DESORPTION AND BIODEGRADATION EXPERIMENTS:

DESORPTION AND BIODEGRADATION EXPERIMENTS:
DESORPTION AND BIODEGRADATION EXPERIMENTS:
1) EFFECT OF APPLICATION SOLVENTS ON NITRIFYING BACTERIA,
2) EFFECT OF SURFACTANTS ON RELEASE AND BIODEGRADATION
OF SIRO N G LY BOUND SOIL RESIDUES OF
ATRAZTNE AND NAPHTHALENE
by
Jerry Lee Miller
Dissertation Submitted to the faculty o f the
DEPARTMENT OF SOIL, W ATER AND ENVIRONMENTAL SCIENCE
In Partial Fulfillment o f the Requirements
For the Degree o f
Doctor o f Philosophy
In the Graduate College
THE UNIVERSITY OF ARIZONA
1997
2
THE UNIVERSITY OF ARIZONA ®
GRADUATE COLLEGE
As members of the Final Examination Committee, we certify that we have
read the dissertation prepared by_______ Jerry Lee Miller_____________
entitled
Desorption and Biodegradation E x p e r i m e n t s : ____________
1) Effect of Application Solvents on Nitrifying Bacteria.
2)_Effect of Surfactants on Release and Biodegradation of
Strongly Bound Soil Residues of Atrazine and Naphthalene
and recommend that it be accepted as fulfilling the dissertation
requirement for the Degree of ____ _____ Doctor of Philosophy
Date
\7J/nr
Date
17 A/oiJ ?7
# 7 /you/ <77
Date
Dr. Mark Iv Brusseau
Date
Dr. Michael D. Bradley
Date
(7 4/ 01/
<?7
Final approval and acceptance of this dissertation is contingent upon
the candidate's submission of the final copy of the dissertation to the
Graduate College.
I hereby certify that I have read this dissertation prepared under my
direction and recommend that it be accepted as fulfilling the dissertation
requirement.
Dissertation Director
Dr. Raina. M. Miller
Date
3
STATEMENT OF AUTHOR
This dissertation has been submitted in partial fulfLUment o f requirements for an
advanced degree at The University o f Arizona and is deposited in the University Library
to be made available to borrowers under rules o f the Library.
B rief quotations from this dissertation are allowable without special permission,
provided that accurate acknowledgment o f the source is made. Requests for permission
for extended quotation from or reproduction o f this manuscript in whole or part may be
granted by the head o f the major department o f the Dean o f the Graduate College when in
his or her judgment the proposed use o f the material is in the interests o f scholarship. In
all other instances, however, permission must be obtained from the author/
////
r
Signed:
4
ACKNOWLEDGMENTS
I would like to personally and publicly acknowledge my debt to Dr. Raina Miller
[no relation : ) ] without whose advice and discipline I would not have achieved this
honor. Eter patience and financial support are greatly appreciated. Drs. Tom Thompson
and M ark Brusseau have also contributed significantly by helping me with editorial
comments (Mark) as well as money (Tom).
I need to say a heartfelt thank you to my friends and co-workers who have helped
me grow over the years. Once again, James Ratchford and Sally Edwards merit mention.
Philip Gekas, John Harrison, John Barrios, Lon Hoffman, Fiona Jordan, Mike Young,
Sheri Musil, Joe Piatt, Eric McGee, AJ Bittner, Jill Champion, M ary Huber, Yimin Zhang,
Guiyun Bai, Adria Bodour, Tom Cooley, Andreas (Andy), Cheryl (and Frasier), Wendy
Kilgore (and Shadow) and many others (esp. from my lab) have all influenced my
decisions as I worked toward this goal. I will greatly miss dog hour over at Himmel Park.
I met many, many wonderful people there (and their dogs). I will also miss volunteer
work for the Native Seed Search seed bank. Then goal o f preservation o f crop genetics
and diversity is indeed laudable.
The support, willing and otherwise, from my family has also been invaluable. I
cannot say enough to thank my M other for Her Support. Mike, Dave, Randy; again I
have no w ords adequate to expression my gratitude for all you have done over the years.
I would also like to thank my father for passing along to me his deep understanding o f the
state where I have spent so many wonderful hours biking, hunting, and camping. Arizona
has no equal in the nation. Todd, Cara Leigh, I am real glad to have been able to spend
this time here with ya’ll. I hope to be able to spend more.
W hat more can I say, I like it here in Tucson and don’t really want to leave, but
the fates have made it clear they are preparing me for another drastic change. Where l go
from here, I do not know. I only know that Selene’s death and the earning o f my
doctorate indicate that it is away from here. I will miss the way desert smells after the
rain, the way sun shines 300+ days a year, the way everything in the desert is well armed.
And I sure as hell will miss hiking in the desert with my dog. The Raleigh Rose Garden
still resides comfortably in my memory and so will southern Arizona. : )
)
5
DEDICATION
I dedicate this dissertation, and indeed, the rest o f my life, in honor o f the memory
o f Selene, my faithful dog o f 5% years, who suffered a tragic end (under the wheels o f a
car) on Friday, 9-19-97. Had it not been for her, I would never have developed both the
discipline and the patience to complete my task here. She taught me so many things. First,
and foremost, she taught me how to deal with my anger so that it did not hurt her and it
did not hurt me. She made me have patience, even when I did not have it in me. By
watching her joy, I learned to love life for each and every moment. She loved me
unconditionally and showed me how to do likewise. The insights that I have gained as a
result o f loving and caring for her mean more to me than any o f the lessons or education I
have received at the University, for only now am I actually wiser.
Oh Selene, I miss you so much! I wake up every sunrise, expecting to feel your
nose on my elbow, but it is no longer there. I miss the way you smell and your happy
smile. I cannot forget how you used to make me laugh and how you used to make me
mad. I will always remember you and I promise to take these lessons forward as a better
man.
When life seemed to have no value, she made it precious.
When I could not see the way forward, she led.
When I faltered, she forgave.
Goodbye, Little Girl - good hunting!
6
TABLE OF CONTENTS
LIST OF ILLU STRA TIO N S............................................ :................... .................
8
LIST OF TA BLES.... .................................................................................................
9
ABSTRACT.....................
10
1.
INTRODUCTION............................................................................................ ....... 12
Effect o f Application Solvents ..................................................................... 13
Sorbed Residue Experim ents......... ................................................................. 14
2.
LITERATURE REVIEW OF BOUND R E SID U E S................................... ........ 16
H istory... ......................................... .
C lay...................................................
Soil Organic M a tte r........................
Effects o f A ging..................
Mechanisms of sorption....
Bioavailability....... .
Removal and Quantification
Aged Residue S tud ies........ :..........
Surfactant S tudies..........................
A trazine...............................
Naphthalene........................,
Summary......................... ................
3.
PRESENT STUDIES
Effect o f Application Solvents.
Introduction...... .........
Materials and Methods
R esults........................
Conclusion..................
Sorbed Residue Experiments..
Introduction................
Materials and Methods
R esults........................
Conclusion..................
4.
16
20
23
24
25
26
28
32
33
35
36,
36
38
38
38
38
39
39
40
40
41
41
42
COUPLED INTERACTIONS BETWEEN CONCENTRATION AND
SPECIES OF NITROGEN AND ORGANIC CONTAMINANTS
.......... ,
43
Introduction.................................. ...................... .;........................................ 43
7
Materials and M ethods...........................
Results and Discussion............................................
Atrazine...............................................................................................
Naphthalene.............................................................
APPENDIX A: EFFECT O f APPLICATION SOLVENTS ON
HETEROTROPfflC AND NITRIFYING POPULATIONS
IN SOIL M ICROCOSM S..............
45
48
48
53
58
APPENDIX B: DESORPTION AND BIODEGRADATION OF STRONGLY
BOUND SOIL RESIDUES OF ATRAZINE AND
NAPHTHALENE USING VARIOUS SURFACTANTS.......... 65
REFERENCES.............................
87
8
LIST OF ILLUSTRATIONS
Figure 1. Extractable Radioactivity, A trazine....................................................................... 49
Figure 2. Bound Residues o f A trazine........... ...................................................................... 50
Figure 3. Production o f 14C 0 2 from A trazine........................................................... ........... 51
Figure 4. Extractable Radioactivity, N aphthalene.................................................................54
Figure 5. Bound Residues o f N aphthalene........... ........................... ...................... ............ 55
Figure 6. Production o f 14C 0 2 from Naphthalene .............................................................. 56
1
LIST OF TABLES
Table 1. Soil Characteristics
10
ABSTRACT
To quantify the interactions, between nitrogenous fertilizers and organic
compounds (both pesticides and industrial pollutants), coupled studies were conducted
between nitrogen species and concentration and organic species and concentration. It was
determined that atrazine and naphthalene rate (1.0, 10 or 100 mg/kg soil) appeared to
have no effect on nitrification and that nitrogen rate (1.0, 10, or 100 mg ammonium
sulfate/kg soil) had little or no effect on degradation o f either atrazine or naphthalene.
During these coupled experiments, it was discovered that the use o f a volatile solvent to
apply the organic compound o f interest resulted in an inhibition o f the nitrifying bacteria.
Tests o f a variety o f solvents revealed that use o f dichloromethane as the application
solvent almost totally inhibited nitrification for 35 days. While methanol inhibited
nitrification for up to 3 weeks, after 35 days production o f nitrate was similar to a water
control. Chlorinated solvents inhibited nitrification more than methanol or acetonitrile.
These results indicate that for coupled organic/nitrogen experiments care must be taken
not to disrupt the more sensitive members o f the microbial community. Experiments were
also performed to remove and biodegrade strongly sorbed residues o f atrazine and
naphthalene. Soils were aged 16 and 8 weeks, respectively, then solvent extracted with
MeOH, resulting in soils containing bound residues o f each compound. Three surfactants
were tested for their ability to enhance removal o f these bound residues; sodium dodecyl
sulfate (SDS)(anionic, CMC= 2336 mg/L), Tween 80 (non-ionic, CMC = 13-15. mg/L),
and p-cyclodextrin. Up to 66% o f the previously unextractable naphthalene residues were
released by 5 washes using a 10,000 mg/L solution o f SDS. Between 6 and 27% o f the
11
)
released naphthalene residues were subsequently biodegraded to C 0 2. Optimum removal
and biodegradation o f naphthalene residues occurred between 4,000 and 8,000 mg/L
SDS. All three surfactants failed to enhance release o f atrazine residues while only SDS
enhanced removal o f naphthalene residues.
,
12
CHAPTER 1
INTRODUCTION
The biodegradation o f organic compounds in the soil is dependant on many
variables including nutrient status. During the course o f research into the interactions
between nitrogen species and concentration and organic contaminant species and
concentration, it was observed that the application solvent used to apply the organic
contaminant o f interest was exhibiting negative effects on the nitrogen cycle. The
experiments investigating the effects o f application solvents on nitrification are described
in detail in Appendix A o f this dissertation. The original research into the coupled
interactions o f nitrogen and organic contaminants provided no publishable conclusions
and is summarized in Chapter 4. The solvent-extracted soils from those experiments were
saved for further experiments to determine if 3 classes o f surfactant were capable o f
enhancing removal o f bound residues o f atrazine and naphthalene which those soils
contained. The desorption and biodegradation experiments are described in detail in
Appendix B o f this dissertation. This chapter will describe the importance and
contribution o f the two lines o f research detailed in the appendices. Chapter 3 will
summarize the experiments and give a synopsis o f the important conclusions. Chapter 2
contains a literature review on the subject o f bound residues.
13
Effect o f Application Solvents
Biodegradation experiments involving organic compounds o f low w ater solubility
often employ a more volatile organic compound, usually a solvent, to apply the organic
contaminants. Various application solvents have been used in a variety o f biodegradation
experiments, including methanol (Zhongbao et a l, 1991; Miller et al., 1997), acetone
(Guerin and Boyd, 1990), and acetonitrile (Hannan, 1995), and extensive use has been
made o f the chlorinated methanes (Miller et al., 1988; Churchill et al., 1995, Tsomides et
al., 1995). The literature indicated that because the nature o f the original w ork was
unique (the coupled interactions between nitrogen and organics) the effect o f application
solvents from standard organic biodegradation experiments on the nitrogen cycle had
largely gone undetected.
As biodegradation experiments become more sophisticated and laboratories try to
develop novel ways to detect or ensure degradation o f organic contaminants, it is critical
to make sure that experimental procedures do not interfere with accurate and precise
interpretation. Because so many labs from around the country use these standard
application techniques, the potential for mistaken interpretation is high. Therefore, work
was designed and performed to determine if the standard method o f conducting organic
biodegradation experiments was having a negative impact on a very specialized, but very
important, class o f soil microbes; the nitrifying bacteria. These experiments provide a
missing tool for critical analysis o f biodegradation data.
14
The objectives were to determine:
-1 )
if the solvents commonly used to apply organic contaminants to soil in organic
biodegradation experiments inhibited nitrification,
2)
which step o f nitiificatiori was inhibited,
3)
was the inhibition reversible,
4)
if the same solvents inhibited the degradation o f naphthalene. .
The published results o f this w ork are presented in Appendix A. This work was a
)
collaborative effort between myself and Michael Sardo, the junior author o f the
publication. Mike did all o f the nitrogen analysis while I analyzed for organic degradation.
Setup and sampling were performed jointly while the compilation, interpretation and
presentation o f the data were performed by myself.
Sorbed Residue Experiments
The soil samples from the coupled nitrogen-organic experiments described in
Chapter 4 were saved and utilized for further research on the residues remaining bound to
them. A review o f the literature on bound residues, their desorption and biodegradation,
can be found in Chapter 2 o f this dissertation. Strongly bound residues o f organic
contaminants have been identified as a reservoir o f long-term contamination. The release
o f these bound residues into groundwater supplies has become an issue o f environmental
concern. Currently there is a large research interest in the development o f methods to
enhance the extraction o f these bound residues and to economically remediate both soils
15
and groundwater. The current experiments add important information regarding aged
residues o f naphthalene and atrazine in soil to that body o f research knowledge. Aged and
strongly bound residues are o f environmental interest due to their potential for release at
(
some time in the more distant future. Therefore experiments were performed to determine
the effect o f various classes o f surfactant on the desorption and biodegradation o f bound
residues o f atrazine and naphthalene. Aged, methanol-extracted soils were used to
simulate strongly bound residues. These experiments provide valuable information on an
environmentally important fraction o f residual soil contamination.
The objectives o f the study were to determine:
1)
whether a model anionic, non-ionic or cyclodextrin surfactant could enhance the
release o f bound residues o f atrazine or naphthalene from aged and methanolextracted soils,
2)
if the released residues could be mineralized to 14C 0 2.
The results from this w ork are presented in Appendix B. All experiments were designed
and carried out by the author.
