), which optimized
organic atomic charges for condensed phase systems. In previous work (Boyd et al.,
2001), we computed quantum mechanical charges for a series of nitroaromatic solutes
and found COMPASS charges to more faithfully reproduce the quantum charge estimates
than did PCF F charges. The charges on organic solute atoms are the most important
force field parameters in determining interactions with the clay surface, interlayer water,
and aqueous cations. In a typical force field, the nuclear charge and the local electron
density around a given atom are subsumed into one atom-centered partial atomic charge,
and the COMPASS charges for nitroaromatic solutes are illustrated in Figure 1.
Commercial software (M81, 2000) was used to compute the classical molecular
dynamics. No constraints were imposed on our systems during their molecular dynamics
except for the contents of the simulation cell: All atoms were completely free to move,

and unit cell shape and volume were unconstrained.

125

 

 

 

 

 

 

 

 

 

 

 

01\ /.-01 Charge
N]
I C1 +0.2390
H,\ /C1\ H] C2 -0.1268
C2 \ Cz/
H Nl +0.6170
/C2 / C2 H1 +0.1268
H, \C/ \H1
1
| 0. -0.4280
/N1‘
0] '0l

 

Figure 1. Charges on para-dinitrobenzene molecule used in simulation study.

The molecular dynamics simulations were run in the NPT ensemble for 50 ps,
much longer than the time required for the system volume and energy to equilibrate.

Trajectories were saved from each molecular dynamics simulation, and snapshots from

 

the post-equilibrium portion of each dynamic trajectory were used to estimate clay-
organic interaction energies by computing the difference between the total energy of the
system and the energies of the hydrated clay and the organic separately. A scripting
program was written to automate the energy calculation for many snapshots of a given
system. We used two different algorithms for converting simulation energies to

predictions for experimental adsorption enthalpies.

126

Computational algorithm 1

The typical experimental scheme for measuring adsorption of aqueous organics to
clays is

Clay(aq) + 0rg(aq) 1::> clay-org(aq) + H20(aq) [1]

Achpt

where chp. is the experimentally determined adsorption energy (for example, the overall
free energy, enthalpy, or entropy of the transfer process). The experimental reaction is
too difficult to simulate directly, since diffusion kinetics are slower than the ns timescales
that are accessible to molecular simulations. Instead, we propose an alternative path that
is more computationally feasible and follows the spirit of many schemes that have been
used to simulate the binding of drugs to polypeptides (Gohlke and Klebe, 2002; Halperin
et al., 2002). We employ the following thermodynamic cycle for computing the energy

of adsorption of an aqueous organic molecule to a hydrated clay mineral:

 

AHexpt

AH 3 AH 5 [2]

Clay-cavity..." + orgmq) + H20 E:> clay-org(aq) + cavity(aq) + H20
AH4

Since the free energy, enthalpy, and entropy are state functions (Atkins, 1998), their
magnitudes should be independent of path, so that

AHCXPI = AH3 + AH4 + AH5 [3]

127

This thermodynamic cycle is very similar in spirit to a recent study by Goss and
Schwarzenbach (2001) where the free energy change AGi for partitioning of a compound
between two phases is decomposed to enable a more mechanistic understanding of the

partition process (Figure 2).

 

Partitimingofsnnllorganicsbetween aclayplnseandanaqueomphase

 

AQ=AH3+AILclay+AHS'AHi-aqmis
Figure 2. Schematic for deconposing the energy changes for tramfer
ofacorrpourxifiomaqueous solutiontotheclaylayer
(after Goss and Schwarzenbach, 2001).

 

 

 

128

Similarly, Scmidt (Schmidt et al., 2002) proposed that the free energy of transfer
of a molecule between two bulk phases 1 and 2, ALZG, depends on the cohesive energy of
both phases and the interactions of the molecule with both surrounding phases (van der
Waals and H-bond interactions). All these interactions need to be taken into consideration
when comparing the partitioning of a set of molecules between various organic phases
and water. The transfer of a molecule into a bulk phase requires a cavity to be formed to
accommodate the molecule (and therefore depends on the size of the molecule). For the
formation of the cavity, the intermolecular interactions of the bulk phase molecules are
interrupted. Thus, the stronger the cohesive interactions in the bulk phase, the more

energy is required to form a cavity in the phase.

