Biodegradation of 2,4-dichlorophenoxyacetic acid (2,4-D) and pentachlorophenol (PCP) in soils
by Ronald Allen Doughten
A thesis submitted in partial fulfillment of the requirements for the degree of Master of Science in Soils
Montana State University
© Copyright by Ronald Allen Doughten (1997)
Abstract:
Contamination of soil and water by organic pollutants is a widespread problem that has resulted in a
great deal of public concern. 2,4-dichlorophenoxyacetic acid (2,4-D) and pentachloropheriol (PCP) are
representative members of a large class of chlorinated pesticides. Furthermore, both compounds are of
local concern in the state of Montana. Contaminant degradation in soils generally occurs by the action
of microorganisms. However, biodegradation is often limited by chemical sorption to soil organic
matter (SOM). To overcome this limitation, surfactants have successfully been used to solubilize
contaminants from soil. The impact of SOM on 2,4-D sorption, availability, and degradation was
investigated in 31 soils with a range of organic matter content. Biodegradation was measured as the
accumulation of 14CO2 produced from 14C-2,4-D treated soils and evaluated against sorption to SOM.
Surfactant influences on solubility and degradation were examined in 14C-PCP labeled soils inoculated
with white-rot fungi. The extent of 2,4-D degradation was negatively correlated with SOM, illustrating
the influence of sorption on bioavailability and degradation. However, the relationship to the partition
coefficient, Kd, was much weaker, suggesting that mass transfer and slow desorption may be the
dominant mechanisms controlling contaminant bioavailability in soils. The extent of PCP degradation,
measured as the accumulation of 14CO2 from 14C-PCP treated soils was unrelated to contaminant
solubility. The addition of 25 mg Tween 80 g soil-1 resulted in enhanced PCP degradation (26%)
relative to a surfactant-free control (17%) despite an overall decrease in the aqueous solubility.
Fungal activity measured as biomass and respiration over the same time period increased, suggesting a
physiological response rather than a chemical response to surfactant addition. We suggest that
contaminant sorption to SOM strongly influences degradation, but mass transfer from the sorbed state
controls bioavailability. Furthermore, the addition of surfactants at sub-critical micelle concentrations
(CMC) can result in increased degradation without increasing overall solubility, and may therefore
represent an effective strategy for the bioremediation of organic-contaminated soils. BIODEGRADATION OF 2,4-DICHLOROPHENOXYACETIC ACID (2,4-D) AND
PENTACHLOROPHENOL (PCP) IN SOILS
by
Ronald Allen Doughten
A thesis submitted in partial fulfillment
o f the requirements for the degree
of
Master o f Science
in
Soils
MONTANA STATE UNIVERSITY—BOZEMAN
Bozeman, Montana
April 1997
11
J ) i +Kl 6
APPROVAL
o f a thesis submitted by
Ronald Allen Doughten
This thesis has been read by each member o f the thesis committee and has been
found to be satisfactory regarding content, English usage, format, citations, bibliographic
style, and consistency, and is ready for submission to the College o f Graduate Studies.
William P. Inskeep
Approved for the Department o f Plant, Soil, and Environmental Science
Jeffrey S. Jacobsen
(SigM ture)
^
v /" > A ?
(Date)
Approved for the College o f Graduate Studies
W
Robert L. Brown
(Signature)
(Ddfe)
f?
Ill
STATEMENT OF PERMISSION TO USE
In presenting this thesis in partial fulfillment o f the requirement for a master’s
degree at Montana State University—Bozeman, I agree that the Library shall make it
available to borrowers under rules o f the Library.
IfI have indicated my intention to copyright this thesis by including a copyright
notice page, copying is allowable only for scholarly purposes, consistent with “fair use”
as prescribed in the U.S. Copyright Law. Requests for permission for extended quotation
from or reproduction o f this thesis in whole or imparts may be granted only by the
copyright holder.
Signature
D ate. / 6 " 6 W
I
(
iv
This thesis is dedicated to my father, Randall Doughten, who was taken from my life at
far too early o f an age. His influence and memory I carry with me always.
V
VITA
The author, Ronald Allen Doughten was bom to Mr. and Mrs. Randall Doughten
on January 23, 1971, in Havre, Montana. He graduated from Havre High School, Havre,
Montana in 1989 with honors. In 1994, Ronald received a Bachelor’s degree in
Chemistry from Montana State University. He graduated with highest honors. In August
o f 1994, he began working towards a Master o f Science degree at Montana State
University-Bozeman under the direction o f Dr. William P. Inskeep.
ACKNOWLEDGEMENTS
I would like to thank my major advisor Dr. William Inskeep for the ideas,
support, and encouragement he has given me as an undergraduate and graduate student at
Montana State University.
I would also like to thank the other members o f my committee, Dr. Jon Wraith,
Dr. Timothy McDermott, and Dr. Anne Camper, for their assistance and use of
laboratories and equipment. I owe thanks also to Dr. Carl Johnston and Michael Beccera
o f Mycotech Corporation for their assistance in my work with the white-rot fungi.
I am grateful to my fellow lab workers, Richard Macur, Dr. Hesham Gaber, Dr.
Robert Grosser, Dr. Shaobai Sun, Clain Jones, Heiko Langner, and countless other
students who have contributed in some part to my life and research in Bozeman.
Finally, I would like to give special thanks to my parents, Mr. and Mrs. Ralph
Prosser, for their unending moral and spiritual support through my education.
Vll
TABLE OF CONTENTS '
APPROVAL....................................................................................................
STATEMENT OF PERMISSION TO U SE .....................................................................
iii
VITA......................................................................................................................................
v
ACKNOWLEDGEMENTS........... .........................................................................................
TABLE OF CONTENTS.....................................................................................................
vi
vii
LIST OF TABLES....................................................................................................................
ix
LIST OF FIGURES.................................................................................................
x
ABSTRACT.............................................................................................................:.............
xv
1. INTRODUCTION.... ..............,......................................................................................
1
The Need for Contaminant Remediation......................................................
Fate o f Organic Chemicals in the Environment..........................................................
3
Strategies for Remediation o f Contaminated Sites.........................................................
4
Traditional Strategies to Remediation........................ :........ .................................
4
Bioremediation..................................................................
Chemical Sorption Limitations to Bioavailability and Biodegradation.......................
9
Chemical Sorption to Soil.............................................................................................
9
Mechanisms o f Sorption to Soil...............................................................................
11
Sorption and Bioavailability........................................................................................
12
The Role o f Surfactants in Remediation..........................................................................
14
Surfactant Basics............................................................................................................
14
Solubility Enhancement o f NOCs in Aqueous Solution..........................................
16
Solubility Enhancement of NOCs in Soil Systems...................................................
17
Biodegradation o f Organic Contaminants in Soils Treated with Surfactants.... 19
Biosurfactants................................................................................................................
21
Objectives............................................................................................................................
23
2. BIODEGRADATION OF 2,4-DICHLOROPHENOXYACETIC ACID (2,4-D)
IN THIRTY-ONE AGRICULTURAL SOILS.......................................................
25
V lll
Introduction......................................................................................................................
Materials and Methods....................................
Chemicals....................................................................................................................
S o ils .............................................................................................................................
Batch Biodegradation.................................................................................................
Batch Sorption...........................................................................................................
Bacterial Plate Counts...............................................................................................
Analytical Methods....................................................................................................
Results and Discussion..................................................................................................
Variation in 2,4-D Degradation Among S o ils.......................................................
Relationship Among 2,4-D Degradation, SOC, and 2,4-D Sorption..................
Relationships between Degradation and Soil Bacteria Numbers.......................
Moisture Effects on 2,4-D Degradation..................................................................
Anaerobic versus Aerobic Degradation..................................................................
Summary...........................................................................................................................
26
29
29
29
32
33
33
34
35
35
39
44
44
47
49
3. MINERALIZATION OF PENTACHLOROPHENOL (PCP) IN SOIL BY
WHITE-ROT FUNGI IN THE PRESENCE OF SURFACTANTS................
Introduction.......................................................................................................................
Materials and Methods....................................................................................................
Chemicals....................................................................................................................
S o il...............................................................................................................................
Surfactants...................................................................................................................
Fungi............................................................................................................................
Batch Biodegradation................................................................................................
PCP Extraction and HPLC Analysis............ ...........................................................
PCP Sorption and Solubility.......................................................
Fungal A ctivity...........................................................................................................
Results and D iscussion...................................................................................................
Effects o f Surfactants on PCP Mineralization.......................................................
Potential Mechanisms o f Surfactant Enhanced PCP Mineralization.................
Summary...........................................................................................................................
Acknowledgements..........................................................................................................
50
51
35
35
56
57
58
58
59
60
61
62
62
66
72
73
4. SUMMARY..................... :........................................................................................... 74
References Cited...................................................................................................................
76
Appendices...........................................................................
85
Appendix A: Additional 2,4-D Data............................................................................
Appendix B : Additional PCP Data..............................................................................
86
91
ix
LIST OF TABLES
Table
1. Soil series, locations, land-use specification, texture, pH, EC, and soil organic
carbon for the 3 1 soils used in the study..............................................................
30
2. Comparisons between the three models used to describe the observed 2,4-D
degradation in six o f the 3 1 soils studied.............................................................
37
3. Characteristics o f the Amsterdam Silt Loam (Typic Cryoboroll)...........................
56
4. Soil characteristics for three PCP contaminated soils used in PCP degradation
studies.......................................................................................................................
92
X
LIST OF FIGURES
Figure
1. Locations o f the nine EPA National Priority Sites in Montana. The shaded labels
are those contaminated with organic chemicals; the others are contaminated
with inorganics.........................................................................................................
2
2. Processes influencing the fate o f organic chemicals in the environment (Weber
and Miller, 1989).....................................................................................................
3
3. Aerobic metabolism o f a substrate to cellular components and COi by an attached
bacterium.......i................................................
6
4. Sorption in soil involves both adsorption to surfaces and partitioning into the bulk
solid............................................................................................................................
9
5. Surfactants can be classified into four groups based upon the charge o f their polar
group— nonionic, anionic, cationic, and zwitterionc— containing no charge, a
negative charge, a positive charge, and both positive and negative charges,
respectively ......................................................... .................................................
15
6. Interactions between sorbed contaminants, soil bacteria and fungi, and surfactants
in enhancing solubility o f NOCs in soil environments......................................
. 18
7. Experimental setup used in the batch biodegradation studies. Compressed air or
nitrogen was passed through a NaOH trip into a manifold that controlled gas
flow to individual flasks containing 14C-2,4-D labeled soil. Degradation was
measured as the accumulation o f 1 CO2 in 0.5M NaOH traps changed
regularly..............................................................................................................••••••
32
8. Degradation curves showing the fraction o f 14C-2,4-D evolved as 14C02 from six
o f the 3 1 soils. The organic carbon content o f the soil is shown adjacent to
each curve.................................. ..............................................................................
35
9. Top figure: Comparison of fitted 2,4-D degradation rate constant (h-1) from the
modified first-order model to SOC. Bottom Figure: Maximum extent of
degradation fitted from the modified first-order model compared to SOC....
40
I
xi
10. Comparison o f the degradation rate constant (h'1) and Pmax, values obtained
from modified first-order m odel............................................................................
41
11. Relationship o f 2,4-D partition coefficients (Kd) to soil organic carbon content
for all 3 1 soils. The slope yields and average Koc value o f 422 L kg'1..........
42
12. Maximum extent o f 14C-2,4-D degradation plotted against soluble 2,4-D in
solution at t = 0 (calculated from Kd). Solid line is the observed relationship
between available and degraded 2 ,4 -D .................................................................
43
13. Relationship between fitted extent o f 2,4-D degradation (Pmax), rate constants,
and DOC to the number o f colony forming units o f bacteria isolated on nonselective media in six soils (bracketing the range in SOC contents observed
for. the soils studied).........................................:......................................................
45
14. Effect o f moisture content on the degradation o f 14C-ul-2,4-D in two soils.
Moisture content o f 0.55 is above saturation; 0.36 is approximately field
capacity.................................................................. ...................................................
46
15. Effect o f soil preconditioning under anaerobic condition on the mineralization
o f 2,4-D in Amsterdam and Brocko soil (0m= 0.46). A: 14COz evolution
from 14C-ul-ring-labeled 2,4-D under aerobic conditions. B: 14COz
evolution from 14C-ul-ring-labeled 2,4-D preconditioned under anaerobic
conditions, then exposed to aerobic conditions..................................................
48
16. Structures o f the nine surfactants used in the study. The chemical formulas for
the five Tween surfactants are given as the R group..........................................
57
17. Experimental setup for measuring PCP mineralization from contaminated soil..
59
18. Mineralization o f PCP in Amsterdam soil by Phanerochaete chrysosporium in
the presence o f 25 mg surfactant g soil'1 .............................................................
62
19. Mineralization o f PCP in Amsterdam soil by a non-Phanerochaete species in'
the presence o f 25 mg. surfactant g soil"1.............................................................
63
20. Fungal respiration estimated from total COz produced from Amsterdam soil
columns treated with a non-Phanerochaete species and five surfactants........
64
21. Mineralization o f PCP (closed symbols) in Amsterdam soil with and without the
addition o f 25 mg Tween 80 g soil"1 by a non-Phanerochaete species,
followed by solvent extraction o f PCP (open symbols) remaining in the soil.
65
22. Mineralization o f PCP by a non-Phanerochaete species in Amsterdam soil
amended with 25 mg Tween 20,40, 60, 80, and 85 g s o il1..............................
67
xii
23. Mineralization o f PCP by a non-Phanerochaete species in Amsterdam soil
amended with 0, 0.05, 0.5, 1, 5, or 25 mg Tween 80 g soil'1.............................
68
24. Aqueous concentration o f PCP and aqueous phase surface tension from
Amsterdam soil equilibrated for 24 h with varying concentrations o f Tween
80....................................................... .........................................................................
69
25. Sorption isotherms o f PCP to Amsterdam soil in the presence o f Tween 80 at
concentrations o f 0, 1.5, 15, and 150 mg Tween 80 g soil'1. The resulting Kd
values are 1.14, 14.04, and « 1 L kg'1, respectively..........................................
70
26. Total COz production from Amsterdam soil treated with a non-Phanerochaete
species and treated with or without 25 mg Tween 80 g soil"1. Also shown is
the amount o f ergosterol present in the soil over the same time period..........
72
27. Degradation o f 2,4-D in 3 1 agricultural soils collected from around Montana.
Soils demonstrated a wide range in both rates and extents o f degradation.
Degradation curves are shown with data points for clarity................................
87
28. Comparison o f three kinetic models used to describe 2,4-D degradation in soil.
The points are the observed data; the solid line is the fit from the first-order
kinetics; the dashed line is the fit from modified first-order; and the dotted
line is the logistic model. Also shown are r2 values from non-linear
regression analysis. The modified first-order and logistic equations
comparably described the observed degradation curves....................... :...........
88
29. Relationship between maximum extent o f 2,4-D degradation and fitted rate
constants (from modified first-order) to soil organic carbon for the cropfallow and native range groups o f soils. A linear regression between extent
o f degradation and SOC is shown for each set o f soils. The relationship was
stronger for the crop-fallow soils than for the native-range soils. No apparent
relationship was evident between the rate constant and SOC...........................
89
30. Comparisons o f aerobic and anaerobic conditions on 14C-Carboxyl-Iabelled 2,4D degradation (0m= 0.46). Top figure is 14COz production from 14Ccarboxyl-labeled 2,4-D under aerobic conditions. Bottom figure is 14COz
production from 14C-Carboxyl-Iabeled 2,4-D preconditioned under anaerobic
conditions then exposed to aerobic conditions. Total extent o f degradation
increased in the Amsterdam soil when incubated under anaerobic conditions
90
Xlll
3 1. Degradation o f PCP in Amsterdam soil by a non-Phanerochaete white-rot
fungus treated with surfactants at 25 mg g soil"1. The naturally-derived
surfactant lecithin and the synthetic surfactant Witconol SN-70 are compared
with no surfactant and Tween 80 treatments. Both lecithin and Witconol
were less effective than either no surfactant or Tween 80 at promoting PCP
degradation...............