16
CHAPTER 2
LITERATURE REVIEW OF BOUND RESIDUES
History
The value o f synthetic organic chemicals as pesticides was apparent by 1927 with
the discovery that ethylene dichloride functioned as a fumigant (Ware, 1978). This was
followed by pentachlorophenol as a termaticide (1936), 2,4, dichlorophenoxyacetic acid
(2,4-D) as an herbicide (1942), and zineb (a dithiocarbamate) as a fungicide (1943)(Ware,
1978). The application o f these and other synthetic chemicals to agricultural soils
skyrocketed soon after World W ar II as military experience and chemical production
capacity was turned to peacetime production o f agrichemicals. By 1965, 150 x 106
kilograms o f pesticides were being manufactured for the agricultural market (Ware,
1978). This marked the first time in history that soils had been exposed to significant
amounts o f novel organic, often highly chlorinated, chemical structures which were not
easily broken down by soil and aquatic microorganisms. As a result, synthetic organic
compounds began to accumulate in the environment. One o f the first episodes o f just such
an environmental accumulation was the foaming o f the Ohio river in 1953 (Swisher,
1987) due to alkyl benzene sulfonate surfactants left undegraded by the microorganisms
in upriver sewage treatment plants. Other environmental accumulations o f synthetic
organic pesticides have occurred, including DDT in remote island ecosystems (Auman et
a l, 1997), PCBs in harbor sediments (Lake et a!, 1996), and, o f course, agrichemicals in
soils. These latter are mostly in the form o f residues bound (usually) to either clay or soil
17
organic matter (SOM). This review summarizes briefly the history o f bound residues, and
in more detail, the current state o f knowledge regarding bound residues, and especially
techniques to release and degrade them.
While it had been understood for some time that when synthetic organic chemicals
were added to soil systems some portion o f the compound did not perform its intended
function, early researchers did not (or could not) look for, and therefore did not find,
significant residues bound to soil. Dissipation was an ambiguous term often used in the
early literature to describe the inability to find the parent compound anymore (Burnside et
a l, 1961; Burnside et a l, 1963; Ghadiri et a l, 1984). The development o f radiolabeled
carbon pesticides made it possible to track a compound thru the entire sequence o f
degradation all the way to C 0 2 (Couch et al., 1965; Kaufman et al., 1965).
Radiolabeling also made it possible to localize and quantify bound residues (McCormick
and Hiltbold, 1966; Benyon et al., 1972). Soils could be mechanically or chemically
fractionated and the radioactive residue within each compartment determined with great
precision (Huang et a l, 1984). Environmental fate studies which included a full mass
balance began in the mid-1960's starting with the s-triazine herbicides (McCormick and
Hiltbold, 1966; Benyon et a l, 1972).
Sorption isotherms performed on separate soil components started to differentiate
binding sites and mechanisms (Weber et a l, 1969c; Weber, 1980b). B oth clay and organic
matter were implicated as sites where organic compounds were sorbed by a variety of
mechanisms (Weber et a l, 1969a; Kozak et al., 1983.). Hydrous metal oxides and
allophanes also contribute to pesticide sorption in soils but their contribution (except in
18
very special cases) is small and they will be discussed only briefly. Cation (and anion)
exchange capacity rapidly became a measure o f the sorptive capacity o f a given soil for
the ionic organic molecules (cationic, protonated bases, weak acids) while the SOM
content has been used extensively to estimate sorption o f the non-ionic organic
compounds (NOCs). Once it was recognized that chemicals bound to soil constituents
formed a significant reservoir o f future contamination, workers started attempting to
remove them for identification and quantification. Bound residues have been defined in
several ways, most often relating to the inability to remove the compound using an
organic solvent. Khan (1982) provides the following definition o f bound residues: “that
unextractable and chemically unidentifiable pesticide residue remaining in fulvic acid,
humic acid, and humin fractions after exhaustive extraction with nonpolar organic and
polar solvents”.
Many schemes have been tried for the removal o f bound organic compounds from
soil. While there are only a few cationic pesticides, they can often be released by the use
o f competitive solutes containing cations such as EDTA or other organic cations (Weber
et a l, 1969b). Non-ionic chemicals are more common as pesticides and the technique
which has seen the most use for removing NOCs from soil is liquid/hquid partition to
redissolve the sorbed organic compound into a similar compound such as a solvent.
Methanol, aqueous solutions o f methanol (atrazine, Huang and Pignatello, 1990),
acetonitrile (atrazine and others, Smith, 1981; primisulfiiron, Miller et al., 1997a),
acetone (Scheunert and Mansour, .1992), and chlorinated methanes (dichloromethane for
3,4-dichloroaniline, Yokel et a l, 1994) are the most common and most successful. Super­
19
critical C 0 2 and methane have proven to be as good as solvent extraction, but often no
better depending on the age o f the samples (Capriel et al., 1986; van der Velde et al.,
1994; Dean, 1996). Various surfactants have also been tried with varying degrees o f
success (Aronstein et a l, 1991; Javert, 1991; Laha and Luthy, 1992; Van Dyke et a l,
1993a; Kommalapati and Roy, 1996). Extraction using cyclodextrin has also been
attempted (Miller et a l, in review). Pyrolysis at various temperatures has produced
somewhat better results (Khan and Hamilton, 1980, Robbat et a l, 1992; Ostiz and Khan,
1994; Richnow et al., 1995). Because all o f these techniques leave some fraction (often
greater than 50%) undisturbed, or unidentified, while employing different theories o f
solvation, the debate about the specific mechanisms o f sorption continues today.
After extraction by one or more o f these techniques, the compound remaining
attached to the soil is referred to as “the bound fraction”. Several techniques have
attempted to remove this fraction. Strong bases and acids fractionate the humic and fulvic
SOM fractions respectively, while the humin fraction can be further be classified as
alcohol soluble or insoluble. In addition to the problem o f sorption o f the parent
compound, sorption o f metabolites is also known to occur (Schiavon, 1988; Clay and
Koskinen, 1990; Brouwer et al., 1992; Smith and Aubin, 1993; Gan et a l, 1994). While
all mechanisms o f sorption will be discussed, this discussion will focus mostly on atrazine
and the NOCs such as naphthalene.
20
Clay
Clays sorb organic compounds through a variety o f interactions. These bonds
range in strength from hydrogen bonds to anionic and cationic bonds, however, the
clearest mechanism o f sorption o f organic compounds by clays is the electrostatic
attraction o f cationic compounds to the overall negative charge o f the clay particle. This
negative charge is created by negatively charged sites from the fracture o f the
ahuninosihcate tetrahedra, so called broken edge sites, and by isomorphous substitution
within the lattice structure itself. Expanding clays such as montmorillonite have large
internal surfaces which are available for cationic bonding as well. These clays often have a
large capacity for sorption (as measured by CEC). In. some cases, hydrophobic organic
compounds were found to sorb so strongly to the clays present in soil that it was thought
that sorption was more or less permanent (Weber and Coble, 1968; W eber et al., 1969a;
Marengo et al., 1997). In particular, the charged pesticides paraquat and diquat not only
engaged in cation exchange with negatively charged sites on clay surfaces but were found
to be capable o f fitting into the interlayer spaces o f the expanding clays (montmorillonite
and vermicuhte) and becoming essentially non-exchangeable or permanently bound,
unavailable even to microbes (Weber and Coble, 1968; Weber et al., 1969a). X-ray
diffraction studies indicated that sorption on montmorillonite occurred when the planar
pyridine ring was parallel to the silicate sheets and fit neatly in between them (Weed and
Weber. 1968). Sorption o f this nature on montmorillonite can become irreversible, as the
interlayer spacing will collapse to approximately the size o f a methyl group as the sorbed
concentration approaches the internal CEC (Weed and Weber, 1968). On the other hand.
.
21
sorption to non-expanding clays is often partially or fully reversible. Paraquat adsorbed to
kaolinite or vermiculite was still capable o f injuring cucumber seedlings (Weber and Scott
1966; W eber et al., 1969a). The practical impact o f cation exchange is that cations with a
stronger affinity for a given site can displace bound cations back into solution. Cationic
pesticides bound by this simple electrostatic mechanism can usually be displaced,
especially by trivalent vs monovalent cations (Na+ < Al3+). Weber et al. (1969b) used N(4-pyridyl)-pyridmium chloride to displace diquat from montmoriUonite, but could do so
i
only when the concentration o f diquat sorbed to the clay was very high and the
concentration o f the displacing cation very high as well. However, magnitude o f sorption
is not necessarily related to the strength o f the bond. Yamane and Green (1972) found
that while ametryne sorbed to clay minerals and soils in greater quantities than atrazine,
atrazine exhibited the more energetic (stronger) bonds. Sorption o f cationic organic
species is highly pH dependant and other ionizable herbicides for which sorption is
dependant on pH include buthidazole, tebuthiuron, ftouridone, metribuzin and prometryu
(Weber, 1980).
Basic herbicides must protonate in order to be positively charged and engage in
cationic bonding. As result, sorption on these sites is also often pH dependant (Green,
1974; W eed and Weber, 1974) and sorption o f these compounds often occurs primarily at
or below the pKa. Atrazine is a weak base and is often protonated in soil solution (Clay
and Koskinen, 1990). They showed increased sorption to a low pH soil when compared
to a high pH soil. However, SOM was not removed in these experiments and may have
contributed to this effect. Laird et al. (1992) have demonstrated that atrazine sorption to
22
smectite clays was primarily as the neutral species (there was no effect o f p H on sorption)
and correlated well (r = 0.82) with surface charge density (SCS) and surface area (SA).
Clay et al. (1988) also showed httle (<30%) effect o f pH on atrazine sorption to 3 soils.
However, Gilchrist et al. (1993) demonstrated using pure clay minerals that atrazine
sorption to expanding clays was due primarily to a second mechanism, intraparticle
diffusion and entrapment in the interlayer spaces o f montmorillonite. Ca2+ was found to
inhibit this interlayer diffusion more than Na+. This specific effect was apparently a result
o f the ordering o f w ater molecules around each cation as opposed to a charge density
phenomena. However, they too, noticed a relationship between SCS and sorption.
Anionic bonds which form with weak acids are similar in strength but with an
inverse dependence on pH when compared to cationic bonds. This anion exchange
capacity (AEC) arises mostly from protonation o f hydroxide surfaces on kaolinite,
amorphous aluminosihcates (allophanes) and hydrous metal oxides at low pH.
Compounds such as 2,4,D, dinoseb and picloram all show anion adsorption (Green,
1974). Imazaquin has been shown to be retained in soils by an anionic binding mechanism
(Regitano et al., 1997).
Hydrogen bonds also are important hi soils both to clays and to SOM. Broken
edges on the clay lattice present an opportunity for hydrogen bonds to form It is thought
that hydrogen bonding accounts for much o f the sorption o f polar but non-ionic organic
compounds to clays (Green, 1974). Binding o f these compounds is often water-content
dependant as w ater competes for polar sites in the soil. W ater can also act as bridge
between positive charges on clays and carboxyl groups o f pesticides. This type o f bonding
23
has been shown for malathion on hydrated montmorillonite saturated with various cations
(Bowman et a l, 1970) with trivalent cations forming the strongest attractions.'Hydrogen
bonding has also been proposed as the mechanism o f binding for phenols in soils (Boyd,
1982). Sorption o f non-ionic compounds to clay can also occur if there is an organic
phase bound to the clay directly (Trairia and Onken, 1991). Interestingly, some
researchers consider residues bound to clays as not “bound” in the classical sense
because appropriate pH and salt concentration can often displace them (Khan, 1982).
Soil Organic M atter
h i the United States, 74 agricultural chemicals have been found in the ground
waters o f 38 states (Williams et a l, 1988) frequently including atrazine (Maas et a l,
1995). Significant amounts o f these compounds are commonly found sorbed to SOM,
forming a reservoir o f contamination which will continue to pose a threat to groundwater
(Yee et a l, 1985). Organic compounds are bound to SOM by a variety o f mechanisms
(Chiou, 1989; Hassett and Banwart, 1989) including covalent bonding (4-chloroaniline,
Scheunert et a l, 1992) and hydrogen bonding (prometryne, Weber, 1980). Li fact, all o f
the binding mechanisms discussed for clay are operational to some degree or another in
SOM as well. SOM contains multiple fimctiorial groups which are capable Of existing in
both disassociated and associated forms and therefore are capable o f participating a
variety o f bonding reactions. Atrazine (a weak base) sorption to SOM has been
postulated as a combination o f complexing with unidentified functional groups o f SOM as
the neutral species and cationic adsorption o f the charged ionic species (Weber et a l,
24
1969c; Hayes, 1970). However, the chemicals o f concern which are found most often
bound to SOM are the NOCs.
Sorption o f NOCs by soil correlates most closely with soil organic matter content
(Karickoff et a l, 1979; Chiou, 1989; Hassett and Banwart, 1989). Binding to the soil has
been shown to decline with depth as SOM declines (Locke and Harper, 1991; Miller et
a l, 1997). The extent o f partitioning out o f the aqueous phase is controlled by the
fraction o f organic carbon (foc) present in the soil and the hydrophobicity o f the
compound, usually measured as the Kow(octanol/water partition coefficient)(Karikoff et
a l, 1979; Jafvert, 1991). Equations have been developed which relate lab derived Kow
values and the foc to the measured distribution o f NOCs between aqueous and sorbed
phases (Karikoff et a l, 1981; Hassett and Banwart, 1989; Jafvert, 1991). An excellent
schematic for the construction and determination o f the appropriate distribution
coefficients is presented in Green and Karickoff (1990). Additions o f organic matter o f
various kinds (surfactants { « C M C } , Liu et al., 1991; n-heterocycles, Traina and Gnken,
1991; humified evergreen oak biomass, Martinez-Inigo, 1992; ripe compost, Kastner et
a l, 1995) has also been shown to increase the sorption o f NOCs to soils. The more highly
composted the material, the greater the increase in bound residues (Martinez-Inigo, 1992;
Kastner et a l, 1995).
Effects o f aging
Levels o f sorption o f NOCs have been shown to increase as contact time between
the contaminant and the soil increases (Karikoff and Morris, 1985; Steinberg et a l, 1987;
25
Pignatello, 1989; Gamerduiger et al, 1991; Smith et al., 1992; Borglin et al, 1996;
Madden et al, 1996). Bi-phasic sorption has been noted for both atrazine and NOCs
wherein some fraction o f the compound comes into a rapid equihbrium with the soil while
some o f the compound continues to sorb more slowly (Wauchope and Myers, 1985;
Gilchrist et al, 1993; Pignatello et a l, 1993). Various multi-compartment models have
been proposed to explain bi-phasic sorption (Karickhoff, 1979; Gamerdinger et a l, 1991;
Pignatello and Huang, 1991) including intra-particle diffusion (Wu and Gschwend, 1986;
Scow and Alexander, 1992) and intra-organic matter diffusion (Brusseau et al 1991).