To relate AH 4 of our thermodynamic cycle to Figure 2, we recognize that

AH 4 = AH i-clay _ AH i-aqueous = AH (clay-cavity —> clay-org) + AH (org(aq) -—) water cavitytaql) [4]

That is, AH4 is the sum of the energy changes for transferring the organic compound out
of a cavity in the aqueous solution and into a cavity in the clay interlayer region. We
estimated each energy term in equation 4 by equilibrating the entire organic-water (clay)
system and then calculating interaction energies between subsets of the system. For
example, we used the following algorithm to calculate AHimy (the interaction energy
between organic compound 1' and its environment in the hydrated smectite interlayer) of
Figure 2:

AH...21y = AH — AH 1 — AH 2; where [5]

to!

129

0 AH to, = Total clay—organic system energy

0 AHl

Energy of system after organic has been deleted

0 AH2 = Energy of the system after clay layers, interlayer cations, and water
have been deleted

0 Compute AHMW for many (2 150) snapshots to compile an estimate for the

ensemble average value

Using this same general algorithm, we also computed AHi.aqueous (the interaction
energy between the organic compound and its environment in aqueous solution), AH 3
(the interaction energy between the clay phase and the water molecules taken from the
clay interlayer to create a cavity the size of the organic compound: AH3 was calculated by
removing water molecules from the smectite with 12.3 A layer spacing but no interlayer
organic compound — see Figure 2), and AH; (the energy required to fill the cavity in
water that previously held the organic compound — see Figure 2). Once the four energy
terms AH..clay, AH.-aqueous, AH3, and AH5 were computed, we were able to estimate AH...pl
as illustrated in cycle [2].

Note that we have neglected two necessary terms in the total energy. Both AH 3
and AH5 can be though of as containing two separate components (Gallicchio et al.,
2000), one for creating a cavity in water and a second component for the interaction of
the cavity contents with its environment. Calculation of the first component, cavitation,
for AH3 presents some special problems: Most studies (Gallicchio et al., 2000; Hdfinger
and Zerbetto, 2003; Sandberg et al., 2002) of cavitation in water work with bulk aqueous

phases that can expand to accommodate the cavity without any fundamental change to

130

the phase at distance from the cavity. In contrast, smectite clays might not be able to add
a cavity to their interlayer volume without swelling and thus becoming a new phase. In
our molecular dynamics methods, we use periodic boundary conditions to allow a
relatively small unit cell (here, just 700 atoms) represent a condensed phase by
replicating the unit cell to infinity. In such a case, the composition of the unit cell is
typically fixed and water cannot leave the interlayer region (for example, in response to a
growing cavity). In principle, we could do a realistic simulation by constructing a
reasonable clay fragment consisting of multiple smectite layers embedded in a much
larger unit cell containing an aqueous solution. Then, as a cavity grew, water could
actually leave the interlayer region and enter the bulk aqueous phase. We leave such a
simulation for the future, as we anticipate very slow simulations due to a huge number of
atoms and also slow kinetics of water diffusion out of the clay. Indeed, with such a
system one might have better luck actually trying to get an organic solute to adsorb into
or desorb from the interlayer region and track the enthalpy changes directly through time.
For the present study, we will use literature estimates for the cavitation components of
both AH3, and AH5, and simulate just the water addition or removal from the cavity.