93
32. Effect o f soil mixing on PCP degradation and fungal respiration by a nonPhanerochaete species. Soil was mixed after 28 d and moisture content
readjusted to field capacity. Mixing resulted in a very slight increase in PCP
degradation and fungal respiration........................................................................
94
33. Degradation o f PCP in the Libby Pole soil by indigenous organisms and a nonPhanerochaete species (F600), both with and without the addition o f 25 mg
Tween 80 g s o i l . None o f the treatments showed considerable PCP
mineralization up to 24 d. Increases in total mineralization after 24 d are
accompanied by very large standard errors, indicating non-uniform variation
among replications...................................................................................................
95
34. Degradation o f PCP in the Montana Pole soil by a non-Phanerochaete white-rot
fungus and the surfactant Tween 80 at five concentrations. Tween 80 had no
effect on the degradation o f PCP, perhaps due to toxicity effects. The
Montana Pole soil contains high levels o f heavy metals (e.g. Cu) and other
toxic materials such as arsenic and may have interfered with growth.............
96
35. Degradation o f PCP in Amsterdam soil by indigenous organisms, a nonPhanerochaete white-rot fungus (F600),' and a fungus isolated from PCPcontaminated soils (tentatively identified as a Penicillium species). The nonPhanerochaete fungus was more effective at PCP degradation (17%) than the
indigenous soil organisms (11%); the Penicillium fungus, however, resulted
in greater than 40% degradation o f PCP after four weeks. More research into
PCP degradation by this organism is warranted.................................................
97
36. PCP degradation in Libby Pole soil inoculated with a non-Phanerochaete whiterot fungus and a Penicillium fungus isolated from soil. Neither fungus
enhanced degradation above that o f the indigenous soil microorganisms in
this particular soil................................................................................................... .
98
XlV
37. Surface tension measurements o f Tween 80 solutions at concentration o f 0, 10,
50,200, 500, 1000,2000,4000, 6000, 8000, 10000, and 20,000 mg L'1. The
diamonds are surface tension measurements made in the absence o f soil; the
CMC is estimated at 7.5 mg L'1. The squares are surface tension
measurements mad after 24 h equilibration o f 15 mL o f surfactant solution
with I g o f Amsterdam soil. The estimated concentration required to achieve
CMC in a soil contain system is 2500 mg L'1 or 2.5 order o f magnitude
higher. The impact o f surfactant sorption to soil is well illustrated by this
comparison................................................................................................................
/
XV
ABSTRACT
Contamination o f soil and water by organic pollutants is a widespread problem
that has resulted in a great deal o f public concern. 2,4-dichlorophenoxyacetic acid (2,4D) and pentachloropheriol (PCP) are representative members o f a large class o f
chlorinated pesticides. Furthermore, both compounds are o f local concern in the state o f
Montana. Contaminant degradation in soils generally occurs by the action o f
microorganisms. However, biodegradation is often limited by chemical sorption to soil
organic matter (SOM). To overcome this limitation, surfactants have successfully been
used to solubilize contaminants from soil. The impact o f SOM on 2,4-D sorption,
availability, and degradation was investigated in 3 1 soils with a range o f organic matter
content. Biodegradation was measured as the accumulation o f 14COa produced from 14C2,4-D treated soils and evaluated against sorption to SOM. Surfactant influences on
solubility and degradation were examined in 14C-PCP labeled soils inoculated with whiterot fungi. The extent o f 2,4-D degradation was negatively correlated with SOM,
illustrating the influence o f sorption on bioavailability and degradation. However, the
relationship to the partition coefficient, Kd, was much weaker, suggesting that mass
transfer and slow desorption may be the dominant mechanisms controlling contaminant
bioavailability in soils. The extent o f PCP degradation, measured as the accumulation o f
14COa from 14C-PCP treated soils was unrelated to contaminant solubility. The addition
o f 25 mg Tween 80 g soil"1 resulted in enhanced PCP degradation (26%) relative to a
surfactant-free control (17%) despite an overall decrease in the aqueous solubility.
Fungal activity measured as biomass and respiration over the same time period increased,
suggesting a physiological response rather than a chemical response to surfactant
addition. We suggest that contaminant sorption to SOM strongly influences degradation,
but mass transfer from the sorbed state controls bioavailability. Furthermore, the addition
of surfactants at sub-critical micelle concentrations (CMC) can result in increased
degradation without increasing overall solubility, and may therefore represent an
effective strategy for the bioremediation o f organic-contaminated soils.
I
CHAPTER I
INTRODUCTION
“As man proceeds toward his announced goal o f the
conquest o f nature, he has written a depressing record o f
destruction directed not only against the earth he inhabits
but against the life that shares it with him.” (Carson, 1962)
The Need for Contaminant Remediation
The introduction o f organic chemicals into the environment has resulted in a great
deal o f public concern on local, national, and international levels. In the United States
alone, approximately 80 billion pounds o f hazardous wastes are produced annually. As
recently as 15 years ago, the United States Environmental Protection Agency (US EPA)
estimated that only 10% o f these wastes were disposed o f safely (Epstein et a l, 1982).
Synthetic chemicals in soil, water, and air have been indicted as causes o f illness among
exposed plants and animals. A classic example was the widespread use o f the pesticide
DDT during the 1940s and 1950s. DDT was found to bioaccumulate in the tissues o f
organisms and was passed through the food chain, resulting in the death o f aquatic life,
fish, and birds (including the bald eagle). Because o f its toxicity to many non-target
species including humans, DDT use was banned in the US in 1972. Although awareness
o f the impact o f hazardous materials in the environment has risen, historical use and
practices have led to large-scale pollution problems. As o f 1993, the US EPA managed
2
the remediation o f 1500-2000 sites under the Comprehensive Environmental Response,
Compensation, and Liability Act (CERCLA or Superfund program), 1500-3000 sites
under the Resource Conservation and Recovery Act (RCRA), and 295,000 underground
storage tanks. In addition, 7300 site cleanups were under the administration o f the US
Department o f Defense and 19,000 sites under the US Department o f Energy.
On a local level, the US EPA recognizes nine national superfund priority sites in
the state o f Montana (Figure I; US ERA, 1995). Four o f these sites are predominantly
contaminated by organic pollutants, including diesel fuel, creosote, and PCP
-------------..
, . '.-------I--------1
Libby Groundwater ,------------------------- , Burlington Northern
|
Contamination
I East Helena Site Livingston Shop Complex
__________>
'
I
7
—■
— -j------------------ :
.II h
j rote i““"v
I
... I
X".....
/Ti TiEEl
- {
r
r X ‘i ’.... J r ^ i
Lr
-
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i
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i
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t
i
, .........-
T4
’» u r .
j. C
wsteuD
M
'i , - - - - - T
- 1 .« “ ^ m u.. I vaxowsri—
"7
Silver Bow Creek/
Butte Area
Montana Pole
and Treating
and Treating
Mouat
Industries
Figure I. Locations of the nine EPA National Priority Sites in Montana. The shaded labels are those
contaminated with organic chemicals; the others are contaminated with inorganics.
3
In addition to these higher profile point sources, the application o f pesticides can
result in non-point source soil and groundwater contamination. Well testing programs
around the state o f Montana between 1984 and 1988 revealed the presence o f seven
different pesticide residues in regional groundwaters (DeLuca, 1989).
Fate o f Organic Chemicals in the Environment
Once introduced into the environment, several factors influence the fate o f an
organic chemical (Figure 2). Surface processes such as runoff,
Water table
Figure 2. Processes influencing the fate of organic chemicals (OC) in the environment (Weber and
Miller, 1989).
4
photodecomposition, and volatilization can initially reduce the amount o f contaminant
entering the soil and groundwater system. Uptake by plants and animals in the soil results
in biological accumulation and/or detoxification. Soil transport processes associated with
water movement through leaching or capillary flow (bottom o f diagram) influence the
movement o f organic chemicals, particularly soluble compounds, and are major factors in
assessing groundwater pollution potential. Abiotic chemical interactions between the soil
and contaminant can result in chemical breakdown and detoxification. Other chemical
processes such as sorption to clays, oxides, or soil organic matter (SOM) can strongly
influence chemical mobility, persistence, and availability to soil microorganisms. Finally,
biological degradation (biodegradation) by soil bacteria and fungi is the primary pathway
o f organic chemical breakdown and detoxification in soil.
Strategies for Remediation o f Contaminated Sites
Traditional Strategies to Remediation
Based upon these soil processes, several strategies have been used for the
treatment o f hazardous chemicals in the environment: (I) Physical treatment (e.g.
adsorption, filtration, insoluble phase extraction) alters the physical state o f a hazardous
material and requires further action to actually remove the contaminant. (2) Chemical
treatment is used to detoxify a waste to a less toxic form and includes such strategies as
stabilization, solidification, and encapsulation. Chemical alteration can result in the
formation o f hazardous by-products that require further remediation. (3) Thermal
5
treatment involves the combustion o f hazardous materials. Incineration costs are often
high; gas emissions and solid residues may require further disposal as hazardous wastes.
(4) Removal and burial comprises the physical removal o f a contaminated medium (i.e.
soil, water), and relocation to a landfill or other holding facility. Hazardous waste
removal does not detoxify the medium, and disposal in landfills often results in further
contamination o f soil and groundwater (Alexander, 1994; Fernando and Aust, 1994;
Thayer, 1991).
Bioremediation
An alternate strategy to remediation o f organic contaminants that has received
great amounts o f attention recently is bioremediation, or the use o f biological agents
(primarily microorganisms) to degrade soil and water contaminants. For purposes o f
remediation, biodegradation can be defined as the biologically mediated alteration o f a
contaminant from a more toxic form to a less toxic form. Microorganisms are the
primary agents responsible for the breakdown o f organic molecules in the environment,
ultimately resulting in the production o f harmless end products such as carbon dioxide
and water. Through a biologically mediated redox process, aerobic organisms use
organic compounds as a carbon and energy source and O2 as an electron acceptor (Figure
3).
Under the appropriate conditions, microorganisms have successfully been used to
remediate hazardous waste sites. Probably the best know example o f the large scale use
o f bioremediation technologies was the effort to clean shorelines impacted by the 1989
Exxon Valdez oil spill in Prince William Sound, Alaska. The application o f an oleophilic
/
X
6
fertilizer, consisting o f oleic acid, urea, and lauryl phosphate, to oil-contaminated shores
Organic C
.Cellular
I J
VZ/
c o m p o n e n t s '^
Substratum
Figure 3. Aerobic metabolism of a substrate to cellular components and CO2 by an
attached bacterium.
stimulated degradation by indigenous microorganisms. A five-fold increase in rates of
biodegradation over untreated sites was observed. The use o f bioremediation resulted in
greater cleanup than physical methods and was conducted at a cost significantly below the
SI million per day expended on soil washing (Pritchard and Costa, 1991).
Most attempts at bioremediation in the US have used two primary approaches: (I)
the stimulation o f indigenous microorganisms through fertilization or (2) the seeding of
organisms into the environment. Stimulation o f indigenous microorganisms involves
furnishing organisms with needed resources that limit degradation o f a specific
contaminant or a suite o f contaminants. The initial steps in the aerobic degradation of
hydrocarbons by bacteria and fungi involve substrate oxidation, a process requiring the
presence o f molecular oxygen. In surface soils and the vadose zone, aerobic conditions
7
normally prevail and can be maintained by adequate surface drainage or a periodic raising
and lowering o f the groundwater. Oxygen concentrations in contaminated aquifers,
however, are very low (O2 concentration at saturation in water is 8 ppm). The addition o f
hydrogen peroxide (H2O2) in stabilized, time release formulations has been used to boost
O2 concentrations in subsurface waters (Atlas and Barth, 1992). A second approach to
stimulating microbial activity involves nutrient addition, the process used in the Alaska
oil-spill. The introduction o f carbon-rich compounds into a soil or water environment (as
occurs in an oil spill) results in a high C to N ratio that is unfavorable for microbial
degradation. The addition o fN , P, or other limiting nutrients has successfully stimulated
bioremediation o f contaminants by indigenous organisms (Atlas, 1995; Atlas and Bartha,
1992).
Seeding is the purposeful introduction o f organisms into the environment for the
sole purpose o f increasing the rate and/or extent o f hydrocarbon degradation above that
capable by native populations. Numbers o f organisms capable o f degrading specific
compounds have been isolated from the environment. Grosser et al. (1991, 1995)
demonstrated a significant increase in polyaromatic hydrocarbon (PAH) degradation in
contaminated soils by the reintroduction o f pyrene degrading microorganisms isolated
from the site. Pyrene degradation by indigenous organisms was approximately 1% in 2 d;
seeding resulted in 55% degradation in 2 d. Likewise, the addition o f chlorobenzoatedegrading bacteria to polychlorinated biphenyl (PCB) contaminated soils resulted in
degradation o f 25% o f the contaminant as compared to only 3% in uninoculated soils
(Hickey et al., 1993).
8
Despite the successes o f some bioremediation strategies, several key limitations
make widespread application difficult and uncertain. Bacteria possess specific enzymes
which are not capable o f remediation o f complex chemical mixtures, as are commonly
found at contaminated sites. Maintenance o f mixed microbial populations capable o f
degrading these mixtures is often difficult or impossible with current knowledge.
Moreover, many seeded organisms are unable to compete with indigenous species, and
the populations required to achieve enhanced degradation cannot be maintained.
Furthermore, many o f the most recalcitrant and toxic compounds are often degraded only
through cometabolism, whereby another C-source is required to maintain microbial
growth. The introduction o f co-substrates into natural environments has met with limited
success as interactions with other organisms in the environment are complex-(Atlas,
1995; Alexander, 1994; Atlas and Bartha, 1992; Thayer, 1991).
Attempts to overcome biological limitations to biodegradation in natural
environments have focused on two approaches. One approach has been to genetically
engineer strains o f bacteria that are competitive in soil environments and have greater
degradation capabilities. In 1981, the first patent for a genetically engineered organism
with contaminant degrading capabilities was granted in the U.S. However, attempts to
modify organisms to produce strains with that can degrade specific pollutants in situ have
met with little success. Furthermore, environmental regulations in most countries restrict
the release o f genetically engineered organisms into the environment (Atlas, 1995).
Another approach has been to use other naturally occurring organisms with
enzymatic systems capable o f degrading a suite o f environmental pollutants. One such
9
genus o f organisms are the white-rot fungi. One species, Phanerochaete chyrsosporium,
uses a non-specific extracellular enzyme system to degrade lignin and has the capabilities
to degrade environmental pollutants as well (Barr and Aust, 1994). The properties o f
white-rot fungi and their applicability to bioremediation will be discussed in further detail
in Chapter 3.
Chemical Sorption Limitations to Bioavailabilitv and Biodegradation
Chemical Sorption to Soil
A major limitation to successful bioremediation is reduced substrate
bioavailability due to chemical sorption. Chemical sorption in soils includes both
adsorption, the retention o f a molecule on the surface o f a solid material, and partitioning.
Partitioning
Adsorption
Organic Matter
/
Figure 4. Sorption in soil involves both adsorption to surfaces and partitioning into the bulk solid.
10
the retention o f solute within the bulk solid (Figure 4). Adsorption occurs primarily to
the mineral, clay, and oxide surfaces in soils; partitioning, especially o f nonpolar organic
chemicals (NOCs), occurs primarily in the organic fraction o f soil. Sorption o f NOCs is
strongly influenced by two factors—the hydrophobicity o f the compound and the amount
o f organic carbon present in the soil.