Two1compartment mathematical models which explicitly account for this non-equihbrium
sorption/desorption are used to model biodegradation data for NOCs in the presence o f
soil constituents (Scow et a l, 1986; Brusseau et al, 1991; Scow and Alexander, 1992).
The fraction which is slow to come to sorptive equihbrium is also slow to come to
desorptive equihbrium and the fraction o f the compound undergoing slow sorption has
been termed irreversibly bound (naphthalene, Al-Bashir et a l 1990), slowly reversible
.
,
(atrazine and metolachlor, Pignatello and Huang, 1991) and variable depending on
compound, sorbent and time (Guerin and Boyd, 1992).
Mechanism o f sorption
The mechanism o f sorption o f NOCs to SOM is thought to be simple diffusion o f
the NOC into the complex, three dimensional matrix o f SOM (Khan, 1982a; Chiou, 1989;
Khan, 1991). This partitioning is affected primarily by two forces; enthalpy related forces
which affect the relative affinity o f the solute (the contaminant) for the sorbent (SOM),
26
and entropy-driven processes where an increase in disorder occurs when the solvated
(
shell o f w ater around the contaminant in solution collapses as the contaminant partitions
out o f the aqueous phase (Hamaker and Thompson, 1972; Chiou 1989; Hassett and
Banwart, 1989; Traina and Onken, 1991). Examples o f enthalpy-related forces include
London-Van der Waals forces, hydrogen bonds, ligand exchange, and dipole-dipole
interactions. For entropy related forces, SOM is significantly disordered already so that
the driving force tow ard order in the water phase easily overcomes the trend toward
disorder in the SOM (ten Hulscher and Comelissen, 1996). This is illustrated by the
decreasing trend in sorption with increasing temperature for many compounds as water
solubility increases, ie. as disorder within the water increases (ten Hulscher and
Comelissen, 1996). Strek and Weber (1982) have shown that activated carbon is a
stronger sorbent than SOM, montmorillonite clay, or Lakeland sand. Activated carbon
can be considered a completely random structure so that the entropy force is virtually nil
(eg., Strek and Weber, 1982).
Bioavailability
I f diffusion is the mechanism o f sorption under consideration then it appears that
molecules which have had considerable time to diffuse into sites deep within SOM may
take as long, if not longer, to diffuse back out. Borglin et al. (1996) showed desorption
times for hexachlorobenzene and 2 PCBs from contaminated sediments can be on the
'
>
order o f several years. Krebs-Yuill et al. (1995) have estimated (based on water solubility
alone) residual tetrachloroethylene can take up to several hundred years to remediate.
27
This slow desorption is known as hysteresis and is possibly partly responsible for the so
called "rebound effect" which often negatively affects pump and treat remediation
projects (Krebs-Yuill et al., 1995). Just as sorption correlates well to organic matter
content, strong hysteresis usually correlates well with high organic matter contents (Gran
et al., 1994).
Efforts to identify and remediate this bound fraction, however, are problematic, as
sorption tends to protect NOCs from most biological and chemical interactions (Gordon
and Milero, 1985; Sims et al., 1991; Guerin and Boyd, 1992). However, it has been
documented for some time that even very strongly sorbed compounds can desorb
naturally, slowly becoming available for some chemical (Macalady, 1984) and biological
reactions (Khan, 1980a; W eber and Coble, 1968; Robinson et al., 1990; Khan, 1991). Lee
and Kyung (1991) have shown uptake o f aged residues o f carbofuran by rice plants while
Singh et al. (1992) have demonstrated y-chlordane uptake into soybeans, peas, millet,
and several clovers from soils aged up to 17 months. The rates o f release into plants,
while usually very small, are o f concern due to other studies which have shown that
pesticide residues in grain can be biologically available when eaten (Akhtar et al., 1992;
Singh et al., 1993). Pesticide residues sorbed to grain products is analogous to sorption
to SOM and the bio availability o f residues in grain and other agricultural products
suggests that compounds bound to SOM may be bioavailable as well.
Binding to various soil components can also increase mobility through soils if the
compound sorbs to soluble soil organic matter (SSOM) (Scheunert et al., 1992) or to soil
particles small enough to be transported to streams (Leonard, 1990; Squillace and
.28
Thurman, 1992; Liess et a l, 1996). As an contrasting idea, the formation o f bound
residues o f atrazine by a white rot fungi has been proposed as detoxification method if
bound residue formation can be sufficiently enhanced (Hickey et al., 1994). Jean-Marc
Bollag has also proposed incorporation into humic materials as a detoxification scheme
(Dec and Bollag, 1988; Dec et al., 1990; Nannipieri and Bollag, 1991)
Removal and Quantification
The most common technique for quantifying NOCs in soil is hquid/liquid
i
partition. This relies on the theory that like dissolves like. Therefore, hydrophobic organic
compounds sorbed to SOM would redissolve into similar compounds such as solvents.
Choice o f solvent is made on the basis o f chemistry o f the contaminant, polar vs non­
polar. Methanol and aqueous solutions o f methanol are often employed for the s-ttiazines
(Nkedi-Kizza et al., 1987; Miller et a l, 1997b) particularly atrazine, where the single
chlorine substituent creates substantial polarity (Erickson and Lee, 1989). Aqueous
butylamine (40% v/v) showed the best response for solubilizing naphthalene when tested
in series o f co-solvents but has not been tested in soil (Li et a l, 1996). Acetonitrile,
acetone, and chlorinated methanes have also been used for various organic compounds
(Smith, 1981; Huang and Pigaatello, 1990; Conte et a l, 1996; Miller et al., 1997a).
Extraction using super-critical C 0 2 operates on the theory that a super-critical
fluid is neither a gas nor a fluid, but has properties o f both (Rizvi et a l, 1986). At the high
temperatures and pressures involved, the C 0 2 flows like a gas but retains the mass flow
properties o f a liquid. That coupled with the small size o f the molecule made super-
I
29
critical extraction a promising technology. Super-critical C 0 2 has since proven to be as
good as solvent extraction, but often no better depending on the age o f the samples (van
:
der Velde et a l, 1994; Dean, 1996). M ost often, an organic cosolvent, such as methanol,
has to be added to achieve comparable results, 5% M eOH for parathion and 4nitrophenol (Wong et al., 1991), 20 uL for organophosphate pesticides (van der Velde et
al., 1994) and 50 uL for triallate (Bernal et al., 1996). Pure, supercritical methane has
also been used (Capriel et al., 1986). The main advantages o f super-critical technology
appear to be smaller to much smaller sample size, lowered solvent use, and ease of
automation. Disadvantages are initial capital costs o f equipment.
Chemical and biological surfactants have been used to release hydrocarbons from
soil and will be reviewed in more detail below. Extraction using cyclodextrin has also
been attempted, but only on strongly bound residues (Miller et al., Chapter 4 in review).
All o f these techniques are still solvent-type extractions which are known to leave behind
a significant fraction (Dao et al., 1979; Huang and Pignatello, 1990). Prolonged or
repeated extraction procedures are usually effective at removing an additional small
increment o f the sorbed compound from the soil but this increase in efficiency must be
balanced against economic considerations o f additional time and or solvents used by the
procedme (Smith, 1981; Conte, 1996). Eschenbach et al (1994) recently proposed the
use o f methanolic hydrolysis (saponification) for the efficient extraction o f sorbed PAHs.
They postulated that the increased release came from the cleavage o f ester bonds within
the SOM matrix. While quantification o f NOCs in soils is most often accomplished by
solvent extraction under varied reaction conditions, other methods have been proposed.
30
Zabik et al. (1991) have demonstrated that passive sampling devices which mimic
SOM (XAD resins or C18 bonded phase silica) will accumulate organic residues in similar
fashion to SOM. These devices can then be removed for quantitative laboratory analysis.
Lake et al. (1996) have demonstrated similar devices (C-18-coated silica particles) for
quantifying release and bioaccumulation o f chlorinated compounds from aged harbor
sediments, however partitioning o f non-chlorinated polyaromatic compounds into the
particles was less successful even though log Kowvalues between the tw o classes o f
compounds were similar. Biotoxicity tests using lettuce seeds and a soil ciliate species
have also been evaluated for as a semi-quantitative screening device for evaluating bound
residues (Bowers et al., 1997) while prairie grasses have been proposed as bioindicators
o f soils contamination by halogenated compounds (Siciliano et al., 1997), however, this
last study did not use aged soils.
The bound residues which remain in the soil after solvent extraction can be further
characterized by sequential extractions using strong base and strong acid. This extracts
the fulvic acid and humic acid fractions, respectively (Mortensen, 1965; Khan, 1980b).
While the fulvic and humic acid fractions often contain large amounts o f radioactivity,
there is often a significant fraction still left irreversibly bound to the humin fraction (Dec
et al., 1990; Haider et al., 1992; Kastner et al., 1995; Richnow et al., 1995). One novel
approach to the analysis o f bound residues is the derivatization o f organic matter using
silylating reagents, resulting in the removal o f 75% o f bound anilazine (a triazine
derivative) residues aged for 8 weeks (Haider et al., 1992). After standard extraction
using N aO H to remove humic and fulvic acids, 43% remained as genuinely unextractable
31
residues remaining in the humin fraction. Silylation o f this fraction resulted in the release
o f 60% o f that fraction leaving only 17% as unextractable. This technique was also
effective for a selection o f chlorinated phenols and anilines. Sing and Agarwal (1992)
have reported the release o f 81% o f the bound residues o f DDT from soils aged 1 year by
treatment o f the methanol-extracted soil with sulfuric acid. The chemical nature o f the
bound residues were unchanged after aging and acid extraction (75% DDT, 21% DDE
and 4% DDD) indicating that binding and extraction did not alter the structure o f the
compound. It should be noted however, that the chlorine-carbon bond o f DDT is an
extremely chemically resistant bond and sulfuric acid treatment o f this type would no
doubt destroy the identity o f the released residues o f many other families o f pesticides.
Another unique attempt to both release and identify bound residues, developed in
the lab o f S.U. Khan, was a thermal desorption unit which could be ramped up to 800° C
and had the output collected in various traps which could then be cleaned up for GC
identification (Khan and Hamilton, 1980, Khan, 1982b; Ostiz and Khan, 1994). This
approach was quite successful and the temperature o f desorption o f various fragments
allowed for suggestions as to the mechanism or functional groups involved in binding.
This author is surprised that more experiments o f this nature have not been conducted.
Thermal desorption GC-MS has also been employed as a rapid field screening mechanism
for organochlorine pesticides (Robbat et a l, 1992).
32
Aged residue studies
Aged residues are those which most probably resemble bound residues from
historical contamination. Aged residues can be extremely difficult to remove by classic
extraction techniques. Capriel et al. (1985) reported the extraction o f atrazine as the
parent compound from soil 9 years after application. However, this study did not attempt
to degrade the released atrazine or quantify the remaining bound residue. Huang and
Pignatello (1990) have investigated the optimization o f atrazine and metolachlor
extraction using standard organic solvents in samples aged for between 8 and 17 months.
As much as 70% o f the added compound was not removed even by hot solvents after 24
hours. Khan and Ivarson (1982c) tested pure cultures with various metabolic enzymes for
their ability to release residues o f prometryh aged for one year. Their results showed that
cellulolytic and lignolytic bacteria released slightly more o f the bound residues than
lipolytic and proteolytic bacteria (27 vs 24%) but mineralized the released residues less
effectively (1.5 vs 2.5%). Khan and Behki (1990) used an aged, extracted organic muck
soil and incubation with a Pseudomonas culture to release and degrade almost 35% o f
sorbed atrazine residues. However, the soil used was very high in organic matter (45%
carbon) and levels o f sorbed atrazine were substantial (54% o f the applied atrazine after 1
year o f aging).
Guerin and Boyd (1992) studied the bioavailability o f residues o f naphthalene
aged up to 24 days. However, they did not extract the soil before inoculation with 1 o f 2
naphthalene degrading cultures and the sorbed phase was assumed to be protected from
degradation. Despite this assumption, it appears that either the sorbed phase was
33
degraded directly or that desorption occurred faster than their predictive model based on
sorption isotherms alone Then data show greater desorption for a Pseudomonas isolate
vs an unidentified gram negative soil isolate. While some Pseudomonas species have been
shown to produce biological surfactants (Jain et a l, 1992; Van Dyke et a l, 1993a,
1993b), the mechanism for the increased desorption shown in this study remains unclear.
They also develop a
(half saturation constant) value for naphthalene for aged residues
in soil. Aronstein et al. (1991) also showed enhanced degradation o f hydrocarbons
(phenanthrene and biphenyl) in the presence o f non-ionic surfactants without an apparent
concurrent increase in desorption.
It should be stressed however, that in all the above studies, there is a significant
sorbed residue left undisturbed (>50%) regardless o f treatment. This means that a
significant percentage o f the NOC is not being addressed by the desorption techniques
currently in use.
Surfactant studies
The first attempts to use surfactants to enhance the removal o f bound organic
residues in soil were in Germany during the 1980's (Scheunert and Korte, 1985;
Oberbremer and Muller-Hurtig, 1989). Other recent work has shown that for a variety o f
NOCs over a wide range o f aqueous concentrations, the attractive forces o f NOCs
towards soil organic matter and surfactant micelles (including SDS) are approximately
equal (Valsaraj and Thibodeaux, 1989; Edwards et a l, 1991; Jafvert, 1991; Jafvert et a l,
1995). Edwards et al, (1991) plots the log Kow against the log K,,, (a derived micellar
34
partition coefficient) and shows an Excellent correlation over a range o f hydrophobic
solutes. This suggests that NOCs should readily partition from soil organic matter into a
surfactant micelle. Therefore, addition o f various surfactants should enhance removal o f
NOCs from soil.
M uch w ork has been published on the removal o f hydrophobic compounds from
soils by chemical (Rittmann and Johnson, 1989; Liu et a l, 1991; Tsomides, 1995) and
biological surfactants (Aronstein et al., 1991; Jafvert, 1991; Jain et al., 1992; Van Dyke et
a l, 1992, 1993a, 1993b; Aronstein and Alexander, 1992, 1993; Scheiberibogen et a l,
1994) as well as culture filtrates assumed to contain biosurfactants (Jain et a l, 1992; Van
Dyke et al., 1993a). Surfactants have proven to be effective at increasing mobilization o f
NOCs both above and below the CMC in freshly spiked soils (Abdul and Gibson, 1991;
Aronstein and Alexander, 1992; Tsomides, 1995). However, none o f these studies
employed aged soils, so the effects o f enhanced removal will less noticeable in samples
with historic contamination. It has also been noted that micelle structure and size also
play a role in dissolution o f organic compounds. Komives et al. (1994) indicate that for
micelle-encapsulated organophosphoms hydrolase, the micelle must be large enough to
contain not only the enzyme but substantial amounts o f water (presumably containing
substrate) as well. The apparent degradation rate constant was found to be dependent on
the partitioning o f the enzyme between the surfactant layer and the micelle water pool
with greater activity observed with larger water pools inside the micelle.