The molar volume of solid 1,4-dinitrobenzene is 103.5 cm3/mol (Lide, 1995),
implying a molecular volume of 172 A3/molecule, which should be a lower limit for the
molecular volume in any other phase since crystalline 1,4-dinitrobenzene packing should
be maximally efficient. This is about six times the molecular volume of liquid water (30
A3/molecule), so any cavity for insertion of 1,4-dinitrobenzene into another phase should
require removal of more than six water molecules. Our clay-water system (containing no

1,4-dinitrobenzene) that gave clay layer spacing of 12.3 A contained 55 water molecules

131

per unit cell while our clay-1,4-dinitrobenzene-water system of the same dam-spacing
required 45. This difference of 10 water molecules per 1,4-dinitrobenzene is reasonable
given packing inefficiency, and was adopted as an empirical estimate for the number of
water molecules that must be removed in order to make a cavity for the nitroaromatic
solutes here. Values of H3 (equation 2) were computed by removing 10 continuous water
molecules out of the 55 originally present in the clay-water system (containing no
organic) with deal-spacing of 12.3 A (same clay layer spacing as our clay-organic-water
system), while the values of H5 (equation 2) were computed by removing 10 waters out
of the 252 originally present in a periodic box of water (containing no clay or organic)

(see Figure 2). The same estimates for H3 and H5 were used for each compound.

Computational algorithm 2
The overall enthalpy change for adsorption of NACs to mineral surfaces from an

aqueous phase can be expressed by combining equation 3 and 4 to give:

AI‘iexpt : A1"I3cav + AH3int + AI'1i-clay _ AHi-aqueous + A1"15cav + AHSint [6]

Summing just the fourth and fifth terms of equation 6 yields approximately the
dehydration enthalpy (that is, the negative of the hydration enthalpy Athd) of the NAC,
since hydration involves creating a cavity in water and placing the organic solute in the

cavity (Goss and Schwarzenbach, 2001). Thus,

AInlexpt : AH3cav + AI"13int + AHi-clay _ AI'lhyd + AHSint [7]

132

The purest measure for hydration of a compound is normally defined (Plyasunov

and Shock, 2000) as transfer of the compound from its ideal gas (where it has no

energetic interactions with its environment) to its aqueous solution. For example:

1,4-dinitrobenzene (g) <——> 1,4-dinitrobenzene (aq) [8]

Note that this reaction is the sum of the dissolution reaction

1,4-dinitrobenzene (s) e 1,4-dinitrobenzene (aq) [9]

and the reaction

1,4-dinitrobenzene (g) <—) 1,4-dinitrobenzene(s) [10]

Reaction 10 is the reverse of the sublimation reaction, so the hydration enthalpy can be

expressed as (Plyasunov and Shock, 2000)

Athd = AHdis — AHsubr [11]

Fortunately, estimates for sublimation enthalpies AH...“ have been published (Chickos
and Acree, 2002) for several of the NACs studied here, and are listed in Table 3. This

allows equation 7 to be rewritten

AI'Iertpt = AH3cav + AH31nt + AHi-clay + AHsubl _ AHdis + AHSint [12]

133

Now, when Athd (equation 7) or AH...“ and AHd,-, (equation 12) can be looked up or
estimated for any compound, it becomes easier to estimate an experimental adsorption

enthalpy for that compound.

Results and Discussion

Algorithm 1 validation

As a first test of our overall method and specific algorithms, we have computed
adsorption enthalpies for the two nitrobenzene derivatives (R-NOz and R-COOCH3,
where R=nitrobenzene) that adsorbed most strongly to the smectite in the experimental
study (Boyd et al., 2001) and compared them with that of p-xylene, which is of similar
size but lacks the polar functional groups that might enable complexation with interlayer

cations so is expected to sorb with a much less favorable enthalpy (Table 1).

134

 

 