Chemical hydrophobicity may be expressed in terms o f an octanol-water partition
coefficient (Kow), or the ratio o f the chemical concentration found in octanol to the
concentration present in water (in an octanol-water system):
where S0 (mg L'1) is the chemical concentration in octanol and Sw (mg L"1) is the
chemical concentration in water. The higher the Kow, the more hydrophobic the chemical,
the lower the aqueous solubility, and the greater the amount o f chemical sorption that will
occur (Chiou, 1989).
SOM is a chemically complex structure resulting from the decomposition and
synthesis that occurs during the degradation o f organic materials, primarily plant tissues
(Brady, 1990); SOM is principally composed o f C, H, and 0 . The surface o f an organic
matter aggregate is hydrophilic and consists o f negatively charged hydroxyl, carboxylic,
and phenolic functional groups. The internal structure is predominantly hydrophobic and
is held together by covalent bonds, van der Waal forces, and dipole interactions. The
quantity o f SOM present directly affects the amount o f NOC sorption that occurs in a soil
(Chiou, 1989).
11
.
Chemical sorption to soils is expressed as a partition coefficient (Kd) that relates
the amount o f chemical in the sorbed phase to the concentration in the aqueous phase:
^ sorbed
'aqueous
where Csorbed is the concentration o f sorbed compound (mg kg'1) and Caqueous is the
aqueous concentration o f compound in equilibrium with the sorbed phase (mg L'1). Since
the primary sorptive phase in soil is organic matter, the partition coefficient is commonly
normalized to organic matter, Kom:
Kd
Y
^z i som
where Xsom is the fraction o f the soil composed o f organic matter. To account for
structural and chemical variations in SOM, the partition coefficient can also be expressed
in terms o f soil organic carbon, (Koc):
where Xoc is the fraction o f organic carbon in the soil. Note that in all three expressions,
larger partition coefficients correspond to greater sorption to soil (Chiou, 1989).
Mechanisms o f Sorption to Soil
Two theories exist as to how organic molecules are sorbed to soil. The first view
maintains that the process is governed by sorption to discrete sites on solid surfaces
located in micropores within the soil matrix. The second theory holds that NOCs
partition into the physical matrix of SOM and are distributed throughout the volume
(Alexander, 1994). Sorption governed by the first mechanism is limited by a fixed
12
number o f sorption sites and exhibits, a plateau effect at high concentrations. Chiou
(1989), however, presented strong evidence in favor o f the partitioning mechanism. He
noted that NOC isotherms showed strong linearity over a large range o f concentrations
and that sorption was strongly dependent on the quantity o f SOM present. Most likely,
both mechanisms contribute to NOC sorption to SOM (Pignatello and Xing, 1996). The
dominating mechanism, however, will influence the thermodynamics and kinetics o f
contaminant removal from the sorbed state.
Sorption o f NOCs to soil generally occurs as a two step process. First, a period o f
rapid, reversible NOC sorption to the soil occurs, which is driven by surface sorption.
This is followed by a second period o f slow sorption whereby the contaminant partitions
into the solid matrix. Over long contact times, the period o f slow sorption can result in
Kd values 30% greater than those obtained after a short equilibration time (Pignatello and
Xing, 1996). Historically contaminated (aged) soils where NOCs have been present for
months or years can be enriched in the slow fraction (Pu et al., 1994; Pignatello, 1989);
consequently, aged chemicals are observed to be highly resistant to degradation as
compared to freshly added chemicals (Hatzinger and Alexander, 1995; Steinberg et al.,
1987; Fu et al., 1994). Poor understanding o f sorption and desorption kinetics and their
impact on bioavailability can result in poor planning o f bioremediation schemes.
(
Sorption and Bioavailability
Reduced bioavailability due to contaminant sorption does not necessarily preclude
degradation, but does severely limit the rate and extent o f contaminant transformation.
Several factors may be responsible for reduced degradation rates o f sorbed compounds by
13
soil organisms: (I) desorption kinetics are slow enough that mass-transfer essentially
limits degradation; (2) sorbed compounds cannot enter the cell and be acted on by
intracellular enzymes; (3.) extracellular enzymes are subject to sorption and lose catalytic
activity; (4) sorbed compounds may be less available and accessible to extracellular
enzymatic attack; (5) additional growth factors and nutrients may be sorbed and less
available for microbial growth and activity; (6) near surface conditions may be depleted
in nutrients or have a lowered pH, making it unfavorable for microbial growth and
survival; and (7) the microbes themselves are attached and consequently have limited
access to the contaminant molecules (Alexander, 1994).
Much discussion has centered on the availability o f sorbed molecules for
microbial degradation. One line o f reasoning contends that organisms use only chemicals
that are in solution; consequently, degradation is rate-limited by desolation kinetics.
Furthermore, some microorganisms excrete extracellular metabolites that facilitate
desorption and degradation. Guerin and Boyd (1992) examined the degradation of
naphthalene sorbed to soil using two naphthalene-degrading organisms and observed both
processes. One organism utilized naphthalene at the rate the compound desorbed, and
degradation was limited by substrate solubility. The rate and extent o f naphthalene
degradation by a second organism, however, exceeded predictions made by desorption
kinetics alone, indicating that the organism was enhancing desorption rates or was
utilizing sorbed substrate. Another theory suggests that bacteria attached to surfaces
degrade sorbed substrates and that the proximity between organism and substrate is
responsible for degradation, Harms and Zehnder (1995) observed that the bioavailability
14
and degradation o f 3-chlorodibenzofuran sorbed to teflon increased with bacterial
attachment to the solid surface, suggesting that organism attachment facilitated
contaminant availability.
.
The Role o f Surfactants in Remediation
Surfactant Basics
Surfactants (SURFace ACTive AgeNTS) have received considerable attention for
their potential to increase the aqueous solubility o f sorbed contaminants and potentially
enhance bioavailability in soil environments. Surfactants are amphipathic molecules,
containing both hydrophilic and hydrophobic moieties; they tend to migrate to surfaces
and interfaces or form new molecular surfaces by forming micelles. In aqueous solution,
surfactants cause a reduction in surface and interfacial tensions by the interaction between
the surfactant and the H-bonding o f water molecules. These properties have led to
widespread production and usage o f surfactants as solubilizing agents in a variety of
applications.
Surfactants can be classified into four categories based upon the charge o f their
polar moiety: (I) cationic, containing a positive charge, (2) anionic, containing a negative
charge, (3) zwitterionic, containing both a positive and negative charge, and (4) nonionic,
containing no charged functional groups (Figure 5). An important parameter of
surfactants is the hydrophile-lipophile balance number (HLB). This parameter is a ratio
between the molecular weight o f the polar portion o f the molecule and the molecular
15
weight o f the entire structure; low HLB numbers are indicative o f surfactants with high
hydrophobicity (Rosen, 1989).
A primary solution characteristic o f surfactants is their ability to form aggregates
called micelles. Micelles are a spherical arrangement o f surfactant monomers formed by
A nionic
Nonionic
HO(C2H4O)xx
^O C 2H4XvOH
L g J ^ H ( O C 2H4)vOH
C H 3- ( C H 2)10- C H 2- O - S O 3'
vC (OC2H4)zR
H
Zw itterionic
C ationic
N - C H 2-C H 2- C O O '
N - ( C 2H4O)x - H 2
C H 37
R
Figure 5. Surfactants can be classified into four groups based upon the charge of the polar
group— nonionic, anionic, cationic, and zwitterionic—containing no charge, a negative charge, a
positive charge, and both positive and negative charges, respectively.
the association o f the hydrophobic groups, creating a nonpolar pseudophase in a
hydrophilic shell. Surfactants with low HLB numbers form micelles with large
hydrophobic cavities as compared to the polar exterior. The concentration at which
micelles form is known as the critical micelle concentration (CMC). At concentrations
below the CMC, surfactants exist in solution primarily as individual monomers;
surfactant monomers also associate with surfaces, forming structures known as
hemimicelles. As the concentration o f monomers increases up to the CMC, the surface
16
tension o f the solution decreases. At concentrations above the CMC, a constant monomer
concentration is maintained in solution by equilibrium with the micelles and
hemimicelles, and the surface tension o f the solution becomes relatively constant.
Another characteristic o f surfactants is their ability, under certain conditions, to
form emulsions between two immiscible liquids. A sudden drop in interfacial tensions
that can accompany the addition o f surfactants in a two-phase liquid system results in the
formation o f fine droplets o f one liquid in another (e.g. oil in water) (Rosen, 1989; Rouse
et al., 1994). Emulsions can play an important role in the degradation o f immiscible
liquids (i.e. oil in water) by increasing the surface area o f the immiscible phase.
Solubility Enhancement of
NOCs in Aqueous Solutions
The aqueous solubility o f NOCs can be enhanced by the addition o f surfactants
above the CMC. The hydrophobic core o f a micelle acts as a partitioning phase for
compounds with low water solubility. Kile and Chiou (1989) observed enhanced
solubility o f both I, I -bis(p-chlorophenyl)-2,2,2,-trichloroethane (DDT) and
trichlorobenzene (TCB) in six commercial surfactant solutions above the CMC. The
solubility enhancement o f the more insoluble compound, DDT, was more pronounced
than that observed for TCB. The effect was also seen to vary with surfactant structure.
With the nonionic surfactants studied, the solubility o f NOCs increased proportionally
with the length o f the nonpolar chain content. Moreover, the solubility o f DDT also
increased at concentrations below the CM C-a phenomenon that has been observed for
only compounds with extremely low water solubility-presumably due to interactions
17
between DDT and the surfactant monomers. However, increases in NOC solubility are
normally directly related to micelle concentration. A linear relationship was observed
" between naphthalene, phenanthrene, and pyrene solubility and surfactant concentration in
aqueous solution (Edwards et al., 1991). Furthermore, surfactants increase the rate of
NOC solubilitzation. Grimberg et al. (1995) found that the rate o f phenanthrene
dissolution from solid to aqueous phase increased with increasing surfactant
concentration.
Solubility Enhancement o f NOCs in Soil Systems
In soil systems, surfactants have been successfully used to solubilize sorbed
contaminants (Figure 6). For example, the soil-water partition coefficient (Kd) for
anthracene was reduced by two orders magnitude with a 1% volume addition of
phenylethoxylate surfactant (Liu et al., 1991). Increased solubility can be used to
positively influence contaminant transport during soil washing or flushing.
Tetrachloroethylene (TCE) and dodecane were successfully flushed from soil columns by
the use o f surfactant solutions (Pennell et al., 1994; Pennell et al., 1993). However, the
rate o f contaminant transport for both compounds was rate-limited by desorption.
Evidence suggests that desorption kinetics in the presence o f surfactants are highly soil
dependent. Deitsch and Smith (1995) examined the rate o f TCE dissolution from both an
organic and a mineral soil treated with 3000 mg L'1 Triton X-100. Desorption from the
organic soil was rate-limited whereas solubilization from the mineral soil could be
approximated assuming instantaneous equilibrium. Regardless o f soil type, desorption
rates are higher with the addition of surfactant than in the absence o f any solubilizing
18
agent. Deitsch and Smith (1995) observed increased contaminant desorption rates from
an organic soil treated with surfactant relative to a no surfactant control.
Contaminant desorption from soils is hypothesized to occur by two mechanisms:
( I) micellular solubilization that is thermodynamically driven by the concentration
Bound Contaminant
Free Contaminant
Surfactant
Micelle / Emulsion
o°’
Surfactant Monomer
Fungi
Bacteria
o
O o
Partitioning
Figure 6. Interactions between sorbed contaminants, soil bacteria and fungi, and surfactants in
enhancing solubility of NOCs in soil environments.
gradient between sorbed and soluble states or (2) sorption and penetration of surfactant
molecules to SOM which causes the organic matrix to swell, permitting a greater flux of
contaminant from the sorbed to aqueous phase. By use of a radial diffusion model, Yeom
et al. (1996) showed that surfactant enhanced PAH release from an organic rich soil-tar
was driven by increased matrix diffusivities and not from partitioning to the micellular
19
pseudophase, suggesting that the influence o f surfactants on desorption is due to the
second mechanism.
Surfactant introduction into soil-water systems is more complex than pure water
systems. Surfactants sorb to the soil, and the quantity o f surfactant required to achieve
CMC can be significantly higher than that in a pure aqueous system. Liu et al. (1991),
using a soil to solution ratio o f 1:8, required a surfactant addition that was 40 times
greater than the CMC in pure water in order to achieve solubilization o f sorbed
anthracene. Furthermore, surfactant sorption can actually result in decreased solubility o f
NOCs. Sorption o f Triton X-100 to a sandy soil resulted in increased soil-solution
partition coefficients (Kd) for three chlorinated hydrocarbon compounds at surfactant
concentrations below the CMC. Once the CMC was achieved, the apparent water
solubility for all three compounds increased (Sun et al., 1995). Surfactant sorption is o f
major consequence in soil systems and needs to be considered in surfactant-aided
remediation strategies.
Biodegradation o f Organic Contaminants
in Soils Treated with Surfactants
The use o f surfactants in bioremediation strategies has met with mixed success. A
few studies have shown that surfactants can solubilize contaminants and result in
enhanced degradation by microorganisms (Thibault et al., 1996; Aronstein and Paterek,
1995; Tiehm, 1995). For instance, Guerin and Jones (1988) noted that phenanthrene
mineralization by a Mycobacterium was enhanced by the use o f Tween surfactants above
the CMC. The mechanism o f contaminant degradation in micelles is unknown, however,
20
it has been suggested that (i) degradation occurs only after the compound has exited from
the micelle or that (ii) cell-micelle contact (and possible fusion between cell membrane
and micelle) may be responsible (Tiehm, 1995).
Contaminant degradation with surfactants at sub-CMC levels has also been
reported. Aronstein et al. (1991) noted that the degradation o f sorbed phenanthrene and
biphenyl occurred in soils treated with surfactants below the CMC. They suggested that
surfactants facilitated mass-transport from the sorbed state and resulted in increased
degradation although contaminant solubility was not enhanced.
In contrast, many studies have shown that the use o f surfactants inhibits
mineralization o f NOCs in soil systems. Some researchers have suggested that NOCs
partitioned in micelles are unavailable to degrading bacteria (Guha and Jaffe, 1996; Laha
and Luthy, 1991). Others have noted a physiological effect o f surfactants on degrading
organisms. Laha and Luthy (1992) reported that the mineralization o f phenanthrene was
completely inhibited at surfactant concentrations above the CMC, but commenced after
dilution o f surfactant concentrations to sub-CMC levels. Roch and Alexander (1995),
however, noted surfactant toxicity to a phenanthrene degrading Pseudomonas species at
concentrations below the CMC. Tiehm (1994) reported that surfactant toxicity to PAH
degrading bacteria was related to surfactant structure; surfactant toxicity decreased with
increasing hydrophilicity. Another theory for the failure o f surfactants to aid in
degradation is the preferential use of surfactant as a substrate to the degrading organisms.
The exact mechanisms responsible for the inhibition o f contaminant degradation in the
presence o f surfactants are not clear. However, it is likely a number o f factors contribute
21
and that degradation rates o f NOCs are dependent on both the properties o f the surfactant
and organism alike.
Biosurfactants
It has been recognized that microorganisms produce surfactants (biosurfactants) as
a mechanism for increasing accessibility to bound subtrates (Rouse et al., 1994; Deziel et
al., 1996; Van Dyke et al., 1993). It has, been suggested that biologically-produced
surfactants may be more effective in remediation strategies for several reasons: (I) unlike
many synthetic surfactants, biosurfactants are degraded by microorganisms and may
stimulate activity; (2) biosurfactants may be less toxic to organisms; and (3)
biosurfactants may be more effective for specific degradation processes in natural
<
environments.
Biosurfactant-producing bacteria are common in soil environments. Deziel et al.
(1996) isolated 23 bacteria from a petroleum-contaminated soil that produced
biosurfactants while growing on mineral salt agar using naphthalene and phenanthrene as
sole C substrates. One o f those isolated organisms, a strain o f Pseudomonas aeruginosa,
increased the water solubility o f naphthalene when grown in liquid media-. Biosurfactants
have been used to increase the solubility and transport o f contaminants in soils.