However, surfactants have also been shown to increase solubility o f some heavy
metals, both in solution and in soils (Tan et a l, 1994; Torrens et a l, (in press)) and
35
mobilization o f sorbed organics may be inadvertent if surfactants are being used to
remediate heavy metal contamination. Conversely, if surfactants are being used to
remediate organic contamination, mobilization o f toxic metals may inhibit biodegradation
o f the organics. In addition, surfactants chosen for the desorption o f organic compounds
destined for bioremediation should be easily biodegradable themselves as well as non­
toxic. Many o f the modem surfactants are, in fact, structured so that microbial
degradation can readily take place. Sodium dodecyl sulfate (SDS) has been shown to
readily biodegraded (Swisher, 1987; Rouse et a l, 1995). Kommalapati arid Roy (1996)
have used a biosurfactant produced from the pericarps o f the fruit o f the soapnut plant,
Sapindus mukorossi. This surfactant was readily biodegradable, showed no toxic effects
on soil microorganisms, and enhanced the removal o f both sorbed naphthalene and
hexachlorobenzene from soil. Rhamnolipids (produced by Pseudomonas bacteria) have
been investigated as natural surfactants for this reason as well. Jain et al. (1992) found
that problems associated with the use o f rhamnolipids in soils included high sorption o f
the rhamnolipid to soil constituents and the high rates required to be effective. Other
researchers have optimized conditions for effective use o f rhamnolipids in soil (Herman,
et al., 1997.
Atrazine
Anionic surfactant (SDS) applied to the surface o f lysimeter plots along with
atrazine did not reduce the formation o f bound residues but did appear to increase
mineralization in the first season (Scheunert and Korte, 1985). There is no available
36
literature on the use o f non-ionic or cationic surfactants for the release o f aged atrazine
residues. Khan and Behki (1990) have incubated aged residues (aged for 1 year) o f
atrazine sorbed onto an organic soil (45% carbon) with a Pseudomonas culture known to
dealkylate atrazine. They showed release and subsequent transformation o f up to 35% o f
the sorbed atrazine. The released atrazine was not mineralized and no attempt was made
to identify any microbially produced surfactants. However, other Pseudomonas species
have been shown to produce rhamnolipid biosurfactants which are effective at mobilizing
NOCs (Van Dyke et a l, 1993a; Scheibenbogen et a l, 1994).
Naphthalene
'
Several studies have used hydrocarbon mixtures including naphthalene or (1 or
2)methyl-naphthalene (Jafvert, 1991; Jain et a l, 1992; Van Dyke et a l, 1993a, 1993b;
Scheibenbogen et a l, 1994). Enhanced removal o f naphthalene resulted from the addition
o f surfactants to freshly spiked soils. However, these treatments often recovered less than
50% o f the added spike and samples were not aged. There appear to be no studies which
have used aged soils, naphthalene residues, and surfactant remediation.
Summary
Residues o f NOCs and other organic compounds bound to soils are bioavailable in
many cases but generally tend to increase in both quantity and recalcitrance with time.
Aged residues represent the most challenging fraction o f bound organic compounds for
remediation. These bound residues can be slowly released and become available not only
37
for biological reactions but for transport to aquifers as well, thus potentially
contaminating groundwater supplies. Soil remediation requires the efficient mobilization
and removal o f these residues and several methods capable o f use at the field scale have
been attempted (co-solvents and surfactants). Many o f the other methods for determining
levels o f bound residues are useful only in the lab or for small samples (thermal
desorption, super-critical C 0 2). Generally, while surfactants o f various kinds have been
shown to be effective for the removal o f fieshly spiked organics, their use oh aged
residues has not been demonstrated.
,
38
CHAPTER 3
PRESENT STUDY
Much o f the work which was completed for this degree is presented in the
publications appended to this dissertation. Following is a summary o f the methods and
significant findings o f that research.
Effect o f Application Solvents
Introduction
During the course o f research into the coupled effects o f nitrogen and organic
contaminants, the application solvent used to apply the organic contaminant to the soil
came under suspicion for inhibiting the transformation o f ammonia to nitrate. Therefore, a
selection o f solvents commonly used in biodegradation experiments were tested for their
effects on the both the heterotrophic and nitrifying communities.
Materials^ and Methods
For each experiment, parallel and identical batch studies were performed due to
the incompatibility between analytical methods for 14C and N. In each study, the carbon
source provided was either unlabeled or 14C-labeled atrazine, naphthalene, or glucose.
Solvent application was made to 5 g o f soil in a 250 ml microcosm and evaporated for a
predetermined length o f time. The nitrogen source provided was applied in all
experiments as (NH4) 2S 0 4 solution. Degradation o f the organic compound was by
39
monitoring 14C 0 2 while nitrogen transformations were monitored by nitrification o f
ammonia to nitrate.
1
i
,
■
Results
All o f the solvents, dichloromethane evaporated for 5 or 60 minutes, chloroform
(trichloromethane), acetonitrile, and methanol significantly inhibited nitrification but did
not inhibit the degradation o f any o f the organics tested. Chlorinated solvents inhibited
nitrification more than methanol or acetonitrile. The inhibition o f nitrification was
reversible with time and destructive binding o f the solvents to the ammonia mono­
oxygenase enzyme was proposed as a potential mechanism.
Conclusion
These results indicate that in soil studies o f volatile organic contaminants such as
naphthalene, the application solvents used to apply the organic compound can effect
different microbial populations in different ways. Both chlorinated solvents, and
dichloromethane in particular, inhibited nitrification in this study even though these
solvents had no effect on degradation o f the carbon source. This inhibition was noticeable
even after the dichloromethane had been evaporated for 1 hour. Although nitrifying and
specific heterotrophic populations were the only groups tested in this study, these results
indicate that specialized groups o f soil microorganisms may be unintentionally impacted
by research techniques. Due to the growing interest in the study o f coupled interactions
o f various microbial populations on processes such as the fate o f fertilizer nitrogen in the
40
presence o f applied pesticides, or the fate o f contaminants in industrial waste streams
containing both carbon and nitrogen species, it is important to understand how these
compounds interact during the degradation o f organic contaminants.
Sorbed Residue Experiments
Introduction
Residues o f organic compounds sorbed to soils and SOM represent a reservoir o f
contamination that constitute a considerable risk to groundwater for the foreseeable
future (Pignatello et a l, 1993). To reduce this risk, it is o f environmental interest to
enhance the release o f these strongly bound residues and to degrade them to innocuous
compounds. Aged and methanol-extracted soils containing 14C-atrazine and 14Cnaphthalene were used to mimic bound residues. Various surfactants were investigated
for their ability to enhance desorption o f bound residues o f atrazine and naphthalene from
2 soils. Surfactants tested were sodium dodecyl sulfate (SDS)(anionic, CMC= 2336
mgZL), Tween 80 (non-ionic, CMC = 13-15 mg/L), and p-cyclodextrin. These were
chosen to represent several classes o f chemistry but which were all still easily degraded
(Swisher, 1987). p-cyclodextrin is a cyclic oligosaccharide produced by the microbial
degradation o f starch and does not form micelles. It is capable o f forming 1:1 inclusion
complexes, with the contaminant partitioned inside the internal cavity o f the cyclic
product. Cyclodextrins do reduce the surface tension o f aqueous solutions (Wang and
Brusseau, 1993).
41
Material and M ethods
x
Strongly bound residues o f atrazine and naphthalene had been generated in
previous experiments and the details on application, aging and .extraction can be found in
Miller et al. (1997c)(Appendix A). Surfactant solutions were prepared so that the
surfactant concentrations delivered ranged over at least 3 orders o f magnitude and
bracketed the critical micelle concentration (CMC). Atrazine desorption was evaluated by
adding 5 g o f contaminated soil to 10 ml o f surfactant solution in 50-ml glass centrifuge
tubes and tumbling for 24 hrs. After centrifugation at 4,000 rp m fo r 30 minutes, duplicate
1 ml samples o f the clear supernatant were quantified using liquid scintillation.
Naphthalene desorption was evaluated in similar fashion except 10 g soil samples and 30
ml o f surfactant solution were used. After desorption was determined, 10 ml o f
supernatant containing released residues was saved for determination o f biodegradation
potential.
Biodegradation potential was evaluated by transferring the 10 ml o f supernatant
into a 15-ml test tube, spiking with 1 ml o f a known mineralizing population and
incubating in a 250-ml microcosm. Tubes fiom both rephcates were placed in the same
microcosm for increased sensitivity o f the assay. A 20-ml scintillation vial containing 1 ml
o f 0.1 M N aOH was also added to each microcosm to trap evolved 14C 0 2.
Results
Tween-80 and (3-cyclodextrin were ineffective at desorbing either compound and
SDS was effective only at desorbing naphthalene. SDS (10,000 mg/L) desorbed 44.0% o f
42
previously nonextractable naphthalene residues vs 13.7% for a water control. However,
SDS inhibited degradation o f released naphthalene residues severely until addition o f a
mineral nutrient solution. Even after nutrient addition, only 6.3 % o f desorbed
naphthalene residues were then degraded to C 0 2 (10,000 mg/L SDS).
Conclusion
The removal (11-19%) o f strongly sorbed residues o f atrazine by water indicates
that desorption from soils which contain sorbed atrazine residues will continue naturally
for months, if not years. Attempts to enhance the removal o f atrazine using surfactants
were not successful. However, SDS enhanced the release o f bound residues o f
naphthalene suggesting that for naphthalene contamination on similar soils, soil washing
using SDS may be an effective remediation strategy. The subsequent biodegradation o f
the released residues also suggests that bioremediation may be an effective means of
efficiently converting released contaminant residues into harmless compounds.
43
Chapter 4
COUPLED INTERACTIONS BETWEEN CONCENTRATION AND SPECIES
OF NITROGEN AND ORGANIC CONTAMINANTS
latroductioii
The discovery o f agrichemicals in groundwater nationwide has prompted research
into the transformations which occur when agricultural chemicals are applied to soils.
Due to the ubiquitous generation o f both nitrogen- and carbon-based contaminants,
increasing interest is being shown in how these classes o f agrichemicals interact with one
,
. • \
another. O f particular interest is the effect o f the addition o f nitrogen on the degradation
o f organic pollutants and the effects o f organic pollutants on the nitrogen cycle. While the
effects o f pesticides on nitrification have been extensively studied, the effect o f nitrogen
on the biodegradation o f organics has received less attention until recently. The effects o f
organic compounds which are not pesticides has also received little attention.
Yee et al. (1985) investigated the release o f bound residues o f prometryne as
affected by nitrogen species. They found that ionic nitrogen formulations (N 0 3" and
NH4+) released greater residues than non-ionic urea. These effects however, were not
apparent after plants (wheat or soybeans) were added to the system and the effect on
degradation o f the organic was not measured. Nitrogen species have also been shown to
have an effect on biodegradation. Lee and Silva (1994) demonstrated enhanced removal
ofPA H s in sedhhent amended with ammonium nitrate or urea. The urea treatment
44
resulted in significant improvements in mineralization o f the PAH compared to the
ammonium nitrate. Hapeman et al (1994) showed that for one o f tw o microbial isolates
i
mineralization o f atrazine was stimulated by the addition o f both ammonia and a carbon
source (com syrup) but not by addition o f either alone. Another isolate showed the
opposite effect: inhibition o f atrazine mineralization when amended with either carbon
alone or carbon plus ammonia while ammonia alone showed no effect. Graves et al
(1994) indicated that degradation o f bis(2-ethylhexyl)phalate was stimulated by addition
o f a degrading-population and/or sewage sludge and nitrogen (ammonia).
Landfarming is a recent remediation technique which relies on stimulating the
native microbial community to degrade low levels o f organic contaminants. This
stimulation is usually achieved by optimizing soil moisture and nutrient status. Nitrogen,
phosphorus and oxygen are the most common additions. Breedveld and Briseid (1994)
have demonstrated a factor o f 4 enhancement o f creosote mineralization using forced air
and nutrient addition (ammonia and phosphorus). Baiba et al (1996) have demonstrated
the effectiveness o f addition o f inorganic nutrients (ammonia and phosphorus) to oil
contaminated sediments in Kuwait. Landfarming vs static piles or windrowing was the
most effective.
The objectives o f the experiments reported here were:
1)
to determine if the addition o f nitrogen (as ammonium sulfate) stimulated the
degradation o f organic pollutants,
2)
,
to determine the effect o f organic contamination on nitrogen transformations.
45
Materials and M ethods
Soils and chemicals. The soil mixture used was a 10:1 mix o f Superstition sand
(sandy, mixed, hyperthermic, Typic Calciorthids) from the Yuma Mesa Agricultural
Center, Yuma, AZ, and Dothan loamy sand (fine loamy, siliceous, thermic, Plinthic
Kandiudult) from the Central Crops Research Station near Clayton, NC, which was
designed to provide a low nitrate background along with a source o f atrazine mineralizers
(Table l)(Miller et a l, 1997). The mixture was prepared by mixing the tw o soils in a
commercial cement mixer for 20 minutes. This resulted in a slight loss o f clay and silt
from the mixed soil (Table 1). All experiments were performed using this soil mix. After
mixing, the air-dried soil was passed through a 2 mm sieve and stored dry in plastic
containers. For experiments, the soil mixture was placed in a plastic bucket and adjusted
to 17% moisture content (w/w) for one week to allow development o f an active microbial
population.
Atrazine and 14C-atrazine (specific activity 14.6 gCi/mg) were obtained from
CIBA Corp, Greensboro, NC. Glucose, 14C-glucose (specific activity 36.0 pCi/mmol),
naphthalene, 14C-naphthalene (specific activity 46.5 pCi/mmol), and all solvents were
obtained from Sigma Chemical Co. (St. Louis, MO). All compounds were uniformly ring
labeled.
Coupled interaction experiments. For all experiments, parallel and identical batch
studies were performed due to the incompatibility between analytical methods for 14C and
15N. A factorial arrangement was used with three rates o f added organic compound and
three rates o f added nitrogen. The biodegradation experiments were carried out in
Table 4-1 Soil characteristics1
Soil
Textural
Soil Texture2
Organic
Class
S and/Silt/Clay
M atter3
-% - —
-%-
—
pH
Vinton
Fine sand
90
7.0
3.0
<0.1
8.4
Dothan
Loamy sand
88
8.1
3.9
1.0
5.8
Superstition
Sand
89
7.9
2.3
0.04
8.81
Mixed Soil
Sand
94
4.8
1.1
0.04
8.58
3Analysis performed by Soil Testing Lab, University o f Arizona.
^Hydrometer method.
3Oven Oxidation method.
47
constructed microcosms (250 rnL glass jars with Teflon lid liners), amended with a total
o f 10 g o f soil (in stages) and adjusted to 17% (w/w) moisture content. Organic
compounds were first added in appropriate solvents to 5 g o f air-dry soil in each
microcosm and the solvent evaporated. Then, 5 g (on a dry wt basis) o f fresh, moist soil
was added to ensure a flourishing microbial population. Atrazine (1, 10, or 100 mg/kg
soil) was applied in 1 ml o f methanol, which was allowed to evaporate overnight.