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135

Note that the clay-organic interaction enthalpies are quite favorable for all nitrobenzene
solutes but not for p-xylene. The value of AH-wy for 1,4-dinitrobenzene adsorption to K-
SWy-Z smectite was recently estimated to be -93 kJ/mol (Li et al., 2004), compared with
our estimate (Table 1) of -1 54 kJ/mol. The same paper estimated AH-{iay for 1,3,5-
trinitrobenzene adsorption to K-SWy-Z smectite at -124 kJ/mol (Li et al., 2004), again
much smaller than our estimate (Table 1) of -235 kJ/mol. The only other estimates we
know of for related systems are quantum mechanical: -38 kJ/mol for 1,3,5-
trinitrobenzene interacting with a single, neutral 2:1 clay surface (Pelmenschikov and
Leszczynski, 1999), and -49 kJ/mol for nitrobenzene interacting with two waters of
hydration at the surface of a Na-smectite fragment (Gorb et al., 2000). The values of
[AH-{lay for 1,4~dinitrobenzene and 1,3,5-trinitrobenzene computed here by molecular
mechanics are about 65 and 90% larger, respectively, than the values estimated from
experimental data (Li et al., 2004), and thus may indicate some deficiency in either the
force field or the structural model used in our simulations. Electrostatic interactions
account for >100% of AH-my for 1,4-dinitrobenzene (Table 2), so the forcefield perhaps
overpredicts the tendency of 1,4-dinitrobenzene to form inner-sphere complexes with K:
On the other hand, the molecular dynamics estimates for AH,.aqueous were done in the
absence of any ionic species, yet the water-organic interaction enthalpies are also strong

and favorable, so -AH-.,-queous values are large and positive for the nitrobenzenes.

136

Table 2. The F reundlich adsorption coefficients for substituted nitrobenzenes and
the adsorption energies of the corresponding substituents.

 

chemical Kf, Freundlich AHelectrostatic AHvdw AH-m AG of

structure coefficient (l/g) solvation
kJ/mol

 

 

 

 

No,_©—NH, 0.0499 -137 40 -98 13.0
140,001+ 0.182 -91 5.0 -86 6.2
No,_©—oc-, 0.452 -75 17 -58 13.8
No,.©—cno 1-05 -107 6.0 -101 10.4
NOz—O—COOCH3 3.28 -103 -5.0 -108 14.4
rum—Gem”, 2-62 -104 5.0 -99 12.0
140,—ch 1.16 -129 24 -106 11.1
8102.0140- 2.87 -l67 14 -154 19.3
ӣ001, 20 19 38 15.8

TNB -236 2.0 -235 15.8

 

 

The values of Aquucous for all solutes in Table 1 are within a few percent of their
analogous values AHtc-ay in the clay phase. This is really quite surprising given that the
clay interlayer contains a 12 mol/L solution of K+ while the bulk aqueous phase
contained only water and the organic solute. From these results, we anticipate that
Algorithm 1 will be difficult to make useful, because ABE-a, and AHi.aqueous differ by less
than their standard deviations. Thus, the overall adsorption enthalpy AHexpt (a relatively
small number) has to be predicted by subtracting two rather uncertain large numbers, a
situation that does not allow for statistically significant predictive power.

Literature estimates are available for comparison with some values of -AH.-aquc0u,
in Table 1, but none of them are direct. Sandberg et al. (2002) predict 73 kJ/mol for
-AG--aqucous of p-xylene. To convert this to a -AH--aque0us value, we note that Gallicchio et
al. (2000) estimate that, for a range of 11 alkanes, the ratio of AHi.aqueous / AGi.aqueous
averages 1.17 :t 0.04. This implies that -AH.-.aqucous should be near 85 kJ/mol for p-xylene.
which is about twice our estimate. Estimates for the values of -AH-.aqueous for the
substituted nitrobenzenes have to begin with nitrobenzene, since we know of no estimates
for di- or trinitrobenzenes. The -AG,-.aqueous for nitrobenzene is estimated as 76 kJ/mol
(Sandberg et al., 2002). Furthermore, adding a nitro-substituent to benzene, toluene, or
phenol result in consistent changes of -AG-.;,qu¢0us of 23 kJ/mol (Sandberg et al., 2002), so
reasonable guesses for -AG--aqueous values of 1,4-dinitrobenzene and 1,3,5-trinitrobenzene
are 99 and 122 kJ/mol, respectively. Scaling these values by 1.17 (Gallicchio et al.,
2000) converts them to crude approximations of 116 and 143 kJ/mol for the -AH,-aquc0u, of
1,4-dinitrobenzene and 1,3,5-trinitrobenzene, respectively, compared with 149 and 227

kJ/mol in Table 1. Our simulated estimates of -AH-.aqueous for p-xylene are too low while