Cyclodextrins, cyclic oligosaccharides produced by the enzymatic degradation o f starch
by bacteria, have shown solubilization potential. A cyclodextrin derivative,
hydroxypropyl-P-cyclodextrin (HCPD), increased the solubility o f six NOCs in solution,
with the greatest solubility enhancements occurring with the more insoluble compounds
(Wang and Brusseu, 1993). HPCD was also successfully used to flush NOCs through
22
two soils (Brusseau et al., 1994); a 10 g L-1HPCD solution was calculated to be 100
times more effective at facilitating pyrene transport through a soil column than water
alone.
Bioremediation strategies involving biosurfactants have been well illustrated by
research conducted using rhamnolipids. Rhamnolipids produced by Pseudomonas
aeruginosa strain UG2 (isolated from an oil-contaminated soil) are produced in relatively
large quantities and effectively solubilize NOCs in pure solution as well as in soil (Van
Dyke et al., 1993; Scheibenbogen et al., 1994). The effectiveness o f the rhamnolipids to
solubilize NOCs in soils was affected by soil type, contaminant aging, and biosurfactant
sorption to soil (Van Dyke et al., 1993).
One remediation approach is to introduce the biosurfactant directly into
contaminated soil. Provident! et al. (1995) introduced partially-purified rhamnolipids into
a phenanthrene contaminated soil and observed increased mineralization by indigenous
microorganisms. When the biosurfactant was introduced with a phenanthrene-degrading
organism, mineralization was inhibited, perhaps due to preferential use o f the surfactant
by the degrading species.
A second remediation strategy is to introduce surfactant-producing organisms into
the environment. In the same study by Provident! et al. (1995), inoculation with
Pseudomonas aeruginosa UG2 alone did not improve mineralization. However, when
the soil was inoculated with both the surfactant-producing organism as well as a
phenanthrene-degrading organism, enhanced mineralization occurred. It was suggested
that inoculation with biosurfactant-producing bacteria alone resulted in the proliferation
23
o f organisms that preferentially used the rhamnlipids as a C-source and therefore
displaced the phenanthrene utilizing species, resulting in decreased degradation.
In summary, the use o f biosurfactants in bioremediation is in its infancy. It is
clear that interactions among biosurfactants, organisms, soils, and pollutants are complex
and not completely understood. Further research to clarify surfactant-microbial
interactions may provide answers needed to evaluate the potential o f biosurfactants in
bioremediation.
Objectives
Given the extent and number o f sites contaminated with NOCs, there is a need to
better define the processes affecting contaminant degradation in soils. Of specific interest
locally is the remediation o f two compounds o f environmental concern around the state o f
Montana, 2,4-dichlorophenoxyacetic acid (2,4-D) and pentachlorophenol (POP). The
current research focused on the degradation o f these compounds in Montana soils.
For 2,4-D the primary objectives were to: (I) compare degradation rates in 31
soils from around the state with widely different SOM contents, (2) determine the
applicability o f three kinetic models for describing observed degradation, and (3)
investigate the relationships between 2,4-D sorption to SOM, bioavailability, and
biodegradation rates.
With respect to POP, the objectives were to: (I) compare the abilities of
indigenous bacteria and white-rot fungi to degrade PCP in contaminated soil, (2)
investigate the effect of surfactant type (i.e. synthetic or naturally-derived) on degradation
24
by white-rot fungi, and (3) examine the potential mechanisms involved in surfactantaided PCP degradation by white-rot fungi.
A common goal in both these studies was to examine the impact o f contaminant
availability on the degradation o f environmentally significant pollutants by bacteria and
fungi in soil.
25
CHAPTER 2
BIODEGRADATION OF 2,4-DICHLOROPHENOXYACETIC ACID (2,4-D) IN
THIRTY-ONE AGRICULTURAL SOILS
2,4-dichlorophenoxyacetic acid (2,4-D) is a widely used herbicide for the control
o f broadleaf weeds and is a representative member o f a large class o f chlorinated
pesticides. Although it is generally assumed that 2,4-D degradation rates are fairly rapid
in surface soils (half-life: 7-10 d), 2,4-D has been detected at low concentrations in
groundwaters around the state o f Montana. The primary goal o f the current study was to
determine relationships among 2,4-D degradation rate parameters, soil organic C (SOC)
contents, and 2,4-D sorption coefficients using mineral soils with a wide range in SOC
contents. The degradation o f I4C-2,4-D was monitored in batch vessels under aerobic
conditions in 3 1 agricultural soils collected in Montana and treated with 2,4-D
concentrations representative o f field application rates. Three kinetic models, including
first-order, modified first-order, and logistic, were fitted to the observed degradation
curves. The modified first-order kinetic model (containing two fitted parameters, the rate
constant, k, and the maximum extent o f degradation, Pmax) best described the observed
production o f 14CO2 from 14C-2,4-D in all soils. The degradation o f 2,4-D across all soils
was characterized by short lag phases (less than I d), a period o f maximum degradation
26
rate, and a wide range in the extents o f degradation (Pmax values ranging from 0.2-0.8).
Fitted rate constants were not correlated with SOC contents, nor with 2,4-D sorption
coefficients, indicating that the extent o f 2,4-D degradation may be limited by sorption to
the soil organic matter (SOM) phase. Estimates o f the total number o f soil
microorganisms culturable on non-selective media did not correlate strongly with fitted
rate constants or values O fPmax. This study suggests that rate constants taken by
themselves (as if often the case for first-order modeling applications) are poor predictors
o f the fate o f 2,4-D in soils due to the fact that extents o f degradation are often limited by
sorption to the SOM phase.
Introduction
2,4-dichlorophenoxyacetic acid (2,4-D) was the first commercial herbicide
introduced in the United States in 1945. In states such as Montana, where significant
acreage is used for cereal grain production, 2,4-D remains one o f the most commonly
used herbicides for broadleaf weed control. Literature values o f the half-life (t,#) and
sorption coefficient (Koc) o f pesticides are often used as input data in fate and transport
models (e.g. LEACHM, Wagenet and Hutson, 1987) for predicting groundwater
contamination potential. 2,4-D normally exhibits low sorption coefficients (Koc = 2.00 x
IO"2 Hi3Lg'1) and relatively short half-lives in soil (ft/2 = 7.1 days; Jury et al., 1987) and is
predicted to be o f low pollution potential with respect to many other commonly used
pesticides (primarily due to rapid degradation rates). However, many well monitoring
surveys, including testing by the Environmental Management Division o f the Montana
27
Department o f Agriculture, have identified 2,4-D as a common contaminant o f regional
groundwaters (DeLuca, 1989).
The degradation o f 2,4-D in soils has been shown to depend on numerous soil and
environmental properties, including soil moisture, temperature, microbial biomass, soil
organic C, soil depth, pore water velocity, and 2,4-D concentrations (Sandmann et al.,
1988). Soil bacteria and fungi have both been reported to degrade 2,4-D (Donnelly et al.,
1993; Yadav and Reddy, 1993; Stott et al., 1983). Although more work has been
conducted to characterize bacteria responsible for 2,4-D degradation in soil, soil
degradation studies generally do not differentiate between fungal and bacterial
degradation. Greer et al. (1990) found that the duration o f lag phases and times required
for complete degradation in soil were linearly dependent on the concentration o f 2,4-D
and bacterial population density. Other researchers have also shown that the length o f the
lag phase increases with increasing 2,4-D concentrations in soil (Veeh et al., 1996; Parker
and Doxtander, 1983; Ou, 1978). Rates o f 2,4-D degradation also vary with soil organic
matter (SOM) contents (Ogram et al., 1985). Veeh et al. (1996) noted increased rates o f
2,4-D degradation in surface soils with greater SOM than in low organic matter
subsurface soils. However, Greer et al. (1992) found that microbial 2,4-D degradation in
a high organic matter soil (14%) was significantly lower than in a low organic matter soil
(1.6%). Ogram et al. (1985) used three mathematical models relating 2,4-D degradation
to substrate availability and concluded that sorbed 2,4-D was unavailable for microbial
degradation, which suggests that higher SOM contents should result in lower extents o f
mineralization.
28
Numerous kinetic models have been used to empirically describe degradation
rates o f 2,4-D in soils, including first-order, modified first-order, and logistic expressions.
Although the simplicity o f the first-order model is convenient for estimating half-life
values and for coupling degradation kinetics to transport models, it often fails to describe
experimentally determined degradation rates in soils (Alexander and Scow, 1989).
Microbial growth models such as the Monod and logistic expressions have also been used
to empirically describe 2,4-D degradation in soils (Veeh et al., 1996; Estrella et al.,
1993). The logistic growth model is particularly robust (three fitted parameters) and is
useful for describing the lag period often observed prior to the onset o f maximum
degradation rates and maximum extent o f degradation. Although microbial growth
models often adequately describe degradation rates o f organic contaminants in soils (due
primarily to a greater number o f fitted parameters) they do not necessarily reflect the
complex soil processes controlling degradation kinetics. For example, the presence o f
soil aggregates and SOM influences mass transfer rates o f organic chemicals to the
aqueous phase and plays an important role in the kinetics o f microbial degradation.
Models based on microbial growth and/or simple first-order kinetics do not account for
these effects (Scow and Hutson, 1992). Scow and Alexander (1992) showed that both the
rate and extent o f phenol degradation were lower in the presence o f aggregates,
demonstrating the importance o f mass-transfer and diffusion processes in soil kinetic
models.
The objectives o f the current study were to (I) determine the rates and extents o f
2,4-D degradation in 3 1 agricultural soils exhibiting a large range in SOM contents, and
(2) evaluate the relationship among SOM contents, 2,4-D sorption coefficients (Kd), and
29
degradation rate parameters obtained from several empirical models used to describe
kinetics o f 2,4-D degradation. Although we are aware that 2,4-D degradation rates have
been extensively studied in soil (Sandmann et al., 1988; Ogram et ah, 1985; Greer et ah,
1992), one o f the unique features o f the current study was the large number o f soils used
(31), which exhibited a wide range in SOM contents.
Materials and Methods
Chemicals
14C-Uniformly ring-labeled and 14C-Carboxyl-Iabeled 2,4-D were purchased from
Sigma Chemical (St. Louis, MO). Radiopurity was verified by high-pressure liquid
chromatography (HPLC) on a C-18 reverse phase column at 254 nm using acetonitrile0.2% phosphoric acid (25:75) with a flow rate o f 1.5 mL min'1 (Rf = 6.5 min). 14C was
analyzed with a Beckman 171 Radioisotope detector (Beckman Instruments, Inc.,
Fullerton, CA).
Soils
Thirty-one agricultural soils from across Montana (Table I) were collected from
the surface horizons (top 30 cm) and stored field moist at 4°C until use. The majority o f
these soils were collected as pairs from adjacent locations representing the same soil
series, but differing in management history: one soil representing a native range site with
presumably no previous application history o f 2,4-D, the other from a crop-fallow site
which had been cultivated for nearly 50 y and had possibly seen several applications o f
2,4-D (Neill, 1994).
Table I . Soil series, locations, land-use specifications, texture, pH, EC, and soil organic carbon for the 31 soils used in the study.
Soil Code
AMST
BACF
BANK
BECF
BENR
BROC
CRCF
CRNR
HICF
HINR
INSK
LACF
LANR
LONN
NECF
PlCF
PlN R
P2CF
P2NR
P3CF
P3NR
* dS m"1
Soil Series
Location
Gallatin
Fergus
Fergus
Daniels
Daniels
Gallatin
Fergus
Fergus
Fergus
Danvers
Fergus
Danvers
Gallatin
Beaverell
Daniels
Turner
Daniels
Turner
Custer
Lonna
Judith Basin
Winiffied
Tumer-Beaverton Roosevelt
Tumer-Beaverton Roosevelt
Roosevelt
Tally
Roosevelt
Tally
Roosevelt
Famuf
Roosevelt
Famuf
Amsterdam
Famuf
Famuf
Farland-Cherry
Farland-Cherry
Brocko
Fairfield-Danvers
Fairfield-Danvers
Land Use
Crop-Fallow
Crop-Fallow
Native-Range
Crop-Fallow
Native-Range
Crop-Fallow
Crop-Fallow
Native-Range
Crop-Fallow
Native-Range
Pasture
Crop-Fallow
Native-Range
Native-Range
Crop-Fallow
Crop-Fallow
Native-Range
Crop-Fallow
Native-Range
Crop-Fallow
Native-Range
-------- %--------
Textural
2 il
Sand Silt Clay
Class
pH EC*
15
28
26
46
45
24
34
34
30
39
37
62
65
16
30
81
81
70
70
69
67
52
44
46
32
34
56
30.
32
36
32
36
19
20
54
32
8
8
15
16
17
17
33 Silt Loam
28 ;Clay Loam
28 Clay Loam
22 Silt Loam
21 Silt Loam
20 Silt Loam
36 Clay Loam
34 Clay Loam
34 Clay Loam
29 Clay Loam
27 Loam
19 Sandy Loam
15 Sandy Loam
30 Silty Clay Loam
38 Clay Loam
11 Loamy Sand
11 Loamy Sand
15 Sandy Loam
14 Sandy Loam
14 Sandy. Loam
16 Sandy Loam
6.9
7.1
7.5
7.8
7.5
8.1
7.3
7.2
5.4
5.3
7.3
6.0
6.4
8.0
7.1
6.4
6.4
7.0
6.4
5.5
6.1
Organic •
C(%)
(L k g 1)
0.26
0.38
0.53
0.24
0.66
NA
0.22
0.28
0.25
0.19
0.58
0.15
0.12
0.28
0.47
0.07
0.50
2.27
2.53
0.97
1.76
0.03
1.34
1.23
1.79
1.83
2.49
0.83
0.93
1.63
2.21
0.64
0.09
0.10
0.14
0.14
0.12
1.17
0.68
0.90
0.61
0.74
0.64
1.52
1.62
0.22
0.37
0.11
0.51
0.62
1.00
0.56
0.88
0.19
0.17
0.72
1.68
0.19
0.17
0.35
0.24
0.12
0.17
Table I. --Continued
Textural
2:1
Clay
Class
PH
Organic
Kd
C
(L kg'1)
Soil Code
Soil Series
Location .
Land Use
Sand
P4CF
Williams-Zahill
Roosevelt
Crop-Fallow
65
17
18
Sandy Loam
5.8 • 0.15
0.26
0.06
P4NR
Williams-Zahill
Roosevelt
Native-Range
63
20
17
Sandy Loam
6.3
0.12
0.38
0.09
POCF
Judith
Judith Basin
Crop-Fallow
29
32
39
Clay Loam
7.2
0.16
1.45
0.49
PONR
Judith
Judith Basin
Native-Range
33
29
38
Clay Loam
7.2
0.21
1.99
0.66
PUCF
Farland-Cherry
Daniels
Crop-Fallow
43
32
25
Loam
7.4
0.21
1.53
0.50
PUNR
Farland-Cherry
Daniels
Native-Range
48
34
18
Loam
7.7
0.33
2.68
1.06
SONN
Sonnet
Custer
' Native-Range
6
57
37
Silty Clay Loam
7.6
0.18
2.11
0.68
SUCF
Willams-Zahill
Daniels
Crop-Fallow
25
33
42
Clay
8.0
0.24
1.21
0.34
SUNR
Willams-Zahill
Daniels
Native-Range
30
36
34
Clay Loam
7.9
0.25
1.79
0.33
WILL
Williams
Custer
Native-Range
20
50
30
Clay Loam
5.6
0.33
3.72
1.16
EC
,
32
Batch Biodegradation
Fifteen g o f sieved, field-moist soil were placed in 25 mL screw-cap erlenmeyer
flasks (in duplicate) equipped with a gas exchange system (Figure I). 14C-Carboxyl-
Figure 7. Experimental setup used in the batch biodegradation studies. Compressed air or
nitrogen was passed through a NaOH trap into a manifold that controlled gas flow to individual
flasks containing l4C-2,4-D labeled soil. Degradation was measured as the accumulation of 14CO2 in
0.5 M NaOH traps changed regularly.
labeled 2,4-D (0.12 pCi) was added to the soil, simulating an application o f 0.2 kg ha'1
incorporated to a depth o f 15 cm. Soil moisture was adjusted to Om= 0.36 (mass basis)
using double-deionized (DD) water. Evolved 14CC^ was trapped in 10 mL 0.5 M NaOH
and analyzed every 12 to 48 h using liquid scintillation analysis (Tri-Carb Liquid
Scintillation Analyzer, Packard Instrument Co.). Soils were not analyzed for metabolites
since previous studies have indicated that accumulation o f metabolites is less than 5-10%
33
o f the added 2,4-D (Ou, 1978; Smith and Aubin, 1991) and considering that removal o f
the carboxyl group is the first step in the degradation pathway (Aislabie and Lloyd-Jones,
1995; Kunc and Rybafova, 1983). At the conclusion o f each experiment, a soil
subsample was analyzed for total 14C by oxidation at 800°C under an O2 atmosphere for
four min (Biological Oxidizer Model OM300, R.J. Harvey Instrument Corp., Hillsdale,
NI). Average total recoveries o f duplicate treatments exceeded 90%.