Naphthalene (1, 10, or 100 mg/kg soil) was applied in dichloromethane evaporated for 8
minutes, which minimized volatilization losses o f naphthalene. Nitrogen compounds were
added in the w ater used to adjust the final moisture content.
fir sample units for quantification o f organic degradation, a 15 mL test tube
containing 10 mL o f 0.1 M N aO H was placed in each microcosm to trap 14C 0 2. At each
sampling time, mineralization was measured by removal o f the traps and quantification o f
1 mL aliqouts in 15 mL o f scintillation cocktail (Scintiverse BD, Fisher, Los Angeles,
CA). Radio assay was by liquid scintillation counting (Tri-Carb 1600, Packard Instrument
Coqp., Meridan, CT). For atrazine, extractable radioactivity was determined by addition
■
o f 50 ml o f methanol to each microcosm and shaking on a rotary shaker (@250 rpm) for
one horn. After filtration through glass fiber filter paper (Watman GFB) and
concentration by rotary evaporation (@ 30° C), 1 ml was radioassayed as described
above. Extractable naphthalene was quantified in a similar fashion except that samples
were not concentrated by rotary eyap oration due to evaporative losses o f the naphthalene
itself. Instead, the specific activity o f the initial application solutions was increased so that
48
samples o f the dilute extraction fluid could be sampled directly. Duplicate samples were
counted to ensure accuracy.
To quantify bound residues, solvent-extracted soil was air-dried and finely
ground. Duplicate 1 g samples were oxidized in a biological oxidizer (Harvey 0X 300) at
875 to 900°C, and the 14C 0 2 trapped in 15 mL o f C 0 2 trapping solution (Harvey
Scintillation Cocktail) and radioassayed.
Results and Discussion
Atrazine
Results. Extractable radioactivity, bound radioactivity and production o f 14C 0 2
are shown in Figures 1, 2 and 3. Extractabihty o f atrazine in this soil mix at time 0 ranged
from 99.9 to 95.2% and averaged 97.2% across all treatments. Extractabihty declined
with time as bound residue formation increased. It was expected that increased rates o f
nitrogen would promote both decreases in extractable radioactivity (due to stimulation o f
microbial activity) and increases in bound residues (due to increases in biomass). This was
not evident from Figure 1. In the 1 mg/kg atrazine treatment, extractabihty was lowest for
I
the 10 mg nitrogen/kg soil samples. This was unexpected and remains unexplained at this
time, hi the 10 mg/kg atrazine treatment, there is a distinct drop in extractabihty for the
100 mg nitrogen/kg soil samples vs the 1 and 10 mg/kg soil treatments and samples from
e
f
ah three nitrogen treatments in the 100 mg/kg atrazine treatment show reduced
extractabihty as weh.
Bound residue formation steadily increased for the duration o f the experiment,
100
80
60
------- N = 1 ppm
----- N = 10 ppm
......... N = 100 ppm
40
20 H
% of the A pplied R adioactivity
A trazine = 1 mg/kg
0 -
100
-
80 —
60 —
40 -
20
-
0
-
100
-
— N = 1 ppm
N = 10 ppm
- N = 100 ppm
A trazin e = 10 mg/kg
80 60 —
40 -
20
—
N = 1 ppm
N = 10 ppm
..... N = 100 ppm
-
0 -
A trazin e = 100 m g/kg soil
0
10
20
Tim e (weeks)
30
Figure 1
Extractable Radioactivity, Atrazine
40
A trazine = 1 mg/kg
------- N = 1 ppm
-------N = 10 ppm
- N = 100 ppm
60 —
% of the A pplied Radioactivity
20
-
A trazine = 10 mg/kg
------- N = 1 ppm
N = 10 ppm
N = 100 ppm
60
40
20
-
A trazin e = 100 m g/kg soil
------- N = 1 ppm
------- N —10 ppm
60 - ......... N = 100 ppm
20
-
Tim e (weeks)
Figure 2
Bound Residues of Atrazine
51
= 1 ppm
---- N = 10 ppm
..... N = 100 ppm
A trazine = 1 m g/kg
IN = 1 ppm
N = 10 ppm
N = 100 ppm
A trazine = 10 m g/kg
N = 1 ppm
-------N = 10 ppm
— N = 100 ppm
A trazine = 100 m g/kg soil
% of the A pplied Radioactivity
n
—
Tim e (days)
Figure 3
Production of l4C01from Atrazine
52
peaking at 35% after 34 weeks (1 mg/kg treatment) (Figure 2). Again, the highest level
o f bound residues was in the 10 mg nitrogen/kg soil samples. The same pattern o f
differences between the 100 mg nitrogen samples from the 10 mg/kg atrazine treatment
and aft 3 nitrogen treatments from the 100 mg/kg atrazine samples was noticed for bound
residue formation as well.
After 38 weeks, 11% o f the added atrazine in the 1 mg kg"1 treatment had been
recovered as 14C 0 2, 9% in the 10 mg kg"1treatment and 4% in the 100 mg kg"1 treatment
(Figure 3). The differences within treatments (esp. the 10 mg/kg atrazine treatment) and
between treatments (100 mg/kg atrazine vs 1.0 and 10 mg/kg) were particularly
noticeable in Figure 3.
Discussion. The differences noted within and between treatments can be
explained by the replacement o f those samples in the nitrogen side o f the experiment.
Application rates o f nitrogen had been miscalculated and these treatment combinations
were replaced about 3 weeks into the study. To maintain the coupled nature o f the
experiments, the same treatment combinations in the atrazine study were also replaced.
However, the results o f these replacements do not seem to be closely correlated with the
results o f the rest o f the study. The reason for this discrepancy is unknown; however, it
may be that it is related to storage o f the soil during the three week interim.
The atrazine mineralization potential o f this soil mix was never fully realized. This
may be a result o f dilution o f the original soil inoculum (10:1 Yuma:Dothan), or a result
o f cold and/or dry storage. Cold, dry storage has been shown to reduce atrazine
mineralization (Baniuso and Houot, 1996). The failure to detect any effects o f the
53
addition o f nitrogen may be related to the failure o f the nitrifying bacteria to convert the
added ammonium to nitrate. This latter possibility was not suspected at the time and is
discussed in more detail in the conclusions. There appeared to be no effect o f the
application solvent on the degradation o f atrazine. The similarity o f degradation rates
across atrazine concentration indicates that a small but functioning mineralizing
population was present in all samples (Figure 3). This supports the possibility that dilution
may have been responsible for the lack o f mineralization although storage may have had
an effect as well.
Naphthalene
'
Results. Totalnaphthalene recoveries ranged from 50.2 to 113.2% with an
average o f 79.6%. Extractable radioactivity declined rapidly for the first 15 days and then
stabilized at a very low level (Figure 4). Bound residues formed immediately and ranged
from less than 10% (100 pg naphthalene) to almost 40% (1 pg naphthalene) at time 0
(Figure 5). With the exception o f the 100 pg naphthalene treatment, bound residues
remained relatively constant between 25 and 35%. Naphthalene was readily mineralized,
with 13-17% o f the added radioactivity recovered as 14C 0 2 after 8 days and an average
across all treatments o f 59.2% after 49 days (Figure 6).
Discussion. While there was little to no effect o f nitrogen rate on extractability or
bound residue formation (except for the 100 kg/kg nitrogen treatment in the 1 mg/kg
atrazine samples at 45 days), there appeared to be a minor effect o f naphthalene rate on
production o f 14C 0 2. The 100 mg/kg nitrogen samples consistently produced greater
54
N aphthalene = 1 m g/kg
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
20
-
% of the Applied Radioactivity
N aphthalene = 10 m g/kg
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
20
60 -
N aphthalene = 100 m g/kg soil
------- N = 1 ppm
-------N = 10 ppm
-------N —100 ppm
40
Time (Days)
Figure 4
Extractable Radioactivity, Naphthalene
55
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
50 -
% of the Applied Radioactivity
30 -==
N aphthalene = 1 m g/kg
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
50 40
30
20 -t
Nc
10
N aphthalene = 10 m g/kg
0
Figure 5
Bound Radioactivity, Naphthalene
56
N aphthalene = 1 mg/kg
% of the Applied Radioacti
40 -
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
N aphthalene = 10 mg/kg
60
40 -
20
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
N aphthalene = 100 mg/kg soil
60
40 -
20
------- N = 1 ppm
-------N = 10 ppm
-------N = 100 ppm
Time (days)
Figure 6
Production of l4CO, from Naphthalene
57
amounts o f 14C 0 2 than the 1.0 or 10 mg/kg nitrogen treatments; however there was
enough variability within the data that this effect was probably not significant.
It was during the course o f the naphthalene degradation experiments that the
effects o f the application solvent on the nitrogen side o f the experiment became apparent.
This discovery prompted the research which is presented in Appendix A. However, it
remains a distinct possibility that the failure in these experiments to see any interaction
between nitrogen addition and enhanced organic degradation was a result o f the
application solvent effect rather than an actual null response. The use o f methanol for
addition o f naphthalene had no effect on the degradation o f naphthalene. Further
experiments with a variety o f solvents also showed no effect on naphthalene degradation
(Appendix A). It is unclear at this time whether the lack o f a response to nitrogen is an
artifact o f the application technique.
Conclusion. Indeed, it may be possible that due to this artifact, the entire time
that experiments were being performed on the interactions between nitrogen and organic
compounds, we may not have actually been testing one o f our two main hypotheses, ie.
that addition o f nitrogen will stimulate biodegradation. Ammonia has been shown to be
toxic to soil microorganisms at higher rates (1.0%) and also to inhibit atrazine
mineralization (Hapeman et a l, 1994). While no effect o f the addition o f nitrogen on the
biodegradation o f either naphthalene or atrazine was ever documented in these studies,
further studies in this area are still warranted in this author’s opinion.
APPENDIX A
The Effect o f Application Solvents on Nitrifying Bacteria
Published in Environmental Toxicology and Chemistry, 16:447-451, 1997.
IS S N 0730-7268
V o lu m e 16, N u m b e r 2. February 1997
ENVIRONMENTAL
TOXICOLOGY AND
CHEMISTRY
An International Journal
A M onthly
Publication of the
Society of
Environmental
Toxicology and
Chemistry
SETACip r e s si
SETAC
60
Environmental Toxicology and Chemistry, Vol. 16, No. 3, pp. 447-451, 1997
© 1997 SETAC
Printed in the USA
0730-7268/97 S6.00 + .00
EFFECT OF APPLICATION SOLVENTS ON HETEROTROPHIC AND NITRIFYING
POPULATIONS IN SOIL MICROCOSMS
J e r r y L. M i l l e r , M i c h a e l A. S a r d o , T h o m a s L. T h o m p s o n and R a i n a M. M i l l e r *
Department of Soil, Water, and Environmental Science, University of Arizona, Tucson, Arizona 85721, USA
{Received 31 January 1996; Accepted 30 July 1996)
A bstract—Agricultural practices may cause contamination of soil and ground water with a combination of organic compounds
(pesticides and fuel) and nitrogen fertilizers. In coupled microcosm studies that monitored the mineralization of naphthalene and
the nitrification of ammonia, it was observed that the solvent (dichloromethane) used to apply naphthalene to the soil inhibited
nitrification, although there was no effect on naphthalene mineralization. Further studies were performed with a series of application
solvents: methanol, acetonitrile, trichloromethane, and dichloromethane. Soil and solvent were allowed to equilibrate with ambient
air for various times before capping and incubation of microcosms. Results indicated that dichloromethane equilibrated for 5 mins
inhibited nitrification for at least 3 weeks relative to the control (water). Acetonitrile and trichloromethane similarly inhibited
nitrification. Methanol and dichloromethane equilibrated for 60 mins also significantly delayed nitrification, although to a lesser
extent. Inhibition of nitrification was not permanent, and nitrification activity was eventually restored in all systems tested. None
of the solvents inhibited mineralization of the added carbon source. These results indicate that special care must be taken to ensure
that applications solvents do not affect the activity of sensitive microbial populations, such as the nitrifiers, that may be part of a
study.
Keywords—Nitrification
Chlorinated solvents
Organic compounds
Biodegradation
organic mineralization. A series of parallel soil microcosm
experiments were performed to determine nitrification of
NHJ and mineralization of atrazine, naphthalene, and glucose.
In addition, the effects of a variety of application solvents on
nitrification and mineralization were determined. Application
solvents tested included acetonitrile, chloroform, dichloro­
methane, methanol, and water.
INTRODUCTION
. Soil biodegradation studies of organic contaminants with
low water solubility usually involve application of the organic
compound in an organic solvent that is then allowed to equil­
ibrate with ambient air to allow solvent evaporation. Generally,
application of the organic/solvent mixture is made to a portion
of the soil, and following solvent evaporation, an additional
portion of fresh soil is added to ensure a flourishing microbial
community. This has proven to be an effective technique for
study of the degradation of organic chemicals by heterotrophic
communities. Various application solvents have been used,
including methanol [1,2], acetone [3], and acetonitrile [4], and
extensive use has been made of the chlorinated methanes [57]. The choice of application solvent depends on the aqueous
solubility and volatility of the organic compound being stud­
ied. For example, naphthalene is a relatively volatile com­
pound. Therefore, it must be applied in a solvent of even
greater volatility, usually a chlorinated solvent such as di­
chloromethane or chloroform.
Research techniques such as those just described generally
focus on measurement of just one type of activity in the en­
vironment, usually heterotrophic activity. However, contami­
nants are generally found as mixtures of many, different com­
ponents, including both carbon and nitrogen species. There­
fore, it is of interest to know how contaminants affect different
microbial populations that may be active in their transfor­
mation. Thus, the objective of this study was to investigate
the coupled interactions between nitrogen, applied as
(NH4)2S 04, transformation, as measured by nitrification activ­
ity, and organic contaminant transformation, as measured by
M ATERIALS AND M ETHODS
Soils and chemicals
The soil used in this study was a 9:1 mixture of two soil
types, a Superstition sand (sandy, mixed, hyperthermic, Typic
Calciorthids) from the Yuma Mesa Agricultural Center, Yuma,
Arizona, USA, and a Dothan loamy sand (fine loamy, siliceous,
thermic, Plinthic Kandiudult) from the Central Crops Research
Station near Clayton, North Carolina, USA. Characteristics of
these soils and the mixture are given in Table 1. The air-dried,
mixed soil was passed through a 2-mm sieve and stored in
plastic containers. Prior to each experiment the soil mixture
was placed in a plastic bucket and adjusted to 17% moisture
content (w/w) to allow development of an active microbial
population. The mixed soil was designed to provide both a
low nitrate background and a source of atrazine degraders.