138

those for the nitrobenzenes appear to be too large in comparison with previous estimates
(Sandberg et al., 2002) that were also based on molecular simulations. A possible
explanation is that force filed parameters used in either the previous study, the present
study, or both were inaccurate or at least disagreed. For example, we used charges of
+0.6170 on N and —0.4280 on O atoms in —N02 groups, while the previous study
(Sandberg et al., 2002) used 0.58 and -0.33, respectively. Apparently, they (Sandberg et
al., 2002) scaled the OPLS-AA forcefield (Price et al., 2001) charges (+0.65 and -0.37)
by 0.9 in order to improve their fits to experimental free energies of hydration. Our
previous quantum mechanical study (Boyd et al., 2001) of 15 small gas-phase
nitroaromatic compounds fit atomic partial charges to the molecular electrostatic
potentials and found that charges on N were +0.66 :t 0.01 and those on 0 were -O.4O i
0.01. In condensed phases like water, these charges would probably be larger in
magnitude due to polarization caused by interactions with solvent molecules.

Relevant estimates for the cavitation component of ABS have been published, but
they are again indirect since AG5 values are more commonly available. Molecular
dynamics simulations have been used several times, the first being (Postma et al., 1982),
to estimate free energies for cavity formation in water, and the numbers come to a
reasonable consensus. Taking cyclohexane as an example (density of the liquid at 20 °C
is 0.779 g/cm3 (Lide, 1995) so molecular volume is 180 A3, similar to dinitrobenzene),
three recent estimates for -AGS, the free energy required to form a cavity in water that can
hold cyclohexane are 66 (Gallicchio et al., 2000), 64 (Sandberg et al., 2002), and 72
kJ/mol (Hdfinger and Zerbetto, 2003). Furthermore, Gallicchio et a1. (2000) estimated

the enthalpic contribution for the first time and found that, for a range of eight C4- to C (,-

139

alkanes, the ratio of AH5 / AGs averages 0.58 i 0.03. Pertinent to Table 1, Sandberg et al.
(2002) predict -68 kJ/mol for A65 of p-xylene, and using AH5 / A05 = 0.58 (Gallicchio et
al., 2000) implies that the cavitation component of AH5 for p-xylene should be near -39
kJ/mol. Consider another way to make this estimate: The cavitation free energy A65 for
a single water molecule is estimated at 15.6 i 0.2 kJ/mol (Sandberg et al., 2002).
Converting this to AH5 using the 0.58 (Gallicchio et al., 2000) conversion factor yields
9.0 kJ/mol per water molecule. Given that the molecular volume of p-xylene (203 A3,
(Lide, 1995) is about seven times that of water, and given that cavitation energies depend
predominantly on molecular volume and surface area (Gallicchio et al., 2000; Hofinger
and Zerbetto, 2003), the cavitation enthalpies AH5 for all the solutes considered here
should be near -60 kJ/mol. A third way to estimate the cavitation energy is through use
of an equivalent spherical radius. Again, cavitation energies like AH5 are thought to
depend more on volume than on any particular shape (Hofinger and Zerbetto, 2003), and
polynomials that are quadratic or cubic in equivalent spherical radius of the cavity often
fit molecular estimates well (Hofinger and Zerbetto, 2003; Sandberg et al., 2002). For p-

xylene, with a volume of 203 A3, the equivalent spherical radius is 3.65 A and a
polynomial fit to alkanes (Sandberg et al., 2002) yields 100 kJ/mol for A65 to form that
sphere in water. Applying AH5 / A65 = 0.58 (Gallicchio et al., 2000) again implies that
AH5 for p-xylene should be about -58 kJ/mol. Since all the solutes of Table 1 are roughly
the same size, -60 kJ/mol should be a reasonable estimate for all AH5 cavitation values
and were also used for the cavitation component of AH5 in the clay since we know of no

relevant estimates there. Thus we assume for now that cavitation energy differences in

140

 

the two phases will be insignificant, which looks like a reasonable assumption given the
magnitudes of some of the other energy terms in Table 1.