Batch Sorption
The sorption o f 2,4-D was measured in 25 mL polypropylene centrifuge tubes
using 15 g soil and 15 mL DD water containing an equivalent 14C ring-labeled 2,4-D
concentration as in the biodegradation experiments. Microbial activity was inhibited
using 0.02% sodium azide, verified by the absence o f 14COz evolution and greater than
90% recovery o f total 2,4-D added to sorption treatments. Samples were equilibrated for
48 h on a table shaker, centrifuged, the supernatant filtered through a 0.45 pm nylon
filter, and analyzed for 14C by liquid scintillation. Blanks containing no soil were used to
verify that no measurable sorption occurred to centrifuge tubes. Sorption was determined
as the difference between total 14C added and 14C in the supernatant.
Bacterial Plate Counts
An estimate o f total bacterial numbers was determined for six o f the 3 1 soils
(bracketing a wide range o f SOC contents, 0.26-3.72%) by plating soil dilutions on Yeast
Extract-Peptone-Glucose media consisting o f 5 g peptone, 5 g yeast extract, 2 g glucose,
15 g bacto agar, and I L water. Ten g o f soil was added to milk dilution bottles
containing 90 mL o f phosphate buffer (1.67 g KH2 PO4 , 1.35 g K2 HPO4 ) adjusted to pH
34
7.0, and shaken on a table shaker for I h. Seven serial 1:10 dilutions were made and 100
pL plated in triplicate. The plates were incubated at 3 O0C for 10 d and colony forming
units (CPUs) determined from the average o f plates containing between 30 and 300
colonies.
Analytical Methods
Three kinetic models were evaluated for their ability to describe 2,4-D
degradation (by non-linear regression). The three models were chosen for their
simplicity and possible application for degradation subroutines in transport models.
First-order kinetics:
- = \ - e kt
C0
where CZC0 is the fraction o f total 14C-2,4-D evolved as 14CO2, k is the first-order rate
constant, and t is time.
Modified first-order kinetics (Ogram et al., 1985):
where P is the fraction o f total 14C-2,4-D evolved as 14CO2, Pmax is the maximum extent
o f mineralization in terms o f total 14C added, k is the rate constant, and t is time.
Logistic Equation (Characklis, 1990):
X =
x
/
where X is the fraction o f 14C-2,4-D evolved as 14CO2, X0 is the initial fraction o f CO2
evolved, X mis the maximum extent o f mineralization, k is the rate constant, and t is time.
35
Results and Discussion
Variation in 2,4-D Degradation Among Soils
Degradation curves for 2,4-D are shown for six o f the 3 1 soils (Figure 8) to
illustrate the short lag phases, the period o f maximum degradation rates, and the wide
SOC (%)
©
Z
©
0.6
0.3
0.2
0
10
20
30
40
50
60
T im e (d ays)
Figure 8. Degradation curves showing the fraction of C-2,4 D evolved as 14CO2from six of the 31
soils. The organic carbon content (SOC, %) of the soil is shown adjacent to each curve.
range in extents o f 2,4-D mineralization (20-80%). Degradation curves for all soils were
typified by relatively short lag phases (< 24 h), and there was no evidence that lag phases
were longer in native range (NR) soils with limited prior exposure to 2,4-D (i.e. no
significant acclimation period by microorganisms prior to the onset o f degradation). The
observed lag phases were shorter than those commonly reported in previous studies. The
36
length o f the lag phase has been shown to increase with increasing initial 2,4-D
concentrations (Greer et al., 1990; Parker and Doxtander, 1983; On, 1978); consequently,
the short lag phases observed in this study may relate to the low initial 2,4-D
concentrations (used to reflect typical 2,4-D application rates).
The first-order kinetic model (one fitted parameter) did not adequately describe
2,4-D degradation in most o f the soils studied (Table 2) and resulted in particularly low r2
values for soils exhibiting higher maximum degradation rates and lower extents o f 2,4-D
degradation (e.g. WILL; Table 2). First-order kinetics assume that all o f the added 2,4-D
is available for degradation. The aqueous 2,4-D concentration in soils with high SOM
can be substantially less than that originally introduced, due to sorption. Consequently,
the first-order model is limited in its capability to describe degradation in soils exhibiting
large amounts o f 2,4-D sorption. The modified first-order equation includes a second
parameter (Pmax term) that empirically accounts for the limited availability o f 2,4-D, and
was more successful in describing the observed degradation kinetics. The logistic
equation includes a third fitted parameter, which is useful in describing the lag phase
commonly observed prior to the onset o f maximum degradation rate. Since only very
short lag phases were observed in the majority o f the soils studied, the logistic equation
did not provide a better description o f 2,4-D degradation kinetics compared to the
modified first-order model.
The logistic and modified first-order expressions resulted in comparable
parameter estimates o f the extent o f 2,4-D degradation (Xmand Pmax) and degradation
rate constants. First-order rate constants were significantly lower than those fitted with
the modified first-order or logistic models, due to the fact that the first-order model does
Table 2. Comparisons among the three models used to describe observed 2,4-D degradation in the 31 soils studied.
Observed
Soil
AMST
BACF
BANK
BECF
BENR
BROC
CRCF
CRNR
HICF
HINR
INSK
LACF
LANR
LONN
NECF
PlCF
PlN R
P2CF
P2NR
P3CF
P3NR
Pmax
0.78
0.66
0.66
0.80
0.72
0.69
0.66
0.75
0.59
0.62 .
0.65
0.79
0.62
0.54
0.62
0.84
0.55
0.81
0.69
0.73
0.67
First-order
k (h"1)
r2
0.005 0.32
0.003 0.00
0.003 0.65
0.005 0.73
0.003 0.65
0.003 0.82
0.003 0.37
0.005 0.27
0.002 0.00
0.003 0.00
0.003 0.10
0.006 0.56
0.002 0.61
0.002 0.64
0.003 0.36
0.014 0.06
0.002 0.00
0.005 0.64
0.003 0.41
0.005 0.19
0.003 0.33
Modified First-order
k (h ') Pmax
r2
0.030 0.74
0.96
0.95
0.020 0.59
0.009
0.011
0.009
0.010
0.013
0.020
0.021
0.021
0.018
0.017
0.009
0.007
0.013
0.039
0.015
0.015
0.013
0.024
0.016
0.61
0.78
0.69
0.66
0.59
0.68
0.52
0.55
0.60
0.74
0.58
0.50
0.58
0.78
0.50
0.75
0.62
0.70
0.63
0.99
0.96
0.99
0.93
0.96
0.96
0.91
0.94
0.97
0.96
0.98
0.98
0.99
0.94
0.95
0.98
0.97
0.94
0.63
Logistic
k(h"‘) Xo
0.163 0.040
0.053 0.081
0.022 0.081'
0.054 - 0.028
0.027 0.073
0.073 0.016
0.030 0.097
0.070 0.058
0.027 0.138
0.049 0.089
0.059 0.056
0.079 0.024
0.029 0.051
0.019 0.064
0.032 0.077
0.139 0.052
0.039 0.070
0.036 0.109
0.036 0.078
0.114 0.019
0.049 0.069
Xm
0.73
0.58
0.58
.0.74
0.66
0.63
0.57
0.66
0.52
0.54
0.58
0.72
0.64
0.46
0.56
0.77
0.49
0.73
0.59
0.68
0.61
r2
0.96
0.91
0.95
0.97
0.96
0.97
0.91
0.93
0.86
0.89
0.94
0.97
0.97
0.94
0.95
0.93
0.91
0.94
0.94
0.96
0.93
Table 2. Continued
First-order
Logistic
Modified First-order
Soil
Observed
Pmax
k(h-')
r2
k(h-')
Pmax
r2
k(h-')
X0
Xm
r2
P4CF
0.75
0.002
0.96
0.003
0.79
0.99
0.011
0.070
0.70
0.99
P4NR
0.71
0.002
0.64
0.007
0.63
0.98
0.012
0.129
0.65
0.94
POCF
0.71
0.004
0.00
0.024
0.65
0.93
0.071
0.068
0.63
0.90
PONR
0.56
0.002
0.17
0.010
0.51
0.93
0.014
0.129
0.50
0.88
PUCF
0.68
0.003
0.49
0.010
0.66
0.98
0.034
0.058
0.62
0.96
PUNR
0.51
0.001
0.49
0.007
0.50
0.98
0.017
0.071
0.47
0.94
SONN
0.40
0.001
0.22
0.010
0.35
0.96
0.024
0.057
0.34
0.90
SUCF
0.61
0.002
0.67
0.006
0.59
0.97
0.024
0.043
0.54
0.96
SUNR
0.53
0.001
0.72
0.005
0.54 . 0.98
0.032
0.032
0.48
0.98
WILL
0.25
0.001
0.00
0.033
0.23
0.058
0.048
0.23
0.88
' 0.93
39
not compensate for extents o f degradation that are less than one. Consequently, first.order rate constants would not be accurate predicators o f the long-term fate o f 2,4-D in
these soils. Both the modified first-order and logistic models adequately fitted the extent
o f degradation (Pmax) obtained from the plateau region o f the 2,4-D degradation curves.
The extent o f degradation is probably a function o f both the amount o f 2,4-D initially
available for degradation, and the amount o f 2,4-D that diffuses to sites o f microbial
activity via mass transfer. Consequently, once the fairly labile 2,4-D is utilized,
degradation rates become limited by slow mass transfer rates from the SOM phase
(Pignatello and Xing, 1996). Scow and Hutson (1992) and Scow and Alexander (1992)
have also shown that degradation rates o f phenol were limited by diffusion rates out o f
well-characterized polyacrylamide gel-exclusion beads, indicating that mass transfer rates
are important in describing degradation rate constants and maximum extents o f substrate
degradation.
Relationships Among 2,4-D
Degradation. SOC. and 2.4-D Sorption
Soil organic carbon (SOC) has been shown to negatively correlate with
degradation rates and extents o f nonpolar organic chemicals in soil environments (Greer
et al., 1992; Ogram et al., 1985; Gordon and Millero, 1985). It is well established that
“aging” o f organic contaminants in soils or in SOC phases further reduces effective
desorption rates and bioavailability to degrading microorganisms (Pignatello and Xing,
1996). In the current study, fitted rate constants from neither the modified first-order nor
the logistic models correlated well with SOC (r2 values <0.01; Figure 9). Further, when
soils were grouped as native-range or crop-fallow, no correlation between degradation
40
rate constants and SOC was observed. However, fitted values o f Pmax from the modified
first-order model (or Xm in the logistic model) were negatively correlated with SOC (r2 =
0.05
♦
0.04
*7
JS
♦
♦
S 0.03
CS
"tS
♦
C
♦
® 0.02
4J
♦
♦
♦
♦*
+ %
♦
a
0.01
♦
►
♦
♦
»
♦
♦
♦
♦
0.00
0.9
o 0.8
r
■3 0.7
♦
♦
♦
♦ ♦
egra
O
ON
♦
*
"O
"I
O
a-
♦
♦ *
*7 0.4
U
°-3
\
♦
g 0.2
—< O
O O
ix g
y = -0.057x + 0.78
R2 = 0.54
0
1
2
3
4
Soil organic carbon (% )
Figure 9. Top figure: Comparison of fitted 2,4-D degradation rate constant (h"1) from the
modified first-order model to SOC. Bottom figure: Maximum extent of degradation fitted from
the modified first-order model compared to SOC.
41
0.54, Figure 9), suggesting that sorption o f 2,4-D in the SOC phase limited
bioavailability. The relationship between extent o f degradation and SOC was much
stronger among the crop-fallow soils (r2 = 0.65) than for native range soils (r2 = 0.48). It
is uncertain whether this is just coincidental or whether it may relate to changes in
characteristics o f the SOM phase after long-term cultivation. Most importantly, among
all soils, no correlation was observed between fitted rate constants and values OfPmax (or
Xmin the logistic model; Figure 10), indicating that the rate constant taken by itself
0.05
0.04
Tc
0.03
ts
e
%0.02
CS
Bi
0.01
0.00
0.0
0.2
0.4
0.6
0.8
1.0
M axim um exten t o f degradation
Figure 10. Comparison of the degradation rate constant (h"1) and Pmai, values obtained from
modified first-order model.
becomes a poor predictor o f the extent o f degradation and the long term fate o f 2,4-D in
soil. This has important consequences for modeling 2,4-D degradation rates in soils, in
that rate constants are often used as the only parameter to describe disappearance of
42
organic compounds; it is clear that this approach would grossly overestimate the
microbial degradation o f 2,4-D in the majority o f soils studied.
We have already indicated that extents o f degradation (Pmax) were negatively
correlated with SOC contents (Figure 9). As expected, the sorption coefficients (Kd) o f
2,4-D across all soils was also correlated with SOC contents (r2 = 0.61, the slope o f this
plot yields an average Koc value o f 400 L kg"1; Figure 11). However, when the extent o f
y = 0.4215x - 0.028
R2 = 0.6066
Sc 1.0
0
1
2
3
4
Soil organ ic carbon (% )
Figure 11. Relationship of 2,4-D partition coefficients (Kd) to soil organic carbon contents for all 31
soils. The slope yields an average Koc value of 422 L kg '.
degradation was plotted against soluble fractions o f 2,4-D (calculated based upon Kd
values obtained from batch experiments) present at the start o f each experiment, only a
weak positive relationship was observed (Figure 12). This plot is essentially the inverse
of a plot OfPmax vs Kd. Interestingly, the relationship between the soluble fraction o f 2,4-
43
~
0.7
X 0.5
y = 0.37x + 0.36
R2 = 0.24
~
0.3
Soluble fraction 14C -2 ,4 -D
Figure 12. Maximum extent of l4C-2,4-D degradation plotted against soluble 2,4-D in solution at
t = 0 (calculated from Kd). Solid line is the observed relationship between available and degraded
2,4-D.
D present at time 0 (from the Kd values) and extents o f degradation was considerably
poorer than the relationship OfPmax vs SOC contents. Although we expect that Kd values
or soluble fraction of 2,4-D are indeed important in defining the extent o f degradation,
the Kd parameter does not account for mass transfer rates o f slow desorption, which may
be the more important mechanistic explanation for the strong correlation between Pmax
and SOC. Although Kd values are reflective o f SOC contents and probably mass transfer
rates o f 2,4-D out o f the sorbing phase as well, they do a poorer job describing the
variation in Pmax values than SOC contents. A more appropriate parameter for explaining
extents of degradation across all soils may be the effective desorption rate constant (e.g.
mass transfer rate). We would expect that this parameter would be highly correlated with
44
SOC contents, and may explain why we observed better correlation between SOC and
extents o f 2,4-D degradation.