Atrazine and [,4C]atrazine (specific activity 14.6 fiCi/mg)
were obtained from CIBA Corporation (Greensboro, NC,
USA). Glucose, [14C]glucose (specific activity 36.0 pGi/
mmol), naphthalene, [14C]naphthalene (specific activity 46.5
IxCi/mmol), and all solvents were obtained from Sigma Chem­
ical Company (St. Louis, MO, USA). All compounds were
uniformly ring labeled.
Biodegradation and nitrogen transformation experiments
For each experiment, parallel and identical microcosm stud­
ies were performed due to the incompatibility between ana-
* To whom correspondence may be addressed.
447
i
61
448
Environ. Toxicol. Chem. 16, 1997
J.L. Miller et al.
Table 1. Soil characteristics1'
Soil .
Textural class
Dothan
Superstition
Mixed
Loamy sand
Sand
Sand
Soil texture15 (%) Organic
matter0
Sand Silt Clay
(%)
88/
89
94
8.1
7.9
4.8
3.9
2.3
1.1
1.0
0.04
0.04
pH
5.80
8.81
8.58
Arizona, Tucson, Arizona, USA.
b Determined by hydrometer method.
c Determined by oven oxidation method.
lytical methods for l4C and nitrogen. In initial studies, atrazine
(1, 10, or 100 mg/kg soil, approx.. 2.5 X 105 dpm/microcosm)
was applied in 1 ml of methanol to 5 g of air-dried soil pre­
viously placed in each microcosm (250-ml glass jars with Tef­
lon® lid liners). The methanol was allowed to evaporate over­
night, then 5 g (dry weight) of fresh, moist soil was added,
the moisture content was adjusted to 17% (w/w), and the mi­
crocosms were sealed. Naphthalene (1, 10, of 100 mg/kg soil,
approx. 2.5 X 105 dpm/microcosm) was applied in dichloromethane to 5 g of air-dried soil in each microcosm. The dichloromethane was allowed to evaporate for 7 min to maximize
removal of the application solvent and minimize volatilization
of naphthalene. Five grams of fresh soil were added, moisture
content was adjusted, and microcosms were sealed. Time zero
recoveries of naphthalene were 64% (applied at a rate of 1
mg/kg soil), 83% (100 mg/kg), and 93% (100 mg/kg).
In subsequent experiments, to test the effect of application
solvents- on nitrification, either naphthalene or glucose was
used as a carbon source. In these experiments, solvent appli­
cation was again made to 5 g of air-dried soil in a microcosm.
Solvent volumes used for application (Table 2) were based on
representative values found in the literature. Microcosms were
allowed to equilibrate with ambient air for a predetermined
length of time, shown in Table 2. Following equilibration, 5
g (dry weight) of moist soil was added, and the appropriate
carbon and nitrogen sources were added in aqueous solution
to bring the final water content to 20% (w/w). Naphthalene
(3.3 mg/kg soil, approx. 2.5 X 10 dpm/microcosm) was applied
in 1.0 ml of a saturated aqueous solution (33 mg/L), while
Table 2. Solvent characteristics
Solvent
Acetonitrile
Methanol
CHCh
CH2Ci2
CH2C1,
Vapor
pressure3
(mm Hg)
78.2
95.5
395.5
1,891.0
1,891.0
Volume
tested
(ml)
. 1.0
1.0
0.1
0.1
0.1
Evapo­
ration
time6
(min)
60
50
30
5
60
Residual
. solvent0
(mg/kg)
ND
ND
ND
19,300
1,800
a [19].
b Evaporation time for acetonitrile, methanol, and chloroform was
defined as double the length of time required so that no visual or
olfactory trace of the solvent remained in the test microcosm. Two
evaporation times were tested for dichloromethane: 5 min, long
enough so that no visual or olfactory traces of solvent remained,
and 60 min, a 12-fold increase.
c Residual solvent was determined by thermal desorption gas chro­
matography as described in “Materials and Methods.”
ND -= not determined.
glucose (100 mg/kg soil, approx. 2.5 X 10 dpm/microcosm)
was applied in 0.5 ml of water.
The residual solvent left in the soil after equilibration was
quantified for dichloromethane using thermal desorption and
gas chromatography. Dichloromethane was added to 5-g soil
samples in microcosms as described above. After equilibration
with ambient air for 5 or 60 min, the soil was placed in crimpcapped vials. Dichloromethane standards were prepared in 5
ml of water and sealed in crimp-capped vials. Samples were
analyzed using a Techmar 7000 headspace autosampler (Cin­
cinnati, OH, USA) to thermally desorb the solvent, followed
by gas chromatography using a Shimadzu model GC-17A gas
chromatograph. The nitrogen source provided was applied in all experi­
ments as (NHJ2SO4 solution to give final soil concentrations
of 0, 10, 40, and 100 mg/kg soil..
In microcosms for determination of organic degradation, a
15-ml test tube containing 10 ml of 0.1 M NaOH was placed
in each microcosm to trap ,4C 02. At each sampling time the
traps were removed, and 1 ml was added to 15 ml of scintil­
lation mixture (Scintiverse BD, Fisher, Los Angeles, CA, USA)
and assayed for radioactivity on a liquid scintillation counter
(Tri-Carb 1600, Packard Instrument Corp., Meridan, CT,
USA).
In samples for the quantitative determination of nitrifica­
tion, NH^-N and NOj-N plus NO7-N concentrations of each
soil sample were determined by extraction with 1.0 M KC1
and by steam distillation with MgO and Devarda alloy as
described by Keeney and Nelson [8].
Reversibility o f nitrification inhibition
Some samples that clearly showed inhibition of nitrification
were incubated for several additional months. These samples
were then tested for the presence of nitrification activity using
the following procedure. One gram of soil was added to a test
tube containing 10 ml of HaO, vortexed for 1 min, and allowed
to settle. A drop of the supernatant was applied to a colori­
metric nitrate test strip that is semiquantitative up to 250 mg/L
for nitrite and 500 mg/L for nitrate (Merckoquant Test Strips,
Merck, Darmstadt, Germany). In addition, 1 ml of supernatant
was added to 25 ml of a nitrification enrichment medium. The
medium was composed of (per liter of medium) 1.0 g KH2P 0 4,
1.0 g Na2H P04, 0.5 g NH4N 0 3, 0.5 g (NH4)2 S 04, 0.2 g
. M gS04-7 H20 , 0.02 g CaCl2 2 H20 , 0.002 g FeCl3, and 0.002
g MnSG4- 2 H20 , at pH 7.2. Flasks were incubated on a rotary
shaker at 200 rpm for 2 months and tested periodically for the
presence of nitrite and nitrate to indicate nitrification activity.
RESULTS
Results showed that the amount of nitrogen present did not
affect biodegradation of any of the carbon sources tested, in­
cluding atrazine or naphthalene (Fig. 1), at any concentration
tested from 1 to 100 mg/kg soil. Furthermore, the presence of
atrazine did not inhibit nitrification (Fig. 2A). However, naph­
thalene at 1 and 10 mg/kg appeared to completely inhibit ni­
trification during the 35-d study (Fig. 2B), while naphthalene
at 1Q0 mg/kg delayed nitrification until at least day 20.
Subsequent studies were performed to further investigate
the effect of naphthalene on nitrification. In these studies the
effect of the naphthalene application solvent, dichloromethane,
was also considered. As shown in Figure 3A, naphthalene
added in water (final concentration 3.3 mg/kg soil) slowed
nitrification in comparison to the control for at least 25 d.
62
Effect of application solvents on nitrification
A
0
-o
E n v ir o n . T o x ic o l. C h e m .
16, 1997
449
Atrazine = 1 mg/kg
10
20
30
40
Time (weeks)
Time (weeks)
>
B
Naphthalene = 1 mg/kg
Time (days)
Fig. 1. The effect of N H ;-N at three rates on the mineralization of
atrazine (A ) and naphthalene (B ) in soil. Presented are the mean values
of triplicate experiments together with their standard errors. O = 1
mg/kg N H ;-N ; □ = 10 mg/kg NH4‘ -N ; A = 100 mg/kg N H ;-N .
However, by day 35, over 90% of the added ammonium had
been converted to nitrate in both the control and naphthalene
experiments. In contrast, microcosms treated with dichloromethane alone showed complete inhibition of nitrification dur­
ing the entire 35-d study. Figure 3B shows the effect of dichloromethane on mineralization of naphthalene. As shown in
this figure, the extent of mineralization of naphthalene applied
in dichloromelhane at a concentration of 100 mg/kg was sim­
ilar to the extent of mineralization of naphthalene applied at
a concentration of 3.3 mg/kg in water, indicating no effect of
the dichloromelhane.
These results suggested that application solvents may ad­
versely effect nitrifying populations in soil. Therefore, further
experiments were conducted with a variety of application sol­
vents to determine whether this method of contaminant ap­
plication could selectively inhibit nitrification. These experi­
ments were performed with glucose, a readily degradable car­
bon source. All application solvents were evaporated for the
length of time shown in Table 2. When water was the appli­
cation solvent, nitrification was complete within approx. 15 d
(Fig. 4A). All other application solvents inhibited nitrification
to some extent, although once nitrification started the rate of
nitrification was similar in all cases. For example, the onset
of nitrification was delayed until at least 19 d by dichloro­
melhane (evaporated for 5 min), acetonitrile, and chloroform.
Nitrification was also delayed to a lesser extent by dichloro­
melhane (evaporated for 60 min) and methanol. Even though
delayed, nitrification eventually was complete by day 27 in all
Time (days)
Fig. 2. The effect of atrazine (A) and naphthalene (B) on the oxidation
of N H J-N (100 mg N H ^-N /kg soiljto NO, -N . Presented are the
mean values of triplicate experiments together with their standard
errors O = 1 mg/kg atrazine or naphthalene: □ = 10 mg/kg: A =
100 mg/kg.
cases except dichloromelhane (evaporated for 5 min), in which
nitrification was still strongly inhibited beyond 27 d. As shown
in Figure 4B. glucose mineralization was not affected by any
solvent.
In all experiments, the loss of ammonium was closely par­
alleled by the production of nitrate (data not shown). It was
therefore surprising that the total recovery of nitrate was great­
er than the added ammonia for the acetonitrile treatment (Fig.
4A). Because acetonitrile also contains nitrogen, it is assumed
that the excess nitrate formed was due to conversion of ace­
tonitrile nitrogen to nitrate. O’Grady and Pembroke [9] have
shown ammonium production from acetonitrile in pure culture
with an Agrobacterium spp. Presumably, this ammonium
would then be available for conversion to nitrate in soil sys­
tems.
Three microcosms that were treated with dichloromelhane
(and naphthalene at 100 mg/kg) and showed strong inhibition
of nitrification even after 35 d were saved and incubated for
an additional 2 months. Nitrate test strips detected nitrite and
a small amount of nitrate (<50 mg/L) after 1 month. After the
second month, nitrate was strongly detected (>250 mg/L) in
all three samples. Enrichment cultures for nitrifying bacteria
inoculated from these samples also showed detectable nitri­
fication activity (>250 mg/L) within 2 months.
DISCUSSION
This study showed no effect of any of the application sol­
vents tested on the degradation of the carbon substrates. Allen-
63
Environ. Toxicol. Chem. 16. 1997
J.L. Miller et al.
A
0
5
10
15
20
25
30
35
0
50 mg nitrogen/kg soil)
5
Time (days)
5
10
15
20
25
15
20
25
30
Time (days)
B
0
10
30
35
Time (days)
Fig. 3. The effect of aqueous naphthalene and dichloromethane on
the oxidation of N H f-N to N O ,'-N (A) and the effect of dichloro­
methane on the mineralization of naphthalene (B). Applied dichlo­
romethane was evaporated for 7 min. Presented are the mean values
of triplicate experiments together with their standard errors. (A) O =
water (control); □ = naphthalene (3.3 mg/kg soil) applied in aqueous
solution; A = CHCI,. (B) O = 3.3 mg/kg naphthalene applied to soil
in aqueous solution (control); □ = 100 mg/kg naphthalene applied
to soil in dichloromethane.
King et al. [10] have also shown no effect of a chlorinated
solvent on degradation of an organic substrate. In their study,
trichloromethane did not inhibit toluene degradation at trichloromethane concentrations up to 4 mg/L. However, in a study
by Narayanan et al. [11], trichloromethane inhibited the an­
aerobic degradation of acetate and acetone at concentrations
of 20 mg/L in a granulated activated carbon reactor. The re­
sidual concentration of dichloromethane left in the soil after
equilibration in this study was found to be 19,300 mg/kg after
5 min and 1,800 mg/kg after 60 min. Interestingly, the amount
of residual solvent (although most likely this is in the sorbed
phase) is one to two orders of magnitude higher than the
amount of contaminant applied (100 mg/kg).
Application solvents tested in this study can be divided into
three categories: those with little effect on nitrification, those
with a moderate effect, and those that delayed nitrification for
2 weeks or longer. The strongest inhibition occurred with the
chlorinated solvents, dichloromethane and chloroform. As
shown in Figure 4A, the effect of dichloromethane was tested
with two evaporation times. Even when dichloromethane was
allowed to evaporate for almost 12 times longer than might
seem necessary empirically (60 min), significant inhibition of
nitrification was observed.
Nitrification is primarily a result of activity of the nitrifying
0
100 mg glucose/kg soil)
4
8
12
16
20
Time (days)
Fig. 4. The effect of various application solvents on the oxidation of
N H f-N (applied at 50 mg/kg soil) to N O ,-N (A) and of glucose
(applied at 100 mg/kg soil) on mineralization (B). Presented are the
mean values of triplicate experiments together with their standard
errors. O = methanol; □ = acetonitrile; A = CHCL,; V = CH.CL,,
evaporated for 5 min; 0 = CH.CL,, evaporated for 60 min; O water (control).
bacteria Nitrosomonas and Nitrobacter spp. Nitrification also
can be carried out by several heterotrophic genera of bacteria
and fungi, but the contribution probably is small in soils with
an active nitrifying population of Nitrosomonas and Nitro­
bacter spp. [12]. The chemoautotrophic nitrifiers are generally
regarded as being relatively sensitive to organic compounds,
and at least one such compound (nitripyrin) is used commer­
cially to inhibit nitrification in the soil [13]). However, many
other organic contaminants, including many herbicides, have
no effect on nitrification [14].