The second components (for the interaction of the cavity contents with its
environment) of AH3 (the enthalpy required to take water out of a cavity in the clay
interlayer) and ABS (the enthalpy change for filling an aqueous cavity with water — see
Figure 2) require comment. These numbers were obtained by applying algorithm [5] to
one or more water molecules, then normalizing the resultant interaction energy to the 10
water molecules considered to be leaving the interlayer cavity or entering the aqueous
phase cavity. The range of values in Table 1 comes from ten or more estimates, each
time choosing different subsets, for both the clay phase and the aqueous phase. The
variability and overlap in these values that is evident in Table 1 would again seem to
preclude their use as part of a sum to estimate relatively small adsorption energies.

Are the estimates (Table l) for interaction of water inserted into an aqueous
cavity reasonable? We know of no previous studies that tried to estimate AH3, but A115
values can be estimated. As discussed above, +9.0 kJ/mol is a reasonable estimate for the
cavitation component of AH5, but what about this interaction component. The sum of the
two components should just be the hydration enthalpy (Gallicchio et al., 2000), which is

-44.0 kJ/mol (Lide, 1995). This implies by difference that the interaction component of
AH5 for insertion of water into an aqueous cavity should be -53 kJ/mol. Our estimates of
-60 to -72 kJ/mol are somewhat more negative than this value, but within reason. Our
value may be a bit large because we did not allow the water molecules bordering the
cavity to reorient their H-bonding patterns during our estimate: We simply deleted a

cavity of water molecules and recalculated the energy.

141

The sum of AH3 and AH5 might be compared with experimental data, since
AH3+AH5 should approximate the energy required to move water out of the smectite
interlayer and into bulk water. That is, the sum of AH3+AH5 refers to the reaction
(equation 2)

C lay(aq) + cavitymm (—> Clay-cavity(aq) + 1120-3,” [13]

which is equivalent to

cavitym) <-—) cavity(c-ay) [14]

which seems equivalent to

H2()(clay) ('9 H20(aq) [1 5]

The measured heat of immersion for the first layer of water loading on K-
smectites is about -1 to -4 kJ/mol of water (Berend et al., 1995), for an average of about
-2.5 kJ/mol, so AH; + AH5 should be about +2.5 kJ/mol of water. Thus, for the 10 water
molecules used to represent each cavity, we would expect the magnitude of the insertion
component of AH3 to be about 25 kJ/mol larger than that of AHs. Instead, there is little
difference and, if anything, the central tendency of AH3 is to have a slightly smaller
magnitude, although any differences are again swamped out by the overall variability.

The preliminary results of Table 1 are encouraging in that a) our algorithm yields
adsorption enthalpies of the correct sign for for agreement with the experimental data (Li
et al., 2004) for 1,4-dinitrobenzene and 1,3,5-trinitrobenzene (and, for what it is worth,

the standard deviations for the computed estimates overlap with the experimental

142

estimates), and b) adsorption enthalpies for nitroaromatics and p-xylene are predicted to
be significantly different, even at this preliminary stage. A significant challenge in the
path to making the algorithm USCfiJI is the size of the error bars at each step because they
cause the overall Eexp- estimates to have an uncertainty that will ofien be larger than the
value itself. Further work is needed to understand and minimize the sources of
uncertainty in the energy estimates.

The interaction energy was dominated by electrostatic rather than van der Waals
terms (Table 2), in accordance with the earlier concepts that complexation interactions
between interlayer cations, K+ in this case, and the nitroaromatic molecules are critical.
These complexation energies also seem to dwarf the aqueous solvation energies of the
nitroaromatics as a contributor to the overall adsorption energy.

To test the validity of the algorithm 1, we used the data from Boyd et a1. (2001)
(Figure 3). The numbers along the right hand side are our simulated interaction energies
between organic and the smectite. Note that they follow a reasonable trend with strong
interaction energy for strongly sorbing ones and smaller interaction energy for poorly
sorbing ones, although some of them are out of order. But, in general, the trend of

predicted energies for experimental sorption results is promising.