Relationships between Degradation
and Soil Bacteria Numbers
Veeh efal. (1996) showed that 2,4-D degradation rates were positively correlated
with SOC among different soil depths o f the same soil profile. It was suggested that
higher OC contents were simply reflective o f higher microbial numbers and activity,
resulting in higher degradation rates for surface soils compared to subsoils (B and C
horizons). Other investigations have shown that the size o f the 2,4-D-degrading
population affects degradation rate. For instance, Cullimore (1981) found that increased
degradation rates corresponded to increases in bacteria numbers in five soils, determined
through a most probable number technique. In the current study, CFU estimates for six
o f the studied soils showed a weak positive correlation with SOC as has been previously
reported. However, no relationships were observed between bacterial plate counts and
maximum extent o f 2,4-D degradation or degradation rate constants (Figure 13). Relative
microbial estimates in soils based upon bacterial growth on non-selective media appear to
be a poor indicator for 2,4-D degradation rates.
Moisture Effects on 2,4-D Degradation
Parker and Doxtader (1983) reported variations in the microbial degradation o f
2.4- D under different soil moisture tensions. They reported greater rates and extents o f
2.4- D degradation in soils with lower soil moisture tension (or higher water contents).
Each o f the 31 soils in. the current study was brought to the same moisture content
45
I 0.8
y = -0.05x + 0.95
R! = 0.34
0.7
S 0.6
0.5
5 0.4
E 0.3
Z 0.1
0.035
0.030
y = 0.0002x - 0.0013
R1 = 0.12
C 0.025
S 0.015
06 0.010
0.000
y = 0.0608* - 1.4859
R2 = 0.4065
Colony forming units p erg soil (10s)
Figure 13. Relationship between fitted extent of 2,4-D degradation (Pmax), rate constants, and
SOC to the number of colony forming units of bacteria isolated on non-selective media six soils
(bracketing the range in SOC contents observed for the soils studied).
gravimetrically. To insure that large variations in observed 2,4-D degradation were not
attributable to moisture effects, two soils were treated at five moisture contents. The rate
46
and extent o f 2,4-D degradation in both soils was relatively insensitive to moisture
variations ranging from 20-40% (Figure 14) although small decreases in the extent o f
mineralization were observed at 0m= 0.22, where low water contents may have reduced
Amsterdam Soil
-*-0.55
Brocko Soil
Time (hrs)
Figure 14. Effect of moisture content on the degradation of 14C-ul-2,4-D in two soils. Moisture
content of 0.55 is above saturation; 0.36 is approximately field capacity.
47
2.4- D mass transfer rates to sites o f microbial activity. Both the Brocko and Amsterdam
Soils exhibited slower rates o f degradation at 6m= 0.55 (hyper-saturated conditions), than
under unsaturated conditions (Figure 14). However, the extent o f 2,4-D mineralization
under saturated conditions was not significantly affected after 40 d o f incubation. In
summary, small variations in soil water contents among soils would not be expected to be
responsible for the differences in 2,4-D degradation observed in the current study.
Anaerobic versus Aerobic Degradation
Finally, a brief study on the impact o f anaerobic conditions on ring cleavage and
2.4- D mineralization was undertaken. Reductive dechlorination o f halogenated
pesticides such as 2,4-D has been reported to occur under anaerobic conditions (Kuhn
and Suflita, 1989). Dechlorination of aromatic rings can act as a detoxification step in
pesticide degradation and help to facilitate ring cleavage. Figure 15 illustrates that the
incubation o f 14C-UL-Hng labeled 2,4-D under a N 2 atmosphere showed almost no 14COi
production as compared with soils under aerobic conditions. When the system was
changed to an aerobic environment, very little variation in extent o f degradation was
observed from that under aerobic conditions. Pretreatment o f 2,4-D with incubation
under N i for 1000 hours followed by aerobic conditions did not appear to appreciable
change 14COi production from uniformly ring labeled 2,4-D as compared to purely
aerobic conditions (Figure 15).
48
Amsterdam
Brocko
U
0.20
r
o . i o
Anaerobic
Q
Aerobic
0.30
Amsterdam
U
Brocko
0.20
0
500
1000
1500
2000
T im e (hrs)
Figure 15. Effect of soil preconditioning under anaerobic conditions on the mineralization of 2,4-D
in Amsterdam and Brocko soils (Om= 0.46). A: CO; evolution from ^C uI ring-IabeIed 2,4-D
under aerobic conditions; B: 14CO2 evolution from '4C uI ring-IabeIed 2,4-D preconditioned under
anaerobic conditions, then exposed to aerobic conditions.
49
Summary
The biodegradation o f 2,4-D showed wide variation among 3 1 agricultural soils.
Modeling 2,4-D degradation with first-order kinetics yielded poor predictions of
degradation behavior and pesticide half-life in soil. Accounting for pesticide availability
through the use o f the modified first-order equation or logistic equation yielded more
accurate predictions o f both the rate and extent o f degradation. SOC appears to be the
dominant influence On bioavailability and degradation, as sorbed 2,4-D becomes less
available for microbial degradation. It has been suggested that soluble 2,4-D is most
easily degraded; however, solubility is controlled by both the extent o f sorption and the
mass transfer rates o f desorption from SOC. Most importantly, the use o f Kd values and
degradation rate constants alone could not accurately predict the degradation and
persistence o f 2,4-D in these soils. Pesticide degradation in soil environments is
governed by a number o f soil and environmental parameters that make degradation
predictions difficult and may require that soil specific evaluations be made on a site by
site basis.
50
CHAPTER 3
MINERALIZATION OF PENTACHLOROPHENOL (PCP) IN SOIL BY
WHITE-ROT FUNGI IN THE PRESENCE OF SURFACTANTS
Pentachlorophenol (PCP) is a common and persistent contaminant o f soils at
wood-preserving facilities. White-rot fungi have been identified as possessing enzyme
systems capable o f degrading PCP; however, degradation is often limited by contaminant
sorption to soils. Surfactants can be used to minimize this limitation by increasing the
bioavailability o f sorbed hydrophobic molecules in soils. Phanerochaete chrysosporium
and a non-Phanerochaete fungus were evaluated for their abilities to degrade 14C-PCP
applied to an agricultural soil amended with 5 surfactants (lignosulfonic acid, sodium
salt; lignosulfonic acid, sodium salt, acetate; hydroxypropyl-(3-cyclodextrin; Tween 80;
and Triton X-100). Each surfactant was applied to the contaminated soil at 25 mg g
soil'1, yielding additions o f C as surfactant similar to native soil organic C contents.
Mineralization was measured as the total amount o f 14CC^ recovered from soil columns
flushed with air. In four weeks, the non-Phanerochaete fungus mineralized a greater
amount o f PCP than either P. chrysosporium or the indigenous microflora. The addition
o f Tween 80 at sub-CMC levels resulted in mineralization o f 26% o f the added PCP
versus 17% in a surfactant-free system. Total extractable PCP decreased to less than 10%
51
for both treatments. Enhanced PCP mineralization by the non-Phanerochaete fungus in
the presence o f Tween 80 appeared to be due to increased fungal activity, as indicated by
increased respiration rates and fungal biomass, and not due to solubility enhancement o f
the contaminant. The use o f surfactants, such as Tween 80, in combination with whiterot fungi may represent a strategy for the bioremediation o f residual PCP in contaminated
soils.
Introduction
Pentachlorophenol (PCP) has been extensively used as a fungicide and insecticide
for the preservation o f wood. Its ability to uncouple oxidative phosphorylation has led to
widespread usage in the environment as a pesticide. Large amounts o f chlorophenols,
PCP in particular, have been detected in soils, natural waters, sediments, plant tissues,
and human urine. PCP concentrations in soils at sawmills and wood-preserving facilities
have been reported at concentrations up to several thousand ppm, with the highest levels
associated with dipping basins or storage areas (Boyd, 1989). PCP is extremely toxic to
plants and has been linked to deaths in fish as well as livestock. Human exposure occurs
primarily through the food chain and it has been estimated that humans are exposed to 16
pg day'1 (Etoxnet, 1997; Boyd et ah, 1989; Hattemer-Frey and Travis, 1989).
Commercial formulations often include other highly toxic chemicals such as dioxins,
which are often more toxic than PCP itself. Because o f its extreme toxicity, PCP use is
being phased out in the US and is currently listed as a restricted use pesticide. However,
due to historic use, chlorinated phenols are ubiquitous in the environment and the US
52
EPA has listed PCP along with five other chlorophenols as priority pollutants (Etoxnet,
1997; H aleetal., 1994).
PCP is a weak acid (pKa = 4.35), and its aqueous solubility varies as a function o f
pH. Its solubility is commonly reported to be 14 mg L'1 at pH 7, but can range from 3 mg
L"1 at pH 3 to 10,000 mg L'1 at pH 9 (Arcand et ah, 1995). Both the protonated and
dissociated forms sorb strongly to soil organic matter (SOM) through a partitioning
process whereby the compound diffuses into the organic matrix (log Koc = 4.51).
Incorporation into SOM by oxidative coupling (a covalent bonding process) results in
immobilization, reduced desorption rates, and microbial resistance (Alexander, 1994;
Boyd, 1989). The persistence o f PCP in soil has been reported to range between 14 d and
5 y, depending on environmental conditions (Hale et ah, 1994).
Microbial degradation is often the primary mechanism o f PCP removal from soil.
Microbial degradation can occur under both aerobic and anaerobic conditions. Although
aerobic bacteria capable of degrading PCP have been isolated from the environment,
aerobic degradation is slow (half-lives o f 45 to >72 d have been reported; Etoxnet, 1997;
Boyd, et ah, 1989; Hale et ah, 1994). Anaerobic organisms are more effective degrading
PCP through a process o f reductive dechlorination followed by ring cleavage (half-lives
o f 15 to 30 d have been reported; Etoxnet, 1997; Boyd, et ah, 1989). Bioremediation o f
PCP contaminated sites has shown significant promise; however, an acclimated microbial
population is a necessary precursor for both aerobic and anaerobic degradation to occur.
A number o f investigators have found that the addition o f PCP degrading bacteria to
contaminated soils resulted in increased rates o f degradation.. Edgehill and Finn (1983)
53
found that the addition o f IO6 Arthrobacter cells capable o f degrading PCP per g o f dry
soil reduced the half-life o f the pesticide from two weeks to less than one day. In
contrast, Mueller et al. (1991) found that indigenous soil microorganisms were ineffective
at removing significant quantities o f PCP from coal-tar creosote contaminated
groundwater over two weeks. In general, PCP is toxic to most bacteria at levels far below
those commonly found in contaminated soils and alternate remediation techniques have
been investigated.
Recently, numerous studies have reported that white-rot fungi may be used for the
biodegradation o f PCP in soils. White-rot fungi degrade lignin, the structural component
o f wood, via a free-radical process that is non-selective and non-stereospecific (Barr and
Aust, 1994; Fernando et al., 1994). Two major peroxidases have been identified in
Phanerochaete chrysosporium that are involved in the degradation process: lignin
peroxidase (LiP) and manganese-dependent peroxidase (MnP). These enzymes are
generally expressed under lignolytic conditions where C, N, or S is limiting. The nonselective nature o f these enzymes has led. to the observation that P. chrysosporium can
degrade a number o f environmentally recalcitrant compounds, including: polychlorinated
biphenyls (PCBs), chlorinated phenols (e.g. PCP), polyaromatic hydrocarbons (e.g.
phenanthrene, anthracene), polychlorinated hydrocarbons (e.g. aldrin and lindane),
chlorinated aromatics (e.g. DDT), and munitions wastes (e.g. TNT, RDX) (Fernando et
al., 1994; Khindaria, 1995; Aust, 1990).
The mineralization and transformation o f PCP by white-rot fungi has been well
documented in both liquid culture and soils (Mileski et al., 1988; Lin et al., 1990; Lamar
54
et al., 1990; Lamar et al., 1994; Brodkorb and Legge, 1992; Lamar et al., 1993; Lamar
and Dietrich, 1990; Morgan et al., 1993). The first step in PCP degradation is
methylation o f PCP to pentachloroanisole (PCA), believed to be a detoxification
mechanism. Recently, Lestan and Lamar (1996) developed a fungal inoculum for the
bioaugmentation o f contaminated soils that successfully introduces white-rot fungi into
soil in alginate pellets. In field tests, up to 90% o f the PCP was degraded using this
process. The use o f white-rot fungi provides several advantages to the degradation of
highly toxic and recalcitrant compounds: (I) they are naturally occurring organisms; (2)
they produce nonspecific extracellular enzymes; (3) they can degrade mixtures of
compounds; (4) they require no prior conditioning to the contaminants; (5) they can be
cultivated on inexpensive growth media; and (6) they can survive under nutrient limiting
conditions.
Despite the promise o f white-rot fungi to degrade environmental pollutants in situ,
the bioavailability o f nonpolar organic chemicals (NOCs) such as PCP may limit the
extent and rate o f degradation. It has been established that both synthetic and naturallyderived surfactants can increase the solubility o f a wide range o f NOCs in soils (Sun et
al., 1995; Sun and Boyd; 1993; Pennell et al., 1993; Kiie and Chiou, 1989; Pennell et al.,
1994; Wang and Brusseau, 1993; Van Dyke et al., 1993; Brusseau et al., 1994).
Solubility enhancement occurs due to partitioning o f the NOC from the soil into
surfactant micelles, a structural arrangement o f surfactant monomers where apolar “tails”
form a nonpolar cavity. The formation o f micelles occurs when surfactant concentrations
exceed the critical micelle concentration (CMC).
.55
The use o f biosurfactants and/or naturally-derived surfactants has received interest
since they are potentially less toxic to microorganisms. Biosurfactants may also serve as
a C source and stimulate biological activity. A recent review o f the influence of
surfactants on the microbial degradation o f organic compounds (Rouse et al., 1994)
concluded that deficiencies in knowledge concerning surfactant-aided degradation
currently limit attempts to successfully incorporate surfactants into remediation activities.
Moreover, the influence o f surfactants on fungal growth and NOC degradation is less
well documented. However, it has been shown that the addition o f surfactants to cultures
of Phanerochaete chrysosporium resulted in increased ligninase production, suggesting
that surfactants may stimulate nonspecific enzymatic degradation (Lestan, et al., 1990;
Ashter and Corrius, 1986).
The objectives o f the current study were to: (I) compare the abilities of
Phanerochaete chrysosporium, a non-Phanerochaete white-rot species, and indigenous
organisms to mineralize PCP in PCP contaminated soil, (2) examine the effect of
surfactant type on fungal degradation o f PCP, (3) examine the influence o f surfactant
concentration on PCP mineralization, and (4) determine the relationship between
chemical and biological effects o f surfactants on fungal degradation o f PCP.
Materials and Methods
Chemicals
14C-Uniformly ring-labeled pentachlorophenol (PCP) was purchased from Sigma
Chemical (St. Louis, MO). Radiopurity was verified by high-pressure liquid.
56
chromatography (HPLC) on an Econosil Cl 8 10 micron reverse phase column (Alltech
Associates, Inc. Deerfield, IL) using acetonitrile-0.2% phosphoric acid (80:20) with a
flow rate o f 1.0 mL min'1 (Rf = 6.6 min) attached to a Beckman 171 Radioisotope
Detector (Beckman Instruments, Inc., Fullerton, CA). Unlabelled PCP was purchased
from Sigma Chemical (St. Louis, MO) and a stock solution o f 10 mg mL"1 prepared in
acetone.
Soil
An uncontaminated agricultural soil, the Amsterdam silt loam (Gallatin Co., MT;
Table 3), was collected from the surface 15 cm, passed through a 2 mm sieve, and stored
Table 3. Characteristics of the Amsterdam Silt Loam (Typic Cryoboroll).