Rasche et al. [15] demonstrated that ammonia oxidation to
nitrite by Nitrosomonas europea in pure culture is inhibited
by several chlorinated solvents due to an interaction with the
ammonia monooxygenase enzyme (AMO). It was found that
if sufficient time was allowed to elapse for de novo protein
synthesis, nitrification activity could be reestablished. In the
current study, the detection of both nitrite and nitrate in sam­
ples after 2 months of additional incubation and the growth
of nitrifying bacteria in liquid culture medium indicate that
inhibition of nitrification may similarly be due to this inter­
action with a key enzyme. Keener and Arp [16] have shown
that the halogenated alkanes are noncompetitive substrates for
AMO and have proposed a model of AMO that contains two
substrate binding sites. In this model, one site will accom­
modate C, and C2 hydrocarbons as well as NH,. while the
other, more hydrophobic site has a wider substrate range that
64
Effect of application solvents on nitrification
includes the halogenated methanes and ethanes. Other recent
work has suggested that AMO inhibition can be both specific
and nonspecific [17]. Specific AMO inhibition occurs when
the AMO enzyme is irreversibly bound by a chlorinated or­
ganic, while nonspecific inhibition is the result of AMO ox­
idation of a chlorinated organic that leads to the formation of
intermediates that are toxic to cell constituents. The time re­
quired for recovery of nitrification activity is shorter for spe­
cific inhibition of the AMO because only one enzyme needs
to be synthesized. The results of this study are consistent with
a nonspecific inhibition of nitrification caused by the solvents
tested.
These results suggest that soil studies of volatile organic
contaminants such as naphthalene applied in chlorinated sol­
vents may be affecting different microbial populations in dif­
ferent ways. Both of the chlorinated solvents, dichloromethane
in particular, inhibited nitrification in this study even though
they had no effect on degradation of the carbon source. Ni­
trification inhibition was evident even after the dichlorome­
thane had been allowed to evaporate for 60 min. This may
have been due to sorption of the dichloromethane by soil [18]
and a subsequent slow release of dichloromethane residues
from the soil. Although nitrifying and specific degrading heterotrophic populations were the only groups tested in this
study, these results indicate that specialized groups of soil
microorganisms may be unintentionally impacted by research
techniques. This is of concern because of the growing interest
in the study of coupled interactions of various microbial pop­
ulations on processes such as the fate of fertilizer nitrogen in
the presence of applied pesticides or the fate of contaminants
in industrial waste streams containing both carbon and nitrogen
species.
Environ. Toxicol. Chem. 16, 1997
4.
5.
6.
7.
8.
9.
10.
11.
12.
13.
14.
15.
16.
Acknowledgement—This work was supported by the Cooperative
State Research Service, U.S. Department of Agriculture, under agree­
ment 92-34214-7433. The authors are grateful to CIBA Corporation
for the generous gift of atrazine, both radioactive and nonradioactive.
REFERENCES
1. Zhongbao, L., S. Laha and R.G. Luthy. 1991. Surfactant sol­
ubilization of polycyclic aromatic hydrocarbons in soil-water sus­
pensions. Water Sci. Technol. 23:475-485.
2. Miller, J.L. 1991. Degradation of atrazine and metolachlor in
soils from four depths in a Dothan loamy sand. M.S. thesis. North
Carolina Slate University, Raleigh, NC, USA.
3. Guerin, W.F. and S.A. Boyd. 1992. Differential availability of
17.
18.
19.
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soil-sorbed naphthalene to two bacterial species. Appl. Environ.
Microbiol. 58:1142-1152.
H annan, P.J. 1995. A novel detection scheme for herbicide res­
idues. Environ. Toxicol. Chem. 14:775-780.
M iller, R.M., G.M. Singer, J.D. Rosen and R. Bartha. 1988.
Photolysis primes biodegradation of benzo [a] pyrene. Appl. En­
viron. Microbiol. 54:1724-1730.
Churchill, S.A., R.A. Griffin, L.P. Jones and P.F. Churchill.
1995. Biodegradation rate enhancement of hydrocarbons by an
oleophilic fertilizer an6 a rhamnolipid biosurfactant. J. Environ.
Qual. 24:19-28.
Tsomides, H .J., J.B. Hughes, J.M . Thomas and C.H. W ard.
1995. Effect of surfactant addition on phenanthrene biodegra­
dation in sediments. Environ. Toxicol. Chem. 14:953-959.
Keeney, D.R. and D.W. Nelson. 1982. Nitrogen-inorganic forms.
In A.L. Page, R.H. Miller and D.R. Keeney, eds., Methods o f
Soil Analysis, 2nd ed., Vol. 2. American Society of Agronomy,
Madison, WI, USA, pp. 643-698.
O ’G rady, D. and J.T. Pembroke. 1994. Isolation of a novel
Agrobacterium spp. capable of degrading a range of nitrile com­
pounds. Biotechnol. Lett. .16:47-50.
Alien-King, R.M., R.W. Gillham and J.F. Barker. 1996. Fate
of dissolved toluene during steady infiltration through unsaturated
soil: I. Method emphasizing chloroform as a volatile, sorptive
and recalcitrant tracer. J. Environ. Qual. 25:279-286.
N arayanan, B., M.T. Suidan, A.B. Gelderloos and R.C. B ren­
ner. 1993. Treatment of VOCs in high strength wastes using an
anaerobic expanded-bed GAC reactor. Water Res. 27:181-194.
Atlas, R.M. and R. B artha. 1993. Microbial Ecology, 3rd ed.
Benjamin/Cummings, New York, NY, USA.
C raw ford, D M. and P.M. Chalk. 1992. Mineralization and im­
mobilization of soil and fertilizer nitrogen with nitrification in­
hibitors and solvents. Soil Biol. Biochem. 24:559-568.
Domsch, K.H. and W. Paul. 1974. Simulation and experimental
analysis of the influence of herbicides on soil nitrification. Arch.
Microbiol. 97:283-301.
Rasche M E., M R. Hyman and D.J. A rp. 1991. Factors limiting
chlorocarbon degradation by Nitrosomonas europaea: Cometabolic inactivatibn of ammonia monooxygenase and substrate
specificity. Appl. Environ. Microbiol. 57:2986-2994.
Keener, W.K. and D.J. A rp. 1993. Kinetic studies of ammonia
monooxygenase inhibition in Nitrosomonas europaea by hydro­
carbons and halogenated hydrocarbons in an optimized whole­
cell assay. A ppl Environ. Microbiol. 59:2501-2510.
Juliette, L.Y., M.R. Hyman and D.J. Arp. 1993. Mechanismbased inactivation of ammonia monooxygenase in Nitrosomonas
europaea by allysulfide. Appl. Environ. Microbiol. 59:37283755.
Shonnard, D.R., R.L. Bell and A.P. Jackm an. 1993. Effects of
non-linear sorption on the diffusion of benzene and dichloro­
methane from two air-dry soils. Environ. Sci. Technol. 27:457465.
Lide, D.R. 1991. CRC Handbook o f Chemistry and Physics, 72nd
ed., CRC Press, Boca Raton, FL, USA.
APPENDIX B
Desorption and biodegradation o f strongly bound soil residues o f atrazine and
naphthalene using various surfactants
Jerry L. Miller
Thomas L. Thompson
M ark L. Brusseau
Raina M. Miller*
Department o f Soil, W ater and Environmental Science
University o f Arizona
Tucson, AZ 85721
C orresponding author:
. Raina M. Miller
University o f Arizona, Tucson, AZ
Phone
(520)621-7231
FAX
(520) 621-1647
E-mail
[email protected]
A manuscript for submission to The Journal o f Environmental Quality
66
ABSTRACT
Nonextractable residues o f organic pollutants bound to soils form a potential
reservoir o f future environmental contamination. Aged and methanol-extracted soils
containing 14C-atrazine and 14C-naphtha!ene were used to mimic these bound residues.
Three surfactants (anionic, cationic, and a cyclodextrin) were investigated for their ability
to enhance desorption o f the bound residues o f atrazine and naphthalene. SDS (10,000
mg/L) desorbed 44-49% o f previously nonextractable naphthalene residues vs 10-14%
for a w ater control. After 5 washes, 66% o f bound naphthalene residues were removed
(10,000 mg/L). To evaluate the biodegradation potential o f the released residues,
surfactant samples containing desorbed naphthalene were inoculated with a known
naphthalene mineralizer and evolution o f 14C 0 2 monitored. SDS inhibited degradation o f
naphthalene with only 6-28% o f desorbed naphthalene residues mineralized to 14C 0 2 vs
38% for a w ater control. Tween-80 and P-cyclodextrin were ineffective at desorbing
either compound. This may be due to size exclusion o f the Tween-80 and p-cyclodextrin
from the soil organic matter.
67
INTRODUCTION
Partitioning o f non-ionic organic compounds (NOCs) into soil organic matter
(SOM) is considered to be an entropy driven process where an increase in disorder occurs
when the solvated shell o f w ater around the contaminant in solution collapses as the
contaminant partitions out o f the aqueous phase (Chiou, 1989; Traina and Onken, 1991).
Bi-phasic sorption has been noted for NOCs and compounds such as atrazine wherein
some fraction o f the compound comes into a rapid equifrbrium with the soil while a
smaller fraction continues to sorb more slowly (Robinson et a l, 1990; Pignatello et al.,
1993; Gaber et al., 1997). Various multi-compartment sorption models have been
proposed to explain these findings, including intra-particle diffusion (Steinberg et al.,
1987; W u and Gschwend, 1986; Scow and Alexander, 1992) and intra-organic matter
diffusion (Brusseau et al., 1991).
Estimates o f the amount o f bound residues remaining in soil are most often made
' '
•
on the basis o f standard sorption isotherms, but because sorption equilibrium is slow to
occur, 24 hr sorption isotherms will always underestimate the sorbed residues o f historic
contamination (Pignatello and Huang, 1991; Pignatello et al., 1993; Xue and Selim,
1995). Capriel et al. (1985) have reported that atrazine was still present as the parent
compound in soils over 9 years after application but they did not attempt to quantify how
much remained. Pignatello and Huang (1991) have investigated the optimization o f
atrazine and metolachlor extraction using standard organic solvents in samples aged
between 8 and 17 months. As much as 70% o f the added compound was not removed
even by hot solvents after 24 hours. Efforts to identify and remediate this ‘bound’
V
68
fraction, however, are problematic, as sorption tends to protect NOCs from most
biological and chemical interactions (Sims et a l, 1991; Guerin and Boyd, 1992; Scribner
et al. 1992). Khan and BehM (1990) used aged and solvent-extracted organic muck soil
and incubation with a Pseudomonas culture to release and degrade almost 35% o f
strongly sorbed atrazine residues. Guerin and Boyd (1992) investigated the bioavailability
o f sorbed naphthalene after aging 25 days, concluding that sorbed naphthalene was
directly available to microbes for degradation. Radesovich et al. (1997), have degraded
aged atrazine residues directly without attempting to desorb them. They noted that while
inoculation with an atrazine mineralizer resulted in production o f 14C 0 2 (85% in freshly
spiked samples), increases in aging time decreased mineralization as atrazine sorbed to
the soil (only 40% 14C 0 2 after aging 90 d).
In all o f the above experiments, a significant fraction o f the NOC sorbed to the
soil was left undisturbed regardless o f treatment (often >50%). This emphasizes that a
significant percentage o f the NOC is not being removed by the techniques currently in use
and that development o f techniques to quantify and remediate these recalcitrant bound
residues is o f environmental interest. The efficient mobilization and removal o f these
residues is also a desirable economic goal. The contamination which remains in aged and
solvent- extracted soils is representative o f this extremely resistant fraction o f strongly
sorbed residues. No studies which address desoxption o f this resistant fraction using
surfactants are available.
Aqueous experiments indicate that for a variety o f NOCs over a wide range o f
aqueous concentrations, the attractive forces o f NOCs towards SOM and surfactant
69
micelles are approximately equal (SDS surfactant, Valsaraj and Thibodeaux, 1989;
Jafvert, 1991). This suggests that NOCs should readily partition from SOM into a
surfactant micelle and that addition o f surfactants to contaminated soil should enhance
removal o f NOCs from soil. Surfactants have proven to be effective' at increasing
mobilization o f NOCs below the critical micelle concentration (CMC) and increasing
solubilization o f NOCs above the CMC for freshly spiked soils (Abdul and Gibson 1991;
Aronstein and Alexander, 1992; Tsomides et al., 1995; Bai et a l, 1997). Wilson and
Clarke (1994) found an increase in solubility o f naphthalene in water using SDS (524 vs
29 mg L"1) and a strong correlation (r = 0.98) between the octanol/water and the
micelle/water partition coefficients for a variety o f other hydrophobic contaminants.
However, none o f these surfactant studies investigated the release or bioavailability o f
aged residues in soils.
h i an effort to increase the removal and efficiency o f remediation o f this
environmentally important fraction o f soil contamination, work was initiated to
investigate the effect o f surfactants on samples o f aged and solvent-extracted soils
containing strongly bound residues o f two common hydrophobic pollutants; atrazine,
(H20 solubility 33 mg/L), and naphthalene (33-35 mg/L). The objectives were 1) to
determine if various surfactants were capable o f increasing the desorption o f these
strongly bound residues, and 2) to determine the biodegradation potential o f the.released
residues.
70
MATERIALS AN® METHODS
'Soils
Soil #1 was a Vinton fine sand (sandy, mixed, thermic, Typic Torrifluvent) from
the University o f Arizona Veterinary Research Property, Tucson, AZ. Soil #2 was a 9:1
mixture o f two soil types, a Superstition sand (sandy, mixed, hyperthermic, Typic
Calciorthids) from the Yuma Mesa Agricultural Center, Yuma, AZ, and a Dothan loamy
sand (fine loamy, siliceous, thermic, Plinthic Kandiudult) from the Central Crops
Research Station near Clayton, NC. Characteristics o f these soils and the mixture are
given in Table 1. The air-dried, mixed soil was passed through a 2 mm sieve and stored in
plastic containers.
Chemicals
Atrazine and [14C]atrazine (specific activity 14.6 pCi/mg) were obtained from
CIBA Corp, Greensboro, NC. Naphthalene, [14C]naphthalene (specific activity .356
pCi/mg), solvents and surfactants were obtained from Sigma Chemical Co. (St. Louis,
MO). (3-cyclodextiin was supplied by Cerestar Co. (Hammond, IN) with no purity
record. All radiolabeled compounds were uniformly ring labeled and 98-99% pure.
Surfactants tested were chosen to represent several classes o f surfactants
including anionic (SDS), non-ionic (Tween 80), and a non-micelle forming surfactant (pcyclodextrin). The critical micelle concentration (CMC) is 2336 m g/L for SDS and 13-15
mg/L for Tween 80 (Sigma, 1991). P-cyclodextrin is a cyclic oligosaccharide produced
by the microbial degradation o f starch and forms 1:1 inclusion complexes, rather than
micelles, with the contaminant, partitioned inside the internal cavity o f the cyclic product.
71
Table A2-1. Soil Characteristics1
Soil
Textural
Soil Texture2
Organic
Class
Sand/Silt/Clay
M atter3
pH
Vinton (soil #1)
Fine sand
90
7.0
3.0
<0.1
8.4
Dothan
Loamy sand
88
8.1
3.9
1.0
5.8
Superstition
Sand
89
7.9
2.3
0.04
8.81
Mixed Soil (soil #2)
Sand
94
4.8
1.1
0.04
8.58
^Analysis performed by Soil Testing Lab, University o f Arizona.