143

 

 

 

 

 

 

 

 

 

 

 

18 —
16 _ ~234zt17
g: 14 ~
a: ——TNB*
5.. 12 — ——p-N02
E 10 p-COOCH3
8 -163i16 p'ggfig3
8 8 - -108:t14 p'
E 99:1:15 _p'OH
is; 6 ‘ ' ——p-CN
< 4 - —p-NH2
-106:l:18
2 K-58i16
" —-86:l:16
0 - 38:19
0 5 10
Equilibrium Conc., mglL

 

 

 

Figure 3: Validity of algorithm with experimental data.

For comparison, p-xylene, which is dimethylbenzene, which has an interaction
energy of +38 kJ/mol. A positive number tells us that it is unfavorable and most of the
nitro compounds are favorable. Xylene cannot form complexes and so its interaction

energy is positive.

(All the calculation energies are for monolayer systems where complexation
interactions will be maximum.)

All-m values were quite favorable for all nitroaromatic compounds but
unfavorable for p—xylene. Thus, the algorithm shows promise for computation of useful
AH-m values, but the adsorption energies depend on factors other than AH-m. For
example, the energies for removal of the organic from aqueous solution must be a
component of the adsorption energy, and this factor shows a range of 6 to 19 kJ/mol (see
Table 2).

In our simulations, the energies for removing organics from their cavities in bulk
water were estimated at +149 i 21 kJ/mol for p-NOZ, +114 i 16 kJ/mol for p-COOC H 3,
+227 1: 23 kJ/mol for TNB, and +40 3: 9 kJ/mol for p-xylene (Table 1). Therefore,
adsorption enthalpies are estimated to be -5 i 27 for p-NOz, +6 i 22 for p-COOCH3, -8 i

29 for TNB, and +78 i 15 for p-xylene.
A goal of each simulation was a 12.3 A d-spacing for the clay-water—organic

system. Using p-COOCH3 as a test for the sensitivity of AHint, four different simulations

were performed at three water contents that each resulted in a 12.3 A d-spacing (Figure

4). Variability in AHint was about 2 kJ/mol for a given water content but ~21 kJ/mol
among the different water contents. Within a given simulation, standard deviations for

AHim among the 150 observations were ~ 17 kJ/mol.

145

 

 

14.93

d-spacing, A

12.87
1225 12.31 1239

A
Ar, c

15.31 12333 - ”-47

 

 

 

42 43 44 45 47 50 55 89

No. of interlayer waters

 

 

 

Figure 4. Sensitivity Analysis.

Algorithm 2 validation

In applying algorithm 2, we can make use of our discovery using algorithm 1 that
AHgm- and AH5-n- cancel each other (Table 1). This simplifies the estimation of an

experimental adsorption enthalpy because only simulations of the clay phase are

necessary, using

AHexpt = AH3cav + AHi-clay + AI"Isubl - AI"Idis [16]

Table 3 shows the results of applying algorithm 2 to the two nitroaromatic
compounds for which we have measured the thermodynamics of adsorption to K-smectite
(Li et al., 2004). Note that, while the predictions are not quantitative, they again have the

proper sign and now have the advantage that the predicted adsorption enthalpy values

themselves are larger than their standard deviations.

146

 

 

 

 

 

 

 

 

 

 

 

 

$8.0. :8 do so ._
A88 8284 use 8828
s : 4 mm 8+ 053-8
«How. 2 H 5. mm ~Hno~ S Hmmm- oo+ mZ-_.
.3:- 2 4mm- 8 In; 2 32- 8+ Noz.Q
axom< 3qu
988.802 8885-”— mar? “3324 8.8.34 23:4 8:53:80

 

anon—«5.8 8:988: 73.8.3898 «88:8 3 N Bantam—a .8 8=3=AE< .m 038,—.

I47

Thus, we conclude that the method for simulating adsorption enthalpies shows
promise. For example, predictions for nitroaromatic compounds versus p-xylene are
significantly different. Further work is needed that will attempt to isolate and minimize

sources of uncertainty to find out how precise the enthalpy predictions can become.

148

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