%
Sand
15
%
Silt
52
%
Clay
33
Textural
Classification
silt loam
Organic
C(%)
1.80
CEC
(mmol charge kg'1)
0.21
pH
(1:1)
6.9
Om
(1/3 bar)
0.26
field moist at 4°C until use. Prior to use, soils requiring sterilization were autoclaved for
2 h, let rest 48 h, autoclaved for 2 h, and oven dried at 105°C for 24 h; unsterilized soils
were air-dried for 48 h. The soil was treated with 2 pCi 14C-PCP kg soil"1 and additional
unlabeled PCP to yield a final concentration o f 100 mg kg"1, delivered by spraying a PCPacetone solution uniformly to soil (20 pL g"1). The 14C concentration in the soil was
verified by oxidation o f triplicate soil subsamples at 800°C under an Oz atmosphere for
four min using a Biological Oxidizer, Model 0X 300 (R.J. Harvey Instrument Co.,
Hillsdale, NI), and analyzed using scintillation analysis on a Packard Tri-Carb Liquid
Scintillation Analyzer, Model 2200CA (Meriden, CT).
57
Surfactants
Nine surfactants [lignosulfonic acid, sodium salt; lignosulfonic acid, sodium salt,
acetate; hydroxypropyl-P-cyclodextrin; Tween 20, 40, 60, 80, 85; and Triton X-100
(Figure 16)] were purchased from Sigma Chemical Co. (St. Louis, MO). Each surfactant
was mixed with deionized water to achieve a concentration that would yield the
appropriate surfactant concentration when the soil was brought to field capacity water
content (mass basis, Om= 0.26).
Lignosulfonic acid,
sodium salt, acetate
HO(C2H4O)w^ ____
O
Hydroxypropyl-Pcyclodextrin
Lignosulfonic acid,
sodium salt
.(OC2H4)xOH
CjIH(OC2H4)yOH
(CH3)3CCH2
O(CH2CH2O)nH
CH2(OC2H4)z R
n = 9-10
Tw een R
20
40
60
80
85
C12H24O2
C16H32O2
ClgH36O2
C18H34O2
(ClgH34O2)3
Triton X -100
Figure 16. Structures of the nine surfactants used in the study. The chemical formulas for the Tween
series are given as the R group.
58
Fungi
Phanerochaete chrysosporium and a non-Phanerochaete species o f white-rot
fungi were provided by Mycotech Corp. (Butte, MT) growing on a solid substrate.
Cultures were stored at 4°C until use. Fungal viability and enzyme production were
periodically checked by growing the fungi on malt agar and Poly R-478 indicator agar
(Freitag, et al., 1992). Discolorization o f the indicator medium signaled positive enzyme
production by the cultures.
Batch Biodegradation
Thirty-five g o f PCP-Iabeled soil were mixed with 0.7 g sawdust, 2.1 g solid
nutrient supplement (containing 3.1% N and 0.4% P by mass in a slow release
formulation), 1.7 g fungal inoculum, and brought to field capacity water content (i.e. 1/3
bar) with double-deionized water or a surfactant containing solution; all treatments were
conducted in triplicate. The soil mixture was packed into 2.54 cm diameter by 6 cm
length acrylic columns and constantly flushed with humidified compressed air (Figure
17). PCP mineralization was measured as the accumulation o f 14COz in 10 mL 0.5M
NaOH traps changed every two d for four weeks. At the conclusion o f each experiment,
soils were mixed and subsamples oxidized in triplicate at 800°C under an Oz atmosphere
for four min using a biological oxidizer. Total 14C recoveries generally exceeded 90%.
Soils were solvent-extracted and analyzed for total PCP by HPLC analysis. PCP aging
and extractability in the Amsterdam soil was followed by extraction o f sterile soils over
59
the same time period.
PCP contaminated soil
Sawdust
Nutrient supplement
W hite-rot fungi
Surfactant
V
Hum idified air
4
0.5 M NaOH
• m ineralization as 14CO2
• respiration estim ated
from total CO2
Figure 17. Experimental setup for measuring PCP mineralization in contaminated soil.
PCP Extraction and HPLC Analysis
Soils were solvent-extracted by mixing 5 g soil, 10 g anhydrous Na 2S 0 4 (dried for
2 hat 105°C), one microspatula OfNa2S2O4, 9.75 mL 1:1 acetone:hexane, and 0.25 mL
concentrated H2SO4 in 22 mL glass scintillation vials. Samples were shaken on a table
shaker overnight then filtered through a 0.45 pm nylon filter. One mL subsamples were
dried under N 2, I mL o f 0.5 M NaOH added to the residue, and samples equilibrated by
shaking overnight. Samples were diluted to a total volume o f 5 mL by the addition o f 4
mL o f 80:20 acetonitrile-0.2% phosphoric acid and passed through a 0.2 pm nylon filter
60
before HPLC analysis. PCP was analyzed by UV absorbance at 254 nm after passing
though an Econosil CIS 10 micron reverse phase column using 80:20 acetonitrile-0.2% ■
phosphoric acid at 1.0 mL min'1 (Rf = 6.6 min.) Extraction recoveries o f PCP from the
Amsterdam soil were approximately 90%.
The original solvent extract was analyzed for total remaining 14C by liquid
scintillation analysis. Metabolite production was followed in selected experiments using
HPLC-radioisotope detection (Beckman 171 Radioisotope Detector). An aqueous
extraction o f selected soils was also conducted, followed by total 14C analysis using liquid
scintillation.
PCP Sorption and Solubility
The sorption o f Tween 80 to Amsterdam soil was measured in triplicate using
batch systems containing I g soil and 15 mL o f solution (0, 10, 50, 200, 500, 1000, 2000,
4000, 6000, 8000, 10000, and 20,000 mg Tween 80 L"1) in glass Corex centrifuge tubes.
The mixtures were equilibrated for 24 hrs on a table shaker, centrifuged, and surface
tension measurements made on the solution using a Cenco-du Nouy Interfacial
Tensiometer (No. 70545, Central Scientific Co., Franklin Park, IL). A standard curve
relating surface tension to Tween 80 concentration was used to determine Tween 80
sorption to Amsterdam soil.
The sorption o f PCP to Amsterdam soil was measured in batch systems (triplicate)
using I g o f contaminated soil, containing 10, 50,100, 500, and 1000 mg kg'1 PCP
(including 2 pCi 14C-Iabel kg soil"1), and 15 mLs o f solution containing 0, 100,1000, and
61
10000 mg L'1 Tween 80. The suspensions were equilibrated for 24 h, centrifuged, and I
mL o f supernatant analyzed for total 14C by scintillation analysis. The equilibrium pH
values ranged from 6.8 to 7.1. Sorption o f PCP was determined as the difference between
the total PCP added and the PCP remaining in solution. Kd values were determined by
linear regression using the relationship Q = KjC, where Q is the sorbed concentration (mg
kg"1) and C is the aqueous concentration (mg L"1), through the measured points.
The solubility enhancement o f PCP in the presence o f Tween 80 and Amsterdam
soil was determined in batch systems (triplicate). Each treatment contained 5 g
contaminated soil (100 mg kg"1 PCP) and 10 mL o f surfactant solution, yielding a final
concentration o f 0, 3.6, 7.3, 36.6, and 183.5 mg Tween 80 g soil"1. The equilibrium pH
values ranged from 6.8 to 7.1. PCP solubility was determined by analyzing 14C-PCP in
solution o f I mL aliquots by scintillation analysis.
Fungal Activity
Fungal respiration was estimated by monitoring total CO2 production as a
function of time from soil columns by titrating 7.5 mL o f trap solution for total alkalinity
with standardized HC1. Fungal biomass was estimated in selected soil columns by
measuring the ergosterol content. -(Ergosterol is a component o f cell walls specific to
fungi and has been used as a measure o f fungal growth in solid substrates; Davis and
Lamar, 1992). Five g soil samples were mixed with 20 mL methanol in 50 mL tubes and
mixed for 3 h. After settling, 5 mL o f supernatant was filtered through a 0.45 pm filter
into clean tubes containing 0.3 g KOH. Tubes were capped and incubated at 75°C for 90
62
min and then allowed to cool. Tubes were shaken vigorously after the addition o f IOrnL
water and 5 mL hexane. After clarification, 4 mL o f supernatant was transferred to a
small vial and dried under N 2. Residue was redissolved in I mL o f methanol and
analyzed by HPLC at 282 nm after passing through a C l8 reverse-phase column using
methanol at a flow rate o f 0.75 mL min"1.
Results and Discussion
Effects o f Surfactants on PCP Mineralization
Mineralization o f PCP in the Amsterdam soil by Phanerochaete chrysosporium
was less than 5% after 27 d with all of the surfactants studied (Figure 18). P.
chrysosporium did not colonize the soil effectively and after 10 d, the columns were
visibly colonized by a contaminant fungus, verified by incubation o f soil samples on malt
Fraction 14C -P C P evolved as 14COz
0 .4 0
—t— Sterile C ontrol
0 .3 5
—
Indigenous
N o surfactant
0 .3 0
-" -L S A
0 .2 5
- * - LSA A
-x- C D
0 .2 0
T w een 80
-" - T r it o n X -1 0 0
0 .1 5
0 .1 0
0 .0 5
0 .0 0 i
10
15
20
T im e (days)
Figure 18. Mineralization of PCP in Amsterdam soil by P lia n e ro c lia e le
25 mg surfactant g soil '.
c h r y s o s p o r iu m
in the presence of
63
agar and Poly-R 478 media. Previous laboratory studies with P. chrysosporium have
generally been conducted at elevated temperatures (30°C or greater), well above the 2123°C used in our laboratory experiments or found in typical field conditions, and may
account for the poor response observed in the current study as compared with other
studies. Indigenous soil microorganisms mineralized nearly 10% o f the added PCP over
four weeks and proved to be more effective at PCP mineralization than P. chrysosporium.
The non-Phanerochaete species was much more effective at colonizing soil (as
evident by visual observation o f hyphal growth) and mineralized between 15-36% o f the
added PCP, depending upon surfactant treatment (Figure 19). Incubation o f soil
subsamples from these columns on malt agar and Poly R-478 media at the termination of
0.45
■Sterile C ontrol
S
0.40
3 0.35
■o
a» 0.30
>
E
0.25
■N o surfactant
LSA
LSA A
■CD
■Tw een 80
yCu o.2o
V
■In d igen ou s
• Triton X -1 0 0
0.15
15
20
T im e (days)
Figure 19. Mineralization of PCP in Amsterdam soil by a n o n -P h a n e ro c h a e te species in the
presence of 25 mg surfactant g soil"1.
64
mineralization experiments confirmed that the predominant species growing in these
columns was a white-rot fungus. In the absence o f surfactants, the non-Phanerochaete
species clearly exceeded the ability o f indigenous soil microorganisms to mineralize PCP.
Although the five tested surfactants had no effect on PCP mineralization in soils
inoculated with P. chrysosporium. PCP mineralization by the non-Phanerochaete species
was enhanced by the addition o f 25 mg Tween 80 g soil"1 (25-36%, range for 3
experiments) relative to a no surfactant control (15-23%, range for 3 experiments). The
presence o f other surfactants (25 mg g"1 o f LS A. LS AA, CD. and Triton X -100) either
had no effect or resulted in slight inhibition o f PCP mineralization compared to the no
surfactant control (Figure 19). With the exception o f Tween 80. none o f the surfactants
T im e (days)
Figure 20. Fungal respiration estimated from total CO2 produced from Amsterdam soil columns
treated with a n o n -P h a n e ro c h a e te species and five surfactants.
65
resulted in significantly higher production o f total CO2 (an indicator o f total fungal
activity), nor was there any evidence that these surfactants inhibited fungal growth
(Figure 20). Further, there was no evidence that the naturally derived surfactants (LSA,
LSAA, or CD) yielded higher CO2 production than Triton X-100 or Tween 80 (synthetic
surfactants).
Measurement o f the total PCP (solvent extractable) remaining in soil as a function
of time in the presence and absence o f surfactant did not reveal the identical trend as
observed with evolution o f 14CC^ (Figure 21). Specifically, Tween 80 did not cause as
rapid a decline in extractable PCP as was expected from mineralization o f PCP. judged
from 14CO2 production. Further, solvent extractable PCP decreased to less than 10% of
1.2
N o su rfactant
I
T w e e n 80
M
- o - N o su rfactant
0.8
- 0 - T w e e n 80
0.6
B
2
Cu
U
0.4 t
0
0.2 I
b
0
-
0
5
10
15
20
25
0.2
30
Tim e (days)
Figure 21. Mineralization of PCP (closed symbols) in Amsterdam soil with and without the
addition of 25 mg Tween 80 g so il' by a n o n -P h a n e ro c h a e te species, followed by solvent extraction
of PCP (open symbols) remaining in the soil.
66
the initial PCP concentration in both the presence and absence o f 25 mg Tween 80 g
soil'1. Initially we were concerned that aging may have affected PCP recoveries;
however, separate experiments showed the PCP recoveries remained greater than 95%
after 4 wks in sterile soils. Although it is clear that the disappearance o f total PCP, as
measured using solvent extraction, was indeed due to fungal degradation, extraction
efficiency and extraction reproducibility remain significant concerns. Consequently, we
place more emphasis on the 14CC^ data as a reliable measure o f PCP mineralization.
Finally, no significant quantities o f labeled nonpolar metabolites, including
pentachloroanisole, were detected using radioisotope detection in either treatment (0 vs
25 mg Tween 80 g soil"1) over the course o f the study.
To test the relative effectiveness o f the Tween family o f surfactants for enhancing
PCP mineralization in soil, five Tween surfactants (Tween 20, 40, 60, 80, and 85) were
added to the Amsterdam soil at 25 mg surfactant g soil'1 in the presence o f the nonPhanerochaete species. After four weeks, only the Tween 80 treatment resulted in
greater PCP mineralization compared to the no surfactant control (Figure 22). Tween 80
has been shown to increase ligninase production in P. chrysosporium, and this may
account for the increased mineralization o f PCP observed in the current study (Lestan et
at., 1990; Ashter and Corrieu, 1986).
Potential Mechanisms of
Surfactant Enhanced PCP Mineralization
The addition o f 25 mg Tween 80 g soil"1 to systems containing the nonPhanerochaete white-rot species consistently resulted in enhanced mineralization ofPCP.
67
No surfactant
Tween 20
Tween 40
Tween 60
-♦-T w een 80
- * - Tween 85
Tim e (days)
Figure 22. Mineralization of PCP by a n o n -P h a n e ro c h a e te species in Amsterdam soil amended
with 25 mg Tween 20, 40, 60, 80, and 85 g soil"1.
The potential mechanisms of enhanced mineralization include (I) an increase in the
solubility and subsequent bioavailability o f PCP as a function of surfactant concentration,
and (2) an increase in fungal activity perhaps causing an increase in production of
enzymes responsible for PCP degradation. In order to test these potential mechanisms,
measurements o f PCP mineralization were coupled with measurements o f PCP sorption
and solubility enhancement and estimates o f fungal activity over a range o f Tween 80
concentrations.
Mineralization o f PCP by the non-Phanerochaete species increased with
increasing levels o f Tween 80 up to 25 mg g soil'1 (Figure 23). As mentioned previously,
the addition o f 25 mg Tween 80 g soil'1 resulted in mineralization o f 26% o f the added
68
14C-PCP to 14CO2 compared to only 17% in the absence o f Tween 80. Higher
concentrations o f Tween 80 were not tested due to solubility limitations o f surfactant in
the quantity o f solution required to bring the soil to field capacity.
It is well established that for micelle-forming surfactants, significant solubility
T w een 80 concentration (m g g soil" )
Tim e (days)
Figure 23. Mineralization of PCP by a n o n -P h a n e ro c h a e te species in Amsterdam soil amended
with 0, 0.05, 0.5, I, 5, or 25 mg Tween 80 g soil"1.
enhancement o f NOCs is not observed unless surfactant concentrations exceed the CMC.