2Hydrometer method.
3Oven Oxidation method.
72
Cyclodextrins do reduce the surface tension o f aqueous solutions (Wang and Bmsseau,
1993). The tw o commercial surfactants are easily degraded (Swisher, 1987; White, 1993)
and cyclodextrins, which are natural products, are assumed to be degradable.
Production of sorbed residues
-
l
Strongly bound residues o f atrazine and naphthalene were generated in previous
experiments; details on application, aging and extraction can be found in Miller et al.
(1997). Briefly, organic contaminants were applied to the soil in an organic solvent which
was then evaporated. Fresh, microbially active soil was added and the samples were
incubated for up to 16 weeks, allowing biodegradation to occur. Afler incubation,
samples were extracted once with methanol, vacuum filtered and air-dried. These samples
z,
were bulked together to provide a large composite source o f soil containing aged bound
residues with a consistent Specific activity. To quantify the extent o f bound residue
formation, contaminated soil was finely ground and duplicate 1 g samples oxidized in a
biological oxidizer (0X 300, Harvey Instrument Corp., NJ) at 875 to 900°C with the
resultant 14C 0 2 being trapped in 15 mL o f C 0 2 trapping scintillation cocktail (Harvey
Corp.) for radioassay (Tri-Carb 1600, Packard Instrument Corp., Meridan, CT).
D esorption experim ents
Surfactant solutions were prepared so that the surfactant concentrations delivered
ranged over at least 3 orders o f magnitude and bracketed the CMC.
Atrazine desorption was evaluated by adding 5 g o f contaminated soil to 10 ml o f
surfactant solution in 50-ml glass centrifuge tubes. Each sample was tumbled for 24 hrs
and then centrifuged at 4,000 rpm for 30 minutes. Duplicate 1 ml samples o f the
73
supernatant were removed and placed in 15 ml o f Scintiverse BD for radioassay.
Naphthalene desorption was determined in similar fashion except 10 g o f soil and 30 ml
o f surfactant solution were used. After desorption was measured, the supernatant was
removed and fresh surfactant solution added to determine whether any further desorption
would occur. This was repeated 5 times. For all samples 10 ml o f supernatant was saved
for experiments to determine biodegradation potential.
Regressions were performed using SigmaPlot v.2.0. All tests employed duplicate
samples to reduce error and all numbers reported are the average o f two observations.
Error bars in the figures represent the standard deviation.
Isolation of m icrobial degraders an d biodegradation of desorbed residues
A bacterial culture capable o f mineralizing naphthalene as a sole carbon source
was isolated from soil #2. Naphthalene mineralizers were obtained by enrichment in the
following mineral salts media: (in g/L; KH2P 0 4, 1.0; Na2H P 0 4, T.0; N H 4N 0 3, 0.05;
(NH4)2S 0 4, 0.5; M gS 04*7 H20 , 0.2; CaCP2H 20 , 0.02; FeCl3, 0.002; M n S 0 4*2 H20 ,
0.002, pH 7.2). Naphthalene was added as the sole source o f carbon (0.1 g). This culture
was the source o f the inocula used for subsequent biodegradation experiments. Ten ml
supernatant samples containing desorbed residues from the desorption experiments were
transferred into 15-ml test tubes and spiked with 1 ml o f the isolated naphthalene
mineralizing population along with 1 ml o f mineral salts media. Tubes were then placed in
microcosms (250-ml glass jars with Teflon lined fids) for incubation. A 20-ml scintillation
vial containing 1 ml o f 0.1 M N aOH was added to each microcosm to trap evolved 14C 0 2.
A t each sampling time the vials were removed, 15 ml o f scintillation cocktail added
74
(Scintiverse BD, Fisher, Los Angeles, CA), and the radioactivity in each sample
determined.
RESU LTS
Q u an titatio n of .sorbed residues
14C-Bound residues in each soil were quantified by oxidation o f triplicate 1 g
samples. It was assumed that all the 14C-bound residues were either atrazine or
naphthalene. Soil #1 contained 307 pg atrazine/g soil (11,700 dpm/g soil) and soil #2
contained 153 pg atrazine/g soil (13,900 dpm/g soil). For naphthalene, soil #2 contained
240 pg naphthalene/g soil (4900 dpm/g soil). Periodic oxidation o f duplicate 1 g samples
indicated that the soil residues o f atrazine were completely stable for the 4 months tested
(data not shown). Soil residues o f naphthalene showed a 10.5% variation over the same
time period.
D esorption experim ents
Table 2 shows the removal o f bound residues o f naphthalene and atrazine from
both soils by the three surfactants. All naphthalene experiments were performed using soil
#2. Trial 1 was an initial screening experiment to determine the range o f surfactant
concentration required and to determine the biodegradation potential o f the released
residues. The success achieved in desorbing naphthalene residues at very high rates o f
i
SDS (10,000 mg/L) prompted a second experiment which bracketed the CMC more
closely (Trial 2). In th is,experiment, the highest rate o f SDS tested (10,000 mg/L)
removed 3.2 times the naphthalene residues as were removed by water alone (44.0 vs
13.7%) in a single wash. Enhanced removal o f bound naphthalene residues started just
/
Table A2-2. Percent Desorption o f Atrazine and Naphthalene from Aged, SolventExtracted Soils
...
SDS concentration (mg/L)
0.0
100
500
1000
Naphthalene
Trial 1 .9 .9 1
11.4
12.3
9.9
Trial 2
13.7
14.5
11.6
12.2
Atrazine
S o il# l
13.7
14.9
2000
4000
8000
10000
48.6
14.4
14.1
24.6
42.9
13.4
44.0
18.7
Tween 80 concentration (mg/L)
Naphthalene
Atrazine
Soil #1
Soil #2
0.0
5
10
15
20
25
100
8.7
7.2
7.7
8.9
8.3
9.9
11.2
12.4
16.3
17.0
16.4
15.4
15.1
14.1
13.1
16.4
Cyclodextrin concentration (% solution w/v)
0.0
0.1
0.25
0.5
0.75
1.0
Naphthalene
15.4
17.4
16.3
17.7
18.8
16.4
Atrazine
S o il# l
11.7
11.2
11.3
11.7
11.7
11.9
Tsach number is the average o f two observations and is reported as a percent o f the
originally sorbed material.
76
below the CMC, increasing linearly with SDS concentration between 2,000 and 8,000
mg/L but then reached a plateau between 8,000 and 10,000 mg/L. Additional washes
with SDS continued to remove radioactivity at all SDS concentrations but at a lower rate
(Figure 1). After 5 washes, a total o f 66.3% o f the bound naphthalene residues could be
removed (10,000 mg/L).
None o f the surfactants tested on either soil enhanced the amount o f radioactivity
released from strongly bound residues o f atrazine vs a control (Table 2). However, all
aqueous solutions, regardless o f the presence, type, or concentration o f a surfactant,
removed between 12 and 19% o f the sorbed atrazine from both soils, material which had
not been desorbed by the methanol extraction procedure. Tween-80 and P-cyclodextrin
also failed to enhance the release o f strongly bound residues o f naphthalene.
B iodegradation of D esorbed N aphthalene Residues
Mineralization o f released naphthalene residues occurred at all levels o f SDS
(Figure 2). Preliminary experiments using samples from Trial 1 had indicated that the
addition o f nutrients was necessary to enhance biodegradation o f the released naphthalene
residues (data not shown). This is consistent with the w ork o f others (Aug and Abdul,
1992; Kommalapati and Roy, 1996). Accordingly, in samples from Trial 2, 1 ml o f the
mineral salts media used to isolate the degrading culture was added to each tube
containing released residues. This was in addition to any nutrients available in the 1 ml
inocula. Microbial growth was present in all samples as evidenced by a brown pigment
and heavy turbidity within all o f the tubes. This was particularly true at higher SDS
concentrations and suggests that SDS was being preferentially utilized as an alternative
77
oZ
W ash # 3 r 2=0.700
W ash #4 r 2=0.144 /
W ash #5 r 2=0.524 / /
W ash #2 n
2000
4000
6000
8000
SDS Concentration (mg/L)
Figure A2-1
Effect o f M ultiple SDS W ashes on Desorption o f
Bound N aphthalene Residues from Soil
10000
78
—o— Control
—o — 100 mg/L
a -.- 500 mg/L
- v
1000 mg/L
—O— 2000 mg/L
—o— 4000 mg/L
—d- - 8000 mg/L
- o - 10,000 mg/L
-
Time (days)
Figure A 2-2
Effect o f SDS C oncentration on the Rate o f 14C 0 2
Produced from D esorbed N aphthalene Residues
79
carbon source (Rouse et al., 1995; White, 1995). In the absence o f surfactant, 12.5 jrg o f
14C 0 2 were produced after 76 days (Figure 2). This represented 38% o f the released
radioactivity. Production o f 14C 0 2 appeared to be inhibited by increasing concentrations
o f SDS. There was a 7 day delay in the production o f 14C 0 2 at all SDS concentrations
except 100 mgZL. Rates o f 14C 0 2 production were comparable during the next 5 weeks
but then appeared to plateau. While the 4,000 mg/L treatment released almost twice the
bound naphthalene residues as did the 100 mg/L treatment (25 vs 14% o f the initially
bound residues), equal amounts o f 14C 0 2 were produced (9.39 vs 9.55 pg naphthalene
mineralized) (Figure 2). Due to the differences in the total amount released, this
represented 16.0 and 27.6% mineralization o f the released residues.
DISCUSSION
D esorption experim ents
The level o f bound naphthalene residues (240 pg/g soil) constitutes a substantial
reservoir o f potential contamination. The enhanced removal ofbound naphthalene
residues by SDS therefore represents a significant reduction in risk to vulnerable
groundwater supplies. Previous studies have demonstrated 15 to 77% removal o f freshly
spiked naphthalene, 1-methylnaphthalene or 2-methylnaphthalene using chemical and
biological surfactants (Jain et a l, 1992; Van Dyke et a l, 1993a, 1993b; Scheibenbogen et
al., 1994). The release o f almost half o f the strongly bound residues o f naphthalene in a
single wash indicates that SDS can be effective on aged contamination as well. The 8,000
and 10,000 mg/L treatments were particularly effective at solubilizing strongly bound
80
residues. The current results are noteworthy because they successfully address one o f the
most recalcitrant fractions o f soil contamination.
O f interest in the atrazine experiments was the fraction (11.7-16.3%) o f strongly
sorbed atrazine which was released by water. The release into aqueous solution o f hound
;.
. . .
. . .
.
'
■
residues o f atrazine is mirrored nationwide in the steady increase in low-level detections
o f atrazine and atrazine degradation products in groundwater throughout American
agricultural regions (Maas et al., 1995; Heimann et al., 1997). The current results suggest
)
that while atrazine is known to participate in varied reactions within soil and SOM, simple
diffusion into the three-dimensional matrix o f the SOM may account for between 10 to
20% o f the behavior o f atrazine in soil (Xing et al., 1996). The roughly 14% o f atrazine
residues released represents enough atrazine (42.1 pg/g o f soil) to be considered a
groundwater pollution problem and the 86% remaining (252 pg/g o f soil) represents a
i
_
'
substantial reservoir o f potential future contamination.
The failure o f the other two surfactants to enhance release o f b ound .residues may
be related to molecular structure. Tween-80 and p-cyclodextrin are both large, relatively
bulky molecules with high molecular weights (MW)(e.g., Laha and Luthy, 1992; Wang
and Brusseau, 1993) and may be restricted from entering into the complicated threedimensional structure o f SOM (Khan, 1980). In contrast, SDS (MW=288) is much
smaller than both cyclodextrim (M W =1300) and Tween-80 (M W =1310), and has a
shorter hydrophobic chain length (Cn vs C17) and smaller hydrophilic head than Tween80, an ethoxylated sorbitan sugar ( Wang and Brusseau, 1993; Jafvert et al., 1995). In
81
addition, SDS is a linear molecule and the differences in size and shape may play a role in
the differences noted in efficacy between these classes o f surfactants.
It is also possible that the failure o f the cyclodextrin and Tween 80 to enhance
desoiption o f atrazine and naphthalene was due to sorption o f the surfactant itself to the
soil. Brusseau et a!. (1994) have reported that cyclodextrins show a low affinity for soil,
including high organic matter soils (12.6% O.C.) so it is unlikely that sorption affected
the cyclodextrin results in this case. Laha and Luthy (1992) reported that the addition o f
1 g o f a silt loam soil containing 1.5% organic carbon increased the effective CMC o f
Tween 80 from 14 mg/L to 2247 mg/L. By plotting their measured effective CMC vs the
organic carbon content o f their soil, the effective CMC for this soil was estimated. The
soil used for the current experiments contained 0.04% organic carbon and the effective
CMC was calculated to be 78 mg/L. The highest rate o f Tween 80 addition was 100
mg/L so sorption could have affected the current results by reducing the concentration o f
y
Tween 80 in solution or by promoting resorption o f released naphthalene to the Tween
sorbed to soil surfaces.
B iodegradation of D esorbed N aphthalene Residues
In this study, initial inhibition o f naphthalene mineralization occurred at all SDS
concentrations except for 100 mg/L (Figure 2). The strongest inhibition occurred at 500
mg/L and 1000 mg/L. This corresponds to lower desorption o f naphthalene as well and
m aybe related to a lack o f micelles (CMC= 2336 mg/L). The optimum SDS
concentration for both release and biodegradation o f bound naphthalene residues
appeared to be between 4,000 and 8,000 mg/L. While high concentrations o f SDS were
82
successful at releasing bound naphthalene residues, only 7-28% (o f the total residues
released) w ere mineralized in samples containing SDS vs 38% in the absence o f surfactant
(control). The rest o f the released radioactivity in the control may have been incorporated
into the microbial biomass produced. Conditions which encourage complete and rapid
degradation can be optimized with further research.
CONCLUSION
The removal o f bound residues o f naphthalene from aged, solvent- extracted soils .
represents a success in attempts to understand and remediate a difficult fraction o f
environmental contamination. As much as 66.3% o f the previously unextractable residues
o f naphthalene were removed by washing with an SDS solution (10,000 mg/L).The
successful removal o f naphthalene aged naphthalene residues suggest that further
experiments with field soils containing historic contamination may be worthwhile.
Biodegradation o f the released naphthalene residues was evident at all SDS
concentrations after addition o f nutrients. However, even though mineralization occurred,
it was inhibited to some degree by SDS concentrations greater than 100 mgZL compared
to a water control. Investigation into conditions which promote the full biodegradation o f
released naphthalene residues in the presence o f surfactants such as SDS should prove
useful for remediation studies. The removal o f strongly sorbed residues o f atrazine by
water alone (11.7-18.7%) indicates that desorption o f atrazine from soils which contain
sorbed residues can continue for months, if not years. Attempts to enhance this removal
using 3 different classes o f surfactant were not successful.
83
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