Batch experiments designed to measure the solubility of PCP in the presence of
Amsterdam soil showed that solubility enhancement did not occur until Tween 80
concentrations exceeded approximately 40 mg g"1. In fact, PCP solubility was lower
relative to a no surfactant system at total surfactant concentrations below 40 mg g"1
(Figure 24). Sun et al. (1995) observed a similar trend for the solubility characteristics o f
69
DDT and PCB in soil as a function o f added surfactant (Triton X -100) concentration.
They concluded that sorbed surfactant monomers resulted in greater sorption at Triton X100 concentrations below effective CMC values, and further demonstrated that solubility
enhancement was not observed until sufficient surfactant had been added to exceed
maximum sorption, followed by a subsequent drop in surface tension due to the formation
o f surfactant micelles.
E
U
g
s
•o
C
O
s
•35
<u
V
3
CA)
T w een 80 co n c en tr a tio n (m g g soil"1)
Figure 24. Aqueous concentration of PCP and aqueous phase surface tension from Amsterdam soil
equilibrated for 24 h with varying concentrations of Tween 80.
Using surface tension measurements o f Tween 80 solutions in the absence and
presence of soil, the effective Tween 80 concentration required to reach CMC in the
Amsterdam soil was estimated at 37.5 mg g soil"1. The highest Tween 80 concentration
70
used in the mineralization experiments was 25 mg g soil"1, well below the concentration
required to result in increased PCP solubility due to the partitioning o f PCP into
surfactant micelles. PCP sorption isotherms conducted in the presence o f Tween 80
demonstrated a similar response (Figure 25). The partition coefficients, Kd, o f PCP to the
Tween 80 cone, (mg g soil )
c 20
■ 150
y
5
0.0
1.0
2.0
3.0
4.0
S o lu tio n P C P co n c en tr a tio n (m g L"1)
Figure 25. Sorption isotherms of PCP to Amsterdam soil in the presence of Tween 80 at
concentrations of 0,1.5,15, and 150 mg Tween 80 g soil"1. The resulting Kd values are 1.14, 2.05,
14.04, and « 1 L kg'1, respectively.
Amsterdam soil increased at Tween 80 concentrations of 1.5 and 15 mg g soil"1, both
below the CMC, and dramatically decreased at 150 mg Tween 80 g soil'1, a concentration
o f total surfactant which resulted in solution concentrations exceeding the CMC of Tween
80. Both the solubility data and the sorption data for PCP as a function o f Tween 80
concentration support the conclusion that enhanced fungal mineralization in the presence
71
o f Tween 80 was not due to solubility enhancement o f PCP. Aronstein et al. (1991)
reported a similar response for phenanthrene degradation by bacteria in soils treated with
surfactants below the CMC, and suggested that sub-CMC levels o f surfactants increased
the rate o f desorption from soils without increasing contaminant solubility. Although
degradation is still limited by contaminant solubility, the extent o f degradation would
increase if surfactant addition resulted in enhanced desorption rates.
An alternate explanation for the increased PCP mineralization observed in the
presence o f Tween 80 at concentrations below the CMC is the stimulation o f fungal
activity by the surfactant. Macur et al. (in preparation) noted increased hexadecane and
phenanthrene mineralization by indigenous organisms in soils treated with Witconol SN70 at concentrations below the CMC. Respiration by the non-Phanerochaete fungus,
estimated by total CO2 production, was greater after 10 d in soils treated with 25 mg
Tween 80 g soil'1 than in surfactant-free systems (Figure 26). Fungal biomass in Tween
80 treated soils, measured as ergosterol present in the soil, was also significantly higher at
IOd than in the surfactant-free soil (Figure 26). The increased respiration and biomass at
10 d corresponds to the period o f most rapid PCP mineralization. Biomass decreased
during the final 14 d o f incubation, corresponding to decreased PCP mineralization and
lower respiration rates. The increased PCP mineralization observed in the Tween 80
treated soils appears to be related to a stimulation o f fungal growth and activity which
occurred at Tween 80 concentrations below the CMC, and not as a result o f PCP
solubility enhancement. Additional experiments using surfactant concentrations up to
200 mg g"1 soil would be warranted to determine if Tween 80 concentrations necessary to
72
achieve significant solubility enhancement o f PCP would result in further increases in
mineralization rates. However, surfactant additions at such high rates (20%. w/w) would
likely not be economically feasible for the remediation o f contaminated soils.
20
-•-T w een 80
16
No surf
-o - Tween 80
-D -
12
8
4
E rg o stero l (H-g g soil"1)
No surf
0
0
5
10
15
T im e (d a y s)
20
25
30
Figure 26. Total CO2 production from Amsterdam soil in the presence of a n o n -P h a n e ro c h a e te species
and treated with or without 25 mg Tween 80 g soil' (filled symbols). Also shown is the amount of
ergosterol present in the soil over the same time period (open symbols).
Summary
A non-Phanerochaete white-rot species exhibited more aggressive soil
colonization and mineralized PCP more effectively than did indigenous organisms or
Phanerochaete chrysosporium. The rates and extents o f PCP mineralization increased
with increasing concentrations o f Tween 80 (up to 25 mg Tween 80 g soil"1), but were
73
unaffected or depressed by other synthetic or naturally-derived surfactants. At a
concentration o f 25 mg Tween 80 g soil'1, approximately 25% o f PCP was mineralized to
CO2 compared to 15% in the absence o f a surfactant. In both cases, PCP concentrations
decreased to less than 10% o f the original concentration after four weeks. The increased
mineralization o f PCP was not a result o f increased PCP solubility since it was shown
that the addition o f 25 mg Tween 80 g soil'1 was insufficient to achieve aqueous
concentrations greater than the CMC. In fact a 25 mg Tween 80 g soil"1 application rate
actually resulted in a small decrease in PCP solubility. The addition o f Tween 80 resulted
in increased fungal activity, as inferred from increased respiration rates, and increased
fungal biomass, based upon ergosterol measurements. Our results show that the use o f
surfactants such as Tween 80 at sub-CMC levels in combination with white-rot fungi may
represent a strategy for enhancing the remediation o f PCP in contaminated soils. Further,
higher concentrations o f surfactant should be tested to determine if increases in PCP
solubility result in greater mineralization rates, and to determine if high surfactant
concentrations result in inhibition o f fungal growth.
Acknowledgements
The project was supported by the Great Plains-Rocky Mountain Hazardous
Substance Research Center and the MSU Center for Biofilm Engineering, created through
Cooperative Agreement ECD-8907039 between the National Science Foundation and
Montana State University.
74
CHAPTER 4
SUMMARY
The availability o f an organic molecule to microorganisms is o f central
importance in soil bioremediation. In the degradation o f 2,4-D (Chapter 2), sorption to
soil organic matter resulted in lowering the amount o f pesticide available for degradation,
resulting in greater quantities o f 2,4-D persisting in higher organic matter soils over low
organic matter soils. Although a weak relationship was observed between the partition
coefficient and extent o f degradation, it was evident that sorption and thermodynamic
data alone are poor predictors o f biodegradation. Bioavailability is controlled by mass •
transfer processes as well as chemical sorption. A more appropriate approach to
predicting contaminant bioavailability and biodegradation would be through the use o f
mass transfer rates that describe the kinetics o f soil desorption processes.
The introduction o f the surfactant Tween 80 at sub-CMC levels into PCP
contaminated soils resulted in increasing the total extent o f mineralization (Chapter 3).
However, the increase was not a result o f increased PCP solubility since the surfactant
addition actually resulted in a slight decrease in total aqueous concentrations. The
enhanced degradation achieved with surfactant addition resulted from either (I) increased
fungal activity o f (2) increased contaminant mass transfer from the sorbed state.
Regardless o f the exact mechanism involved, the addition o f surfactants at sub-CMC
75
concentrations can result in increased contaminant degradation, suggesting the utility of
surfactants in bioremediation strategies.
Despite the promises that bioremediation offers for cleaning the environment,
some key issues remain unanswered and limit widespread application o f these
technologies. It is clear that microorganisms are the primary agents responsible for
contaminant degradation, however, poor understanding o f microbial communities and
their interactions limit our ability to assess bioremediation potential and shape
remediation schemes. Furthermore, surfactants appear to facilitate greater contaminant
degradation in some instances, but microbial responses are varied, suggesting the need
for greater investigation into bioavailability and biodegradation enhancement through
these chemical agents.
A fuller understanding o f the physical, chemical, and microbiological issues
involved in soil environments would provide useful information for assessment o f
bioremedation potential in contaminated areas. It is encouraging to recognize that as our
knowledge o f the environment continues to grow, we are being equipped with powerful
tools to help clean our world o f our past mistakes. If mankind has learned one lesson
from his legacy o f pollution, it is an increased level o f environmental awareness.
Hopefully, it is this collective wisdom, and not our follies of the past, that will be passed
on to future generations. As we continue to look at relationships among our species and
the planet and strive to understand the processes taking place in soils, we are really laying
the groundwork for a brighter, cleaner future.
76
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85
APPENDICES
APPENDIX A
ADDITIONAL 2,4-D DATA
Fract. 14C-Z,4-D evolved as l4COz
87
200
400
600
800
T im e (hrs)
Figure 27. Degradation of 2,4-D in 31 agricultural soils collected from around Montana. Soils
demonstrated a wide range in both rates and extents of degradation. Degradation curves are
shown without data points for clarity.
88
♦
Observed Data
------First-order r' = 0.7303
-----Modified first-order r" = 0.9585
. of 0.2
----- Logistic r ' = 0.9703
Time (hrs)
Figure 28. Comparison of three kinetic models used to describe 2,4-D degradation in soil.
The points are the observed data; the solid line is the fit from the first order model; the
dashed line is the modified first-order model; and the dotted line is the logistic model. Also
shown are r2 values from non-linear regression analyses. The modified first-order and
logistic equations comparably described the observed degradation curves. First-order
kinetics orovided noor fits.
89
0.045
Crop-Fallow
0.040
y = -0.1051x + 0.8031
R2 = 0.6492
0.035
0.030 e
0.025
I
0.020
§
0.015
S
CC
0.010
♦ Extent of degradation
□ Rate constant
0.005
0.000
Soil organic carbon (%)
0.035
Native Range
0.030
y = -0.0987% + 0.7078
R2 = 0.483
0.025 -T
0.020 I
I
QQ
0.015
Si
o.oio 5
♦ Extent of degradation
□ Rate constant
0.005
0.000
Soil organic carbon (% )
Figure 29. Relationship between maximum extent of 2,4-D degradation and fitted rate
constants (from modified first-order) to soil organic carbon for the crop-fallow and
native range groups of soils. A linear regression between extent of degradation and SOC
is shown for each set of soils. The relationship was stronger for the crop-fallow soils than
for the native-range soils. No apparent relationship was evident between the rate
constant and SOC.
90
Amsterdam
Brocko
Time (hrs)
Anaerobic
Aerobic
Amsterdam
Brocko
O
500
1000
1500
2000
T im e (hrs)
Figure 30. Comparison of aerobic and anaerobic conditions on 14C-Carboxyl-Iabeled 2,4-D
degradation (8^, = 0.46). Top figure is 14CO2 production from 14C-Carboxyl-Iabeled 2,4-D under
aerobic conditions. Bottom figure is 14CO2 production from 14C-Carboxyl-Iabeled 2,4-D
preconditioned under anaerobic conditions then exposed to aerobic conditions. Total extent of
degradation increased in the Amsterdam soil when incubated under anaerobic conditions.
91
APPENDIX B
ADDITIONAL PCP DATA
Table 4. Soil characteristics for three PCP contaminated soils used in PCP degradation studies.
%
Sand
Soil
15
Amsterdam
43
Libby Pole
Montana Pole 79
%
Silt
52
43
13
%
Clay
33
14
9
Textural class
Silt loam
Loam
Loamy sand
Organic
C(%)
1.8
2.0
2.3
CEC
(mmhos
cm'1)
0.21
0.33
0.22
pH
(1:1)
6.9
NA
8.1
Water
content
(1/3 bar)
0.26
0.23
0.12
PCP
concentration
(mg kg'1)
100
445
273
93
No surf
Tween 80
Lecithin
Witconol SN-70
0.0
*
T im e (d ays)
Figure 31. Degradation of PCP in Amsterdam soil by a n o n -P h a n e ro c h a e te white-rot fungi treated
with surfactants at 25 mg g soil"1. The naturally-derived surfactant lecithin and the synthetic
surfactant Witconol SN-70 are compared with no surfactant and Tween 80 treatments for
comparison. Both lecithin and Witconol were less effective than either no surfactant or Tween 80
at promoting PCP degradation.
C um ulative COz evolved (m m ol)
F r a c t 14C -P C P evolved as 14C O z
94
Tim e (days)
Figure 32. Effect of soil mixing on PCP degradation and fungal respiration by a nonspecies. Soil was mixed after 28 days and moisture content readjusted to field
capacity. Mixing resulted in a very slight increase in PCP degradation and fungal respiration.
P h a n e r o c h a e te
95
Indigenous, no surf
Indigenous, Tween 80
F600, Tween 80
T im e (d ays)
Figure 33. Degradation of PCP in the Libby Pole soil by indigenous organisms and a nonspecies (F600), both with and without the addition of 25 mg Tween 80 g soil'1. _None
of the treatments showed considerable PCP mineralization up to 24 d. Increases in total
mineralization after 24 d are accompanied by very large standard errors, indicating non-uniform
variation among replications.
P h a n e r o c h a e te
96
Tween 80 concentration (mg g soil" )
T im e (d ays)
Figure 34. Degradation of PCP in the Montana Pole soil by a n o n -P h a n e ro c h a e te white-rot fungus
and the surfactant Tween 80 at five concentrations. Tween 80 had no effect on the degradation of
PCP, perhaps due to toxicity effects. The Montana Pole soil contains high levels of heavy metals
(e.g. Cu) and other toxic materials such as arsenic and may have interfered with fungal growth.
c
97
Sterile
Indigenous
F600
Penicillium
0
5
10
15
20
25
30
T im e (d ays)
Figure 35. Degradation of PCP in Amsterdam soil by indigenous organisms, a n o n -P h a n e ro c h a e te
white-rot fungus (F600), and a fungus isolated from PCP-contaminated soils (tentatively identified
as a P e n ic illiu m species). The n o n -P h a n e ro c h a e te fungus was more effective at PCP degradation
(17%) than the indigenous soil organisms (11%); the P en icilU m fungus, however, resulted in
greater than 40% degradation of the PCP after four weeks. More research into PCP degradation
by this organism is warranted.
Indigenous
F600
PeniciIlium
Fract.
C -P C P evolved as
COz
98
T im e (d ays)
Figure 36. PCP degradation in Libby Pole soil inoculated with a n o n -P h a n e ro c h a e te white rot
fungus and a P e n ic illiu m fungus isolated from soil. Neither fungus enhanced degradation above
that of the indigenous soil microoganisms in this particular soil.
99
Surface tension o f pure solution
Surface tension in the presence
o f I g Amsterdam soil
30 —
0
5000
10000
15000
20000
T w een 80 co n cen tra tio n (m g L '1)
Figure 37. Surface tensions measurements of Tween 80 solutions at concentrations of 0, 10, 50, 200,
500, 1000, 2000, 4000, 6000, 8000,10000, and 20,000 mg Tween 80 L"1. The diamonds are surface
tension measurements made in the absence of soil; the CMC is estimated at 7.5 mg L"'. The squares
are surface tension measurements made after 24 h equilibration of 15 mL of surfactant solution
with I g of Amsterdam soil. The estimated concentration required to achieve CMC in a soil
containing system is 2500 mg L"1or 2.5 orders of magnitude higher. The impact of surfactant
sorption to soil is well illustrated by this comparison.
3